Handbook of Environmental Materials Management [1 ed.] 9783319736440, 9783319736457

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Chaudhery Mustansar Hussain Editor

Handbook of Environmental Materials Management

Handbook of Environmental Materials Management

Chaudhery Mustansar Hussain Editor

Handbook of Environmental Materials Management With 811 Figures and 417 Tables

Editor Chaudhery Mustansar Hussain Department of Chemistry and Environmental Science New Jersey Institute of Technology Newark, NJ, USA

ISBN 978-3-319-73644-0 ISBN 978-3-319-73645-7 (eBook) ISBN 978-3-319-73646-4 (print and electronic bundle) https://doi.org/10.1007/978-3-319-73645-7 Library of Congress Control Number: 2019933557 © Springer Nature Switzerland AG 2019 This work is subject to copyright. All rights are reserved by the Publisher, whether the whole or part of the material is concerned, specifically the rights of translation, reprinting, reuse of illustrations, recitation, broadcasting, reproduction on microfilms or in any other physical way, and transmission or information storage and retrieval, electronic adaptation, computer software, or by similar or dissimilar methodology now known or hereafter developed. The use of general descriptive names, registered names, trademarks, service marks, etc. in this publication does not imply, even in the absence of a specific statement, that such names are exempt from the relevant protective laws and regulations and therefore free for general use. The publisher, the authors, and the editors are safe to assume that the advice and information in this book are believed to be true and accurate at the date of publication. Neither the publisher nor the authors or the editors give a warranty, express or implied, with respect to the material contained herein or for any errors or omissions that may have been made. The publisher remains neutral with regard to jurisdictional claims in published maps and institutional affiliations. This Springer imprint is published by the registered company Springer Nature Switzerland AG The registered company address is: Gewerbestrasse 11, 6330 Cham, Switzerland

I would like to dedicate this handbook to My beloved Holy Prophet (May the blessings and the peace of Allah be upon him (Muhammad)) “Mera Pyarey Aqa: App Sall Allāhu ʿAlay-Hi Wa-Sallam”

Preface

Environmental management is the management of the interaction and impact of human societies on the environment. It offers research and opinions on the use and conservation of natural resources, protection of habitats, and control of hazards. Modern environmental management of materials helps to minimize environmental impacts by reducing the release of toxic substances to the environment and by limiting human exposure. It also helps to reduce pressures on resources by diminishing the quantities of materials that need to be extracted. Beyond this, environmental management of materials supports sustainable decision-making by balancing the social, environmental, and economic considerations throughout the life cycle of a product or material, ensuring that negative impacts are not shifted from the production process to the consumption phase, or vice versa. Modern environmental management of materials therefore encourages the consideration of the impacts of policies that affect a given target area, thereby promoting consideration and possible identification of policy incoherence where this may be the case. Going for green future and establishing a resource-efficient economy is therefore a major environmental challenge today. In this context, putting in place policies that ensure environmental materials management building on the principle of the 3Rs – Reduce, Reuse, Recycle – is crucial. Environmental materials management can help both to improve the environment, by reducing the amount of resources that human economic activity requires as well as diminishing the associated environmental impacts, and to improve resource security and competitiveness. However, till today, the advanced comprehensive understanding and real-world applications of these environmental materials management strategies are still at distance. This handbook summarizes recent progresses and developments for modern environmental materials management at both experimental and theoretical model scales. To capture the comprehensive overview of the modern environmental materials management and to offer reader a systematic and coherent pattern of the topic and focused up-to-date reference, the handbook is divided into several parts, where each Part comprises of several chapters, starting with introduction where modern inventive perspective of environmental materials management is grasped. Part 1 discusses major environmental issues and problems. Part 2 describes various site remediation technologies that help to manage the modern environment. Parts 3, 4, and 5 elaborate several waste management techniques for hazardous, radioactive, and special vii

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Preface

wastes. Part 6 is dedicated to environmental analysis, while Part 7 talks about various environmental treatment technologies. Parts 8, 9, and 10 describe environmental management of biomaterials, mathematical modeling, and management of nanomaterial-based waste. Part 11 defines fate and transport of pollutants in the environment. Parts 12 and 13 explain environmental risk assessment and socioeconomic factors for modern environmental materials management. The last Part clarifies the role of pollution prevention in terms of sustainability and green world. The future of modern environmental materials management is the last content to summarize the book. The aim of this book is to deliver the recent advancements in environmental materials management techniques because of modern social demands. This handbook is intended for a very broad audience working in the fields of advanced environmental management as a part of materials science, environment, green chemistry, sustainability, environmental engineering, environmental sciences, etc. This handbook will be an invaluable reference source for the libraries in universities and industrial institutions, government and independent institutes, individual research groups, and scientists working in the field of environment. Overall, this handbook is planned to be useful for advanced undergraduate, graduate students, researchers, and scientists who are searching for new and advanced technologies for environmental materials management. The editor and contributors are well-known researchers, scientists, and true professionals from the academia and industry. On behalf of Springer, we are very grateful to the authors of all chapters for their outstanding and passionate efforts in the making of this handbook. Special thanks to Miss Anita Lekhwani (Senior Editor), Miss Alexa Steele (Editor, Major Reference Works), Miss Aparajita Basu, and the Springer Reference Editorial Team for their enthusiastic support and help during this project. In the end, all thanks to Springer for publishing the handbook. Chaudhery Mustansar Hussain, Ph.D. The Editor

Contents

Volume 1 Part I 1

2

3

4

5

Major Environmental Issues and Problems

..............

Assessment of Some Aspects of Provisioning Sewerage Systems: A Case Study of Urban Agglomerations in Ganga River Basin . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Saurabh Shukla and Vinod Tare Economic Assessment of Provisioning a Comprehensive Solid Waste Management System: A Case Study of Urban Agglomerations in Ganga River Basin . . . . . . . . . . . . . . . . . . . . . Smriti Gupta, Saurabh Shukla, and Vinod Tare Decentralized Integrated Approach of Water and Wastewater Management in Rural West Bengal . . . . . . . . . . . . . . . . . . . . . . . Pankaj Kumar Roy, Somnath Pal, Arunabha Majumder, Gourab Banerjee, and Asis Mazumdar

1

3

35

65

Major Issues on Beneficial Utilization of Solar Energy in India . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Debanjan Sannigrahi

89

Major Environmental Issues and Problems of South Asia, Particularly Bangladesh . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . G. N. Tanjina Hasnat, Md. Alamgir Kabir, and Md. Akhter Hossain

109

6

Air Pollution . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Saliha Saadet Kalender and Güler Bilen Alkan

149

7

Status of Particulate Matter Pollution in India: A Review . . . . . . Geetanjali Kaushik, Arvind Chel, Satish Patil, and Shivani Chaturvedi

167

8

Global Environmental Issues Alsharifa Hind Mohammad

............................

195

ix

x

Contents

9

10

11

12

13

14

15

16

Fluoride Contamination in Groundwater and the Source Mineral Releasing Fluoride in Groundwater of Indo-Gangetic Alluvium, India . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Abhishek Saxena Identification of the Source Mineral Releasing Arsenic in the Groundwater of the Indo-Gangetic Plain, India . . . . . . . . . . . . . . Pooja Goel Industrial Solid Waste Management in a Developing Country Governorate and the Opportunities for the Application of Cleaner Production Principles . . . . . . . . . . . . . . . . . . . . . . . . . . . Aida O. Al-Batnij, Issam A. Al-Khatib, and Stamatia Kontogianni Indoor Air Pollution Around Industrial Areas and Its Effect: A Case Study in Delhi City . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Jitendra Kumar Nagar, Raj Kumar, J. P. Shrivastava, and Geetanjali Kaushik Scenario of Landfilling in India: Problems, Challenges, and Recommendations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Swati, Indu Shekhar Thakur, Virendra Kumar Vijay, and Pooja Ghosh Environmental Health Problems Due to Air Pollution Exposure: A Case Study of Respiratory and Associated Morbidities Among Traffic Police Personnel in Aurangabad City of Maharashtra . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Suchirai Gaikwad, N. N. Bandela, Geetanjali Kaushik, and Chaudhery Mustansar Hussain

247

285

303

321

337

Microplastics as Contaminant in FreshWater Ecosystem: A Modern Environmental Issue . . . . . . . . . . . . . . . . . . . . . . . . . . Muafia Shafiq, Abdul Qadir, and Chaudhery Mustansar Hussain

355

Impact of Material Chemistry on the Performance characteristics of a Coal Handling Plant . . . . . . . . . . . . . . . . . . . Kumar Harshit, Syed Ali Hussain Jafri, and Pallav Gupta

379

Part II 17

209

Environmental Site Remediation . . . . . . . . . . . . . . . . . . . . .

Technologies for Treatment of Colored Wastewater from Different Industries . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Shraddha Khamparia and Dipika Jaspal

415

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Contents

18

19

20

21

xi

Application of Novel Microbial Consortia for Environmental Site Remediation and Hazardous Waste Management Toward Low- and High-Density Polyethylene and Prioritizing the Cost-Effective, Eco-friendly, and Sustainable Biotechnological Intervention . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sinosh Skariyachan, Meghna Manjunath, Apoorva Shankar, Nikhil Bachappanavar, and Amulya A. Patil Micro-remediation of Metals: A New Frontier in Bioremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A. Banerjee, M. K. Jhariya, D. K. Yadav, and A. Raj Biostimulation and Bioaugmentation: An Alternative Strategy for Bioremediation of Ground Water Contaminated Mixed Landfill Leachate and Sea Water in Low Income ASEAN Countries . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Jubhar C. Mangimbulude and Ronald Kondo Lembang Bacterial Cell-Mineral Interface, Its Impacts on Biofilm Formation and Bioremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . Hamid M. Pouran

431

479

515

535

22

Bioremediation of Mined Waste Land . . . . . . . . . . . . . . . . . . . . . Nisha Rani, Hardeep Rai Sharma, Anubha Kaushik, and Anand Sagar

557

23

Soil Pollution and Remediation . . . . . . . . . . . . . . . . . . . . . . . . . . . Nataša Stojić, Snežana Štrbac, and Dunja Prokić

583

Part III 24

25

26

Hazardous Waste Management . . . . . . . . . . . . . . . . . . . . . .

Development of an Environmentally Sustainable Approach for Safe Disposal of Arsenic-Rich Sludge . . . . . . . . . . . . . . . . . . . Pankaj Kumar Roy, Arunabha Majumder, Somnath Pal, Gourab Banerjee, Malabika Biswas Roy, Jayanta Debbarma, and Asis Mazumdar

617

619

Household Hazardous Waste Quantification, Characterization, and Management in Developing Countries’ Cities: A Case Study . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Waseem M. W. Al-Tamimi, Issam A. Al-Khatib, and Stamatia Kontogianni

635

Investigation of the Chemical Content of Two Specific Streams in Municipal Waste: The Case of Hazardous Household Waste and Dental Waste . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Stamatia Kontogianni, Nicolas Moussiopoulos, and Issam A. Al-Khatib

657

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27

28

29

Contents

Management of Hazardous Paper Mill Wastes for Sustainable Agriculture . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A. K. Sannigrahi Management of Municipal Solid Waste in Morocco: The Size Effect in the Distribution of Combustible Components and Evaluation of the Fuel Fractions . . . . . . . . . . . . . . . . . . . . . . . . . . A. Ouigmane, Otmane Boudouch, Aziz Hasib, and M. Berkani Hazardous Waste Management with Special Reference to Biological Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Soumya Nair and Jayanthi Abraham

Part IV

Management of Radioactive Wastes . . . . . . . . . . . . . . . . . .

30

Management of Radioactive Wastes . . . . . . . . . . . . . . . . . . . . . . . Bouchra El Hilal, Mohammed Hussein Rafeq Khudhair, and Ahmed El Harfi

31

Actinide Speciation in Environment and Their Separation Using Functionalized Nanomaterials and Nanocomposites . . . . . N. Priyadarshini, K. Benadict Rakesh, and P. Ilaiyaraja

Part V Management of Special Wastes: CO2, CH4, NOX, SO2, Carbon Particles, and Oil Spills . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 32

33

34

35

679

701

715

741 743

771

819

Prospects and Issues of Phosphorus Recovery as Struvite from Waste Streams . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sampriti Kataki and D. C. Baruah

821

Global Status of Nitrate Contamination in Groundwater: Its Occurrence, Health Impacts, and Mitigation Measures . . . . . Saurabh Shukla and Abhishek Saxena

869

Remediation of Industrial and Automobile Exhausts for Environmental Management . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sukanchan Palit and Chaudhery Mustansar Hussain

889

Management of Residues from Air Pollution Control Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Christof Lanzerstorfer

907

Contents

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Volume 2 Part VI 36

37

38

Environmental Analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . .

Tailor-Made Molecular Traps for the Treatment of Environmental Samples . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Rüstem Keçili, Özlem Biçen Ünlüer, and Chaudhery Mustansar Hussain Removal of Pharmaceutically Active Compounds from Contaminated Water and Wastewater Using Biochar as Low-Cost Adsorbents, An Overview . . . . . . . . . . . . . . . . . . . . . . . Adel Al-Gheethi, Efaq Ali Noman, Radin Mohamed, Mohd Adib Mohammad Razi, and M. K. Amir Hashim Microbial Risk Associated with Application of Biosolids in Agriculture . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Adel Al-Gheethi, Efaq Ali Noman, Radin Mohamed, Abd. Halid Abdullah, and M. K. Amir Hashim

927

929

951

961

39

Monitoring and Risk Analysis of PAHs in the Environment . . . . Karishma Hussain, Raza R. Hoque, Srinivasan Balachandran, Subhash Medhi, Mohammad Ghaznavi Idris, Mirzanur Rahman, and Farhaz Liaquat Hussain

973

40

Water Quality Assessment of an Unexplored Tropical Freshwater System in Thiruvananthapuram, India: A Multivariate Statistical Approach . . . . . . . . . . . . . . . . . . . . . . . 1009 Anila P. Ajayan, Jan W. Rijstenbil, and K. G. Ajit Kumar

41

Nanomembranes for Environment . . . . . . . . . . . . . . . . . . . . . . . . 1033 Sukanchan Palit and Chaudhery Mustansar Hussain

42

Exergy and Life Cycle-Based Analysis . . . . . . . . . . . . . . . . . . . . . 1057 Niloufar Salehi, Morteza Mahmoudi, Alireza Bazargan, and Gordon McKay

43

Magnetic and Nanostructural Properties of Cobalt–Zinc Ferrite for Environmental Sensors . . . . . . . . . . . . . . . . . . . . . . . . 1079 A.-H. El Foulani, R. C. Pullar, M. Amjoud, K. Ouzaouit, and A. Aamouche

44

New Trends in Environmental Analysis . . . . . . . . . . . . . . . . . . . . 1097 Awad Ageel Al-rashdi

45

Integrated Evaluation of Quantitative Factors Related to the Environmental Quality Scenario . . . . . . . . . . . . . . . . . . . . . . . . . . 1117 Gustavo Marques da Costa, Annette Droste, Darlan Daniel Alves, and Daniela Montanari Migliavacca Osório

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Contents

46

Characterizing the Cell Surface Properties of Hydrocarbon-Degrading Bacterial Strains, a Case Study . . . . . . 1139 Hamid M. Pouran, Steve A. Banwart, and Maria Romero-Gonzalez

47

Preparation, Characterization, and Heavy Metal Ion Adsorption Property of APTES-Modified Kaolin: Comparative Study with Original Clay . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1167 Bahia Meroufel and Mohamed Amine Zenasni

48

Concept Note on Method Development for Speciation and Measurement of Arsenic (As) in Its Valence States (As (III) and As (V)) in Solid and Semisolid Organic Environmental Samples . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1191 A. M. M. Maruf Hossain

49

Determination of Select Heavy Metals in Air Samples from Aurangabad City . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1199 Onkar Jogdand, N. N. Bandela, Geetanjali Kaushik, and Arvind Chel

50

Recent Advances in Membrane Extraction Techniques for Environmental Samples Analysis . . . . . . . . . . . . . . . . . . . . . . . . . 1209 Hadi Tabani, Saeed Nojavan, Kamal Khodaei, and Alireza Bazargan

51

Environmental Toxicology and Air Pollution: A Comparative Analysis of Different Methods and Studies . . . . . . . . . . . . . . . . . . 1243 Gustavo Marques da Costa, Larissa Meincke, Darlan Daniel Alves, Ane Katiussa Siqueira Frohlich, Sandra Manoela Dias Macedo, and Daniela Montanari Migliavacca Osório

52

The Effects of Atmospheric Pollution in Respiratory Health . . . . 1271 Sandra Magali Heberle, Gustavo Marques da Costa, Nelson Barros, and Michele S. G. Rosa

Part VII

Environmental Treatment Technologies . . . . . . . . . . . . . .

1287

53

Advanced Treatment Technologies . . . . . . . . . . . . . . . . . . . . . . . . 1289 Manviri Rani and Uma Shanker

54

Treatment of Domestic Gray Water by Multicomponent Filters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1341 R. M. S. Radin Mohamed, Adel Al-Gheethi, Muhammad Shabery Sainudin, and M. K. Amir Hashim

55

Development In-House: A Trap Method for Pretreatment of Fat, Oil, and Grease in Kitchen Wastewater . . . . . . . . . . . . . . . . . . . . 1351 R. M. S. Radin Mohamed, Adel Al-Gheethi, A. N. Welfrad, and M. K. Amir Hashim

Contents

xv

56

Environmental Treatment Technologies: Adsorption . . . . . . . . . . 1367 Subramanyam Busetty

57

Making Artificial Beachrock Through Bio-cementation: A Novel Technology to Inhibition of Coastal Erosion . . . . . . . . . . 1399 Md. Nakibul Hasan Khan and Satoru Kawasaki

58

Potential of Biogas Technology in Achieving the Sustainable Developmental Goals: A Review Through Case Study in Rural South Africa . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1423 T. E. Rasimphi, D. Tinarwo, and W. M. Gitari

59

A Review on Treatment of Pharmaceuticals and Personal Care Products (PPCPs) in Water and Wastewater . . . . . . . . . . . . 1433 Mukesh Goel and Ashutosh Das

60

Constructed Wetland: A Green Approach to Handle Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1445 Ashutosh Das and Mukesh Goel

61

How to Improve Selectivity of a Material for Adsorptive Separation Applications . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1469 Vipin K. Saini and Aparajita Shankar

62

Advanced Pretreatment Strategies for Bioenergy Production from Biomass and Biowaste . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1507 C. Veluchamy, Ajay S. Kalamdhad, and Brandon H. Gilroyed

63

Green Infrastructure: Cost-Effective Nature-Based Solutions for Safeguarding the Environment and Protecting Human Health and Well-Being . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1525 Daniel Jato-Espino, Luis A. Sañudo-Fontaneda, and Valerio C. Andrés-Valeri

64

The Modified Bardenpho Process . . . . . . . . . . . . . . . . . . . . . . . . . 1551 Ehsan Banayan Esfahani, Fatemeh Asadi Zeidabadi, Alireza Bazargan, and Gordon McKay

65

Diagnostic and Treatment by Different Techniques of Leachates from Municipal Solid Waste in Morocco Using Experimental Design Methodology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1593 Meriem Abouri, Salah Souabi, and M. Abdellah Bahlaoui

66

Application of Liquid Chromatography-Mass Spectrometry for the Analysis of Endocrine Disrupting Chemical Transformation Products in Advanced Oxidation Processes and Their Reaction Mechanisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1633 Jin-Chung Sin, Sze-Mun Lam, and Abdul Rahman Mohamed

xvi

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67

Olive Mill Wastewater: Treatment and Valorization Technologies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1659 Yahia Rharrabti and Mohamed EI Yamani

68

Studies on Water Quality of Mokokchung District, Nagaland, India, and Removal of Trace Elements Using Activated Carbon Prepared from Locally Available Bio-waste . . . . . . . . . . . . . . . . . 1687 Daniel Kibami

69

Advancement of Photocatalytic Water Treatment Technology for Environmental Control . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1719 Sze-Mun Lam, Jin-Chung Sin, and Abdul Rahman Mohamed

70

Techniques for Remediation of Paper and Pulp Mill Effluents: Processes and Constraints . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1747 Smita Chaudhry and Rashmi Paliwal

71

Biofilm-Based Systems for Industrial Wastewater Treatment . . . 1767 Meryem Asri, Soumya Elabed, Saad Ibnsouda Koraichi, and Naïma El Ghachtouli

72

Role of Bioremediation as a Low-Cost Adsorbent for Excessive Fluoride Removal in Groundwater . . . . . . . . . . . . . . . . 1789 Abhishek Saxena and Anju Patel

73

Biotechnological Approach for Mitigation Studies of Effluents of Automobile Industries . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1821 N. N. Bandela, P. N. Puniya, and Geetanjali Kaushik

74

New Techniques for Treatment and Recovery of Valuable Products from Olive Mill Wastewater . . . . . . . . . . . . . . . . . . . . . 1839 Reda Elkacmi and Mounir Bennajah

75

Role of Earthworms in Managing Soil Contamination . . . . . . . . 1859 Payal Garg, Geetanjali Kaushik, Jitendra Kumar Nagar, and Poonam Singhal

76

The Application of Membrane Bioreactors (MBR) for the Removal of Organic Matter, Nutrients, and Heavy Metals from Landfill Leachate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1879 Susan Hayeri Yazdi, Ali Vosoogh, and Alireza Bazargan

Volume 3 Part VIII 77

Environmental Management of Biomaterials

.........

1899

Sustainable Biomedical Waste Management . . . . . . . . . . . . . . . . 1901 Sukanchan Palit and Chaudhery Mustansar Hussain

Contents

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xvii

Agrowaste Materials as Composites for Biomedical Engineering . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1925 Geetanjali Kaushik, Poonam Singhal, and Arvind Chel

Part IX Environmental Modeling (Mathematical Modeling and Environmental Problems) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

1941

79

A New Method to Estimate the Instantaneous NOx Emissions from Road Traffic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1943 Hicham Gourgue

80

Water Balance Models in Environmental Modeling . . . . . . . . . . . 1961 Khodayar Abdollahi, Alireza Bazargan, and Gordon McKay

81

Modeling the Feasibility of Employing Solar Energy for Water Distillation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1977 Hisham A. Maddah

82

Applications of Soft Computing Methods in Environmental Engineering . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2001 Kaan Yetilmezsoy

83

Modelization of Trihalomethanes Formation in Drinking Water Distribution Systems in France . . . . . . . . . . . . . . . . . . . . . . . . . . . 2047 Otmane Boudouch, C. Galey, C. Rosin, and A. Zeghnoun

Part X Environmental Nanotechnology: Management of Nano-waste (Nanomaterials) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

2061

84

Nano-wastes and the Environment: Potential Challenges and Opportunities of Nano-waste Management Paradigm for Greener Nanotechnologies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2063 Sherif A. Younis, Esraa M. El-Fawal, and Philippe Serp

85

Homeopathic Nanomedicines and Their Effect on the Environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2135 P. Nandy, P. Bandyopadhyay, M. Chakraborty, A. Dey, D. Bera, B. K. Paul, S. Kar, A. Gayen, R. Basu, S. Das, D. S. Bhar, S. Manna, R. K. Manchanda, A. K. Khurana, and D. Nayak

86

Environmental Nanotechnology . . . . . . . . . . . . . . . . . . . . . . . . . . 2159 Junaid Saleem, Usman Bin Shahid, and Gordon McKay

87

Engineered Nanomaterials in the Environment, Their Potential Fate and Behavior and Emerging Techniques to Measure Them . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2191 Hamid M. Pouran

xviii

Contents

88

Environmental Nanotechnology and Education for Sustainability: Recent Progress and Perspective . . . . . . . . . . . . . . . . . . . . . . . . . . 2205 Abdelaziz El Moussaouy

89

Nanotechnology Interaction with Environment . . . . . . . . . . . . . . 2233 Rigers Bakiu

90

Nanowaste Classification, Management, and Legislative Framework . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2257 Rigers Bakiu

91

Current Status and Perspectives in Nanowaste Management Astrid Campos and Israel López

Part XI

Fate and Transport of Pollutants-LCA

. . . 2287

................

2315

92

Transport and Fate of Mercury (Hg) in the Environment: Need for Continuous Monitoring . . . . . . . . . . . . . . . . . . . . . . . . . 2317 Zia Mahmood Siddiqi

93

Life-Cycle Assessment of Construction Materials: Analysis of Environmental Impacts and Recommendations of Eco-efficient Management Practices . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2337 Leonor Patricia Güereca, Daniel Jato-Espino, and Esther Lizasoain-Arteaga

94

Dispersion, Photochemical Transformation, and Bioaccumulation of Pollutants in the Vicinity of Highway . . . . . . . . . . . . . . . . . . . . 2373 Gennady Gerasimov

Part XII

Environmental Risk Assessment . . . . . . . . . . . . . . . . . . . . .

2395

95

Metabolic Toxicity and Alteration of Cellular Bioenergetics by Hexavalent Chromium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2397 Sudipta Pal and Kanu Shil

96

Recent Advances in Toxicology of Gold Nanoparticles . . . . . . . . 2425 Siva Prasad Bitragunta, S. Aarathi Menon, and P. Sankar Ganesh

97

Performance Evaluation of Global Environmental Impact Assessment Methods Through a Comparative Analysis of Legislative and Regulatory Provisions . . . . . . . . . . . . . . . . . . . . . 2441 Lekha Sridhar and Vaibhav Gupta

98

Environmental Impact of Steel Industry Andrea Di Schino

. . . . . . . . . . . . . . . . . . . 2463

Contents

xix

99

Spatial Variation in the Grain Size Characteristics of Sediments in Ramganga River, Ganga Basin, India . . . . . . . . . . . . . . . . . . . 2485 Mohd Yawar Ali Khan

100

Integrated Assessment of Environmental Factors: Risks to Human Health . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2497 Gustavo Marques da Costa and Michele dos Santos Gomes da Rosa

Volume 4 Part XIII

Environmental Policy, Laws, and Economics . . . . . . . . . .

2513

101

Implications of International Environmental Laws: A Close Enquiry at the International Levels to Protect the Environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2515 Noor Mohammad

102

The Role of Local Government Laws in Bangladesh for Promoting Environmental Justice in the Union Parishads of Bangladesh: A Case Study . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2537 Noor Mohammad

103

Research Prioritization in Aerosol Geo-Engineering . . . . . . . . . . 2559 A. M. M. Maruf Hossain

104

Environmental Law and Policy in the Russian Federation: An Overview . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2569 Ekaterina A. Belokrylova

105

Critical Assessment of Existing Environmental Legislation and Policies in India, Its Benefits, Limitation, and Enforcement . . . . 2591 Anirban Dhulia and Rajiv Ganguly

Part XIV Pollution Prevention, Sustainability, and Green World . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

2603

106

A Safety Promotion Approach for Handling of Chemicals in the Working Environmental Management . . . . . . . . . . . . . . . . . . 2605 A. Umamaheswari, S. Lakshmana Prabu, M. Rengasamy, and G. Venkatesan

107

The Importance and Potential of Duckweeds as a Model and Crop Plant for Biomass-Based Applications and Beyond . . . . . . 2629 Hieu X. Cao, Paul Fourounjian, and Wenqin Wang

xx

Contents

108

Enhancing Resilience of Vulnerable Coastal Areas and Communities: Mangrove Rehabilitation/Restoration Works in the Gambia . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2647 Jean-Hude E. Moudingo, Gordon Ajonina, Diyouke M. Eugene, Ansumana K. Jarju, Kwasu Jammeh, Foday Conteh, Saul Taal, Lamin Mai Touray, Modou Njei, and Saiko Janko

109

Assessment of the Biogas Potential in the Vhembe District of Limpopo: A Case Study of Waste-to-Energy Conversion Technology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2693 T. E. Rasimphi, D. Tinarwo, and W. M. Gitari

110

Biosynthesis and Assemblage of Extracellular Cellulose by Bacteria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2703 Sumathi Suresh

111

Wastewater Management to Environmental Materials Management . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2745 Musarrat Parween and A. L. Ramanathan

112

The Evolution of the Paradigm of Pollution Prevention and Sustainability . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2769 W. I. Jose

113

Bioremediation of Hormones from Waste Water . . . . . . . . . . . . . 2801 Anushree Suresh and Jayanthi Abraham

114

An Introduction to Sustainable Materials Management . . . . . . . . 2831 Mohaddeseh Khorasanizadeh, Alireza Bazargan, and Gordon McKay

115

Air Pollution Prevention Technologies Saliha Saadet Kalender

116

Camelina sativa: An Emerging Biofuel Crop . . . . . . . . . . . . . . . . 2889 Shivani Chaturvedi, Amrik Bhattacharya, Sunil Kumar Khare, and Geetanjali Kaushik

117

Sustainability of Waste Glass Powder and Clay Brick Powder as Cement Substitute in Green Concrete . . . . . . . . . . . . . . . . . . . 2927 O. M. Olofinnade, A. N. Ede, and C. A. Booth

118

Dynamic Agro-economic Modeling for Sustainable Water Resources Management in Arid and Semi-arid Areas . . . . . . . . . 2949 Hayat Lionboui, Tarik Benabdelouahab, Aziz Hasib, Fouad Elame, and Abdelali Boulli

119

Photovoltaic Systems and Equipments for the Rural and Urban World . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2975 Mustapha Melhaoui, Kamal Hirech, Ilias Atmane, and Khalil Kassmi

. . . . . . . . . . . . . . . . . . . . . 2871

Contents

xxi

120

Seed Germination and Propagation for Regeneration of Some Medicinal Plants Growing Wild in Semiarid Region of Algeria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3007 Souad Mehalaine

121

Impact of Climate Change and Land Use Change Scenarios on Water Resources in Tha Chin River Basin: A Case Study of Suphan Buri Province, Thailand . . . . . . . . . . . . . . . . . . . . . . . . . 3031 Sathaporn Monprapussorn

122

An Evaluation of Usage of Methyl Esters of Jatropha and Fish Oil for Environmental Protection . . . . . . . . . . . . . . . . . . . . . . . . . 3041 S. Sendilvelan and K. Bhaskar

123

Modern Environmental Materials, Pollution Prevention, Sustainability, and Green World . . . . . . . . . . . . . . . . . . . . . . . . . . 3069 Mohammed Hussein Rafeq Khudhair, Bouchra El Hilal, M. S. Elyoubi, and Ahmed Elharfi

124

Stabilization of Black Cotton Soil Using Waste Glass . . . . . . . . . 3099 Niraj Singh Parihar, Vijay Kumar Garlapati, and Rajiv Ganguly

125

Air Quality Status and Management in Tier II and III Indian Cities: A Case Study of Aurangabad City, Maharashtra . . . . . . . 3115 Geetanjali Kaushik, Satish Patil, and Arvind Chel

126

Modern Air Pollution Prevention Strategies in the Urban Environment: A Case Study of Delhi City . . . . . . . . . . . . . . . . . . 3137 Richa Dave Nagar and Geetanjali Kaushik

127

Environmental Management and Sustainable Development: A Vision for the Future . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3163 Sukanchan Palit and Chaudhery Mustansar Hussain

128

Modern Social Media in Environmental Management and Sustainability . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3181 Poonam Singhal, Indranil Bose, Geetanjali Kaushik, and Chaudhery Mustansar Hussain

Index . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3203

About the Editor

Dr. Chaudhery Mustansar Hussain (born in 1975, in Pakistan, has lived in the USA from 2009, and naturalized in the USA recently) is currently an Adjunct Professor, an Academic Advisor, and Director of Laboratories in the Department of Chemistry and Environmental Science at the New Jersey Institute of Technology (NJIT), Newark, New Jersey, USA. Dr. Hussain is an experienced Environmental Consultant who worked for several projects sponsored by the Asian Development Bank. Dr. Hussain completed his Ph.D. in International Program in Environmental Management from Chulalongkorn University, Bangkok, Thailand. He is truly an educator with several years of experiences and has privileged to present the knowledge, awareness, and familiarity of environmental management from industry to higher education research. His editorial experience includes Editor of the Journal of Chemistry (Hindawi Publishing), Guest Editor of ASME Journal of Nanotechnology in Engineering and Medicine, Editor of BAOJ Chemistry, and many more. Dr. Hussain is the Author of numerous papers in peer-reviewed international journals as well as prolific Author and Editor of several scientific monographs and handbooks published with Springer, Elsevier, the Royal Society of Chemistry, John Wiley & Sons, CRC Press, etc. To date, he has authored/edited more than 20 books and handbooks and contributed toward more than 35 book chapters. His research interests include environmental management, environmental analysis, analytical chemistry, environmental engineering, application of nanotechnology and advanced materials for various industries, etc. Dr. Hussain can speak several languages including English, Urdu, Punjabi, and Thai. xxiii

Contributors

A. Aamouche Applied Sciences National School (ENSA), Cadi Ayyad University, Marrakech, Morocco S. Aarathi Menon Department of Biological Sciences, BITS Pilani, Hyderabad Campus, Hyderabad, Telangana, India Khodayar Abdollahi Department of Engineering Measures for Nature Development, Faculty of Natural Resources and Earth Sciences, Shahrekord University, Shahrekord City, Iran Abd. Halid Abdullah Micro-pollution Research Centre (MPRC), Department of Water and Environmental Engineering, Faculty of Civil and Environmental Engineering, Universiti Tun Hussein Onn Malaysia, Batu, Pahat, Malaysia Meriem Abouri Laboratory of Process Engineering and Environment, Faculty of Sciences and Techniques, Mohammedia, University – Hassan II, Mohammedia, Morocco Jayanthi Abraham Microbial Biotechnology Laboratory, School of Biosciences and Technology, VIT University, Vellore, Tamil Nadu, India Anila P. Ajayan Environmental Biology Division, Department of Botany, Mahatma Gandhi College, University of Kerala, Thiruvananthapuram, Kerala, India K. G. Ajit Kumar Kerala State Biodiversity Board, Thiruvananthapuram, Kerala, India Gordon Ajonina Cameroon Wildlife Conservation Society (CWCS), CWCS Coastal, Forests and Mangrove Programme, Mouanko, Littoral Region, Cameroon Aida O. Al-Batnij Faculty of Graduate Studies, Birzeit University, Birzeit, Palestine Adel Al-Gheethi Micro-Pollutant Research Centre (MPRC), Department of Water and Environment Engineering, Faculty of Civil and Environmental Engineering, University Tun Hussein Onn Malaysia (UTHM), Parit Raja, Johor, Malaysia xxv

xxvi

Contributors

Güler Bilen Alkan Barbaros Hayrettin Naval Architecture and Maritime Faculty, Iskenderun Technical University, Hatay, Iskenderun, Turkey Issam A. Al-Khatib Institute of Environmental and Water Studies, Birzeit University, Birzeit, West Bank, Palestine Awad Ageel Al-rashdi Department of Chemistry, Center for Scientific Research (QCSR), Al-Qunfudah University College, Umm Al-Qura University, Makkah Al-Mukarramah, Al-Qunfudah, Saudi Arabia Waseem M. W. Al-Tamimi Faculty of Graduate Studies, Birzeit University, Birzeit, West Bank, Palestine Faculty of Graduate Studies, Birzeit University, Hebron, Palestine Darlan Daniel Alves Programa de pós-graduação em Qualidade Ambiental, Universidade Feevale, Novo Hamburgo, RS, Brazil M. K. Amir Hashim Micro-Pollutant Research Centre (MPRC), Department of Water and Environment Engineering, Faculty of Civil and Environmental Engineering, University Tun Hussein Onn Malaysia (UTHM), Parit Raja, Johor, Malaysia M. Amjoud Laboratory of Condensed Matter and Nanostructures, Faculty of Science and Technology, Cadi Ayyad University, Marrakech, Morocco Valerio C. Andrés-Valeri GITECO Research Group, University of Cantabria, Santander, Cantabria, Spain Fatemeh Asadi Zeidabadi Department of Chemical and Petroleum Engineering, Sharif University of Technology, Tehran, Iran Meryem Asri Laboratoire de Biotechnologie Microbienne, Faculté des Sciences et Techniques, Université Sidi Mohamed Ben Abdellah, Fes, Morocco Ilias Atmane Faculty of Science, Department of Physics, Team: Materials, Electronics and Renewable Energies MERE, Laboratory Electromagnetism Signal Processing and Renewable Energy LESPRE, Mohamed Premier University, Oujda, Morocco Nikhil Bachappanavar R & D Centre, Department of Biotechnology Engineering, Dayananda Sagar Institutions, Bengaluru, India Visvesvaraya Technological University, Belagavi, Karnataka, India Dayananda Sagar Institutions, Bengaluru, Karnataka, India M. Abdellah Bahlaoui Laboratory of Materials, Catalysis and Development of Natural Resources, Faculty of Sciences and Techniques, Mohammedia, University – Hassan II, Mohammedia, Morocco Rigers Bakiu Faculty of Agriculture and Environment, Agricultural University of Tirana, Tirana, Albania

Contributors

Srinivasan Balachandran Department of Environmental BharatiSantiniketan, Santiniketan, West Bengal, India

xxvii

Studies,

Visva-

Ehsan Banayan Esfahani Department of Chemical and Petroleum Engineering, Sharif University of Technology, Tehran, Iran N. N. Bandela Department of Environmental Sciences, Dr. Babasaheb Ambedkar Marathwada University, Aurangabad, Maharashtra, India P. Bandyopadhyay Centre for Interdisciplinary Research and Education, Kolkata, India A. Banerjee Department of Environmental Science, Sarguja University, Ambikapur, Chattisgarh, India Gourab Banerjee School of Water Resources Engineering, Jadavpur University, Kolkata, West Bengal, India Steve A. Banwart School of Earth and Environment, University of Leeds, Leeds, UK Nelson Barros Energy, Environment and Health Research Unit, FP-ENAS - UFP: University Fernando Pessoa, Porto, Portugal D. C. Baruah Energy Conservation Laboratory, Department of Energy, Tezpur University, Tezpur, Assam, India Department of Mechanical and Industrial Engineering, University of South Africa, Pretoria, South Africa R. Basu Centre for Interdisciplinary Research and Education, Kolkata, India Alireza Bazargan Department of Civil Engineering, K. N. Toosi University of Technology, Tehran, Iran Ekaterina A. Belokrylova Environmental Law and Policy Department, Science of Institute of Law, The Udmurt State University, Izhevsk, Russia Tarik Benabdelouahab National Institute of Agricultural Research, Research Center of Tadla, Tadla, Morocco K. Benadict Rakesh Department of Physics, Indian Institute of Technology Madras, Chennai, Tamil Nadu, India Mounir Bennajah Department of Process Engineering, National School of Mineral Industries of Rabat, BP, Rabat, Morocco D. Bera Centre for Interdisciplinary Research and Education, Kolkata, India M. Berkani Laboratory of Spectro-Chemistry Applied and Environment, University Sultan MoulaySlimane, BeniMellal, Morocco D. S. Bhar Centre for Interdisciplinary Research and Education, Kolkata, India

xxviii

Contributors

K. Bhaskar Department Automobile Engineering, Rajalakshmi Engineering College, Chennai, India Amrik Bhattacharya Department of Chemistry, Indian Institute of Technology, Delhi, India Özlem Biçen Ünlüer Faculty of Science, Department of Chemistry, Anadolu University, Eskişehir, Turkey Siva Prasad Bitragunta Department of Biological Sciences, BITS Pilani, Hyderabad Campus, Hyderabad, Telangana, India Biotechnology Division, Environment Protection Training and Research Institute, Hyderabad, Telangana, India C. A. Booth University of the West of England, Bristol, UK Indranil Bose Indian Institute of Management, Calcutta, India Otmane Boudouch Transdisciplinary Team of Analytical Sciences for Sustainable Development, University Sultan MoulaySlimane, BeniMellal, Morocco Environmental and Agro-Industries Processes Team, University Sultan Moulay Slimane, Beni Mellal, Morocco Abdelali Boulli Laboratory of Environment and Valorization of Agro-Resources, Sultan Moulay Slimane University, Beni Mellal, Morocco Subramanyam Busetty SASTRA Deemed University, Thanjavur, India Astrid Campos Universidad Autónoma de Nuevo León, UANL, Facultad de Ciencias Químicas, Laboratorio de Materiales I, San Nicolás de los Garza, Nuevo León, Mexico Hieu X. Cao Institute of Biology/Plant Physiology, Martin-Luther-University of Halle-Wittenberg, Halle, Saale, Germany M. Chakraborty Centre for Interdisciplinary Research and Education, Kolkata, India Shivani Chaturvedi Department of Chemistry, Indian Institute of Technology Delhi, New Delhi, India Smita Chaudhry Institute of Environmental Studies, Kurukshetra University, Kurukshetra, India Arvind Chel MGM’s Jawaharlal Nehru Engineering College, Mahatma Gandhi Mission, Aurangabad, Maharashtra, India Foday Conteh Department of Water Resources, Abuko, Gambia Gustavo Marques da Costa Programa de pós-graduação em Qualidade Ambiental, Universidade Feevale, Novo Hamburgo, RS, Brazil Universidade Feevale, Novo Hamburgo, Brazil

Contributors

xxix

Michele dos Santos Gomes da Rosa Pontifícia Universidade Católica do Rio Grande do Sul – PUCRS, Porto Alegre, Brazil Ashutosh Das Centre for Environmental Engg., PRIST Deemed University, Thanjavur, Tamil Nadu, India S. Das Centre for Interdisciplinary Research and Education, Kolkata, India Jayanta Debbarma PWD (Water Resources), Agartala, Tripura, India A. Dey Centre for Interdisciplinary Research and Education, Kolkata, India Anirban Dhulia Department of Civil Engineering, Jaypee University of Information Technology, Himachal Pradesh, Waknaghat, India Andrea Di Schino Dipartimento di Ingegneria, Università di Perugia, Perugia, Italy Annette Droste Programa de pós-graduação em Universidade Feevale, Novo Hamburgo, RS, Brazil

Qualidade

Ambiental,

A. N. Ede Covenant University, Ota, Ogun State, Nigeria Mohamed EI Yamani Polydisciplinary Faculty of Taza, Taza, Morocco Naïma El Ghachtouli Laboratoire de Biotechnologie Microbienne, Faculté des Sciences et Techniques, Université Sidi Mohamed Ben Abdellah, Fes, Morocco Ahmed El Harfi Laboratory of Agro resources Polymers and Process engineering (LAPPE), Team of Macromolecular and Organic Chemistry, Faculty of sciences, Ibn Tofail University, Kenitra, Morocco Bouchra El Hilal Laboratory of Agro resources Polymers and Process engineering (LAPPE), Team of Macromolecular and Organic Chemistry, Faculty of sciences, Ibn Tofail University, Kenitra, Morocco Operation Unit of the Radioactive Waste, Center of Nuclear Studies of Maamora (CENM) (CNESTEN), Kenitra, Morocco Abdelaziz El Moussaouy Laboratory of Dynamics and Optics of Materials, Department of Physics, Faculty of Sciences, Mohammed University; CRMEF-O, Oujda, Morocco Soumya Elabed Laboratoire de Biotechnologie Microbienne, Faculté des Sciences et Techniques, Université Sidi Mohamed Ben Abdellah, Fes, Morocco Fouad Elame National Institute of Agricultural Research, Research Center of Agadir, Agadir, Morocco Esraa M. El-Fawal Analysis and Evaluation Department, Egyptian Petroleum Research Institute, Nasr City, Cairo, Egypt Ahmed Elharfi Laboratory of Agro Resources Polymers and Process engineering (LAPPE), Team of Macromolecular and Organic Chemistry, Faculty of sciences, Ibn Tofail University, Kenitra, Morocco

xxx

Contributors

Reda Elkacmi Department of Chemistry and Valorisation, Faculty of Sciences Ain-Chock, HASSAN II University of Casablanca, BP, Casablanca, Morocco Department of Process Engineering, National School of Mineral Industries of Rabat, BP, Rabat, Morocco M. S. Elyoubi Laboratory of Chemistry of Solid State, Faculty of Science, Ibn Tofail University, Kenitra, Morocco Diyouke M. Eugene Cameroon Wildlife Conservation Society (CWCS), CWCS Coastal, Forests and Mangrove Programme, Mouanko, Littoral Region, Cameroon A.-H. El Foulani MSISM Research Team, Department of Physics, Polydisciplinary Faculty of Safi, Cadi Ayyad University, Safi, Morocco Paul Fourounjian Waksman Institute of Microbiology, Rutgers University, Piscataway, NJ, USA Ane Katiussa Siqueira Frohlich UFMG, Belo Horizonte, Brazil Suchirai Gaikwad Department of Environmental Sciences, Dr. Babasaheb Ambedkar Marathwada University, Aurangabad, Maharashtra, India C. Galey Agence nationale de santé publique 12 rue du Val d’Osne 94415, Saint Maurice Cedex, France Rajiv Ganguly Department of Civil Engineering, Jaypee University of Information Technology, Waknaghat, Himachal Pradesh, India Payal Garg Centre for Rural Development and Technology, Indian Institute of Technology Delhi, New Delhi, India Vijay Kumar Garlapati Department of Biotechnology and Bioinformatics, Jaypee University of Information Technology, Waknaghat, Himachal Pradesh, India A. Gayen Centre for Interdisciplinary Research and Education, Kolkata, India Gennady Gerasimov Institute of Mechanics, Moscow State University, Moscow, Russia Pooja Ghosh Centre for Rural Development and Technology, Indian Institute of Technology, New Delhi, India Brandon H. Gilroyed School of Environmental Science, University of Guelph Ridgetown Campus, Ridgetown, ON, Canada W. M. Gitari Department of Ecology and Resource Management, University of Venda, Thohoyandou, South Africa Mukesh Goel Centre for Environmental Engg., PRIST Deemed University, Thanjavur, Tamil Nadu, India Pooja Goel Centre for Advanced Study in Geology, University of Lucknow, Lucknow, India

Contributors

xxxi

Hicham Gourgue Sustainable Innovation and Applied Research Laboratory (SIARL) Universiapolis, International University of Agadir, Agadir, Morocco Leonor Patricia Güereca Instituto de Ingeniería, Universidad Nacional Autónoma de México, Coyoacán, Ciudad Universitaria, Mexico City, Mexico Pallav Gupta Department of Mechanical Engineering, A.S.E.T., Amity University, Noida, Uttar Pradesh, India Smriti Gupta Faculty of Civil Engineering, Shri Ramswaroop Memorial University, Lucknow, India Vaibhav Gupta Council on Energy, Environment and Water (CEEW), New Delhi, India Kumar Harshit Department of Mechanical Engineering, A.S.E.T., Amity University, Noida, Uttar Pradesh, India Aziz Hasib Laboratory of Environment and Valorization of Agro-Resources, Sultan Moulay Slimane University, Beni Mellal, Morocco G. N. Tanjina Hasnat Department of Land Administration, Faculty of Land Management and Administration, Patuakhali Science and Technology University, Dumki, Patuakhali, Bangladesh Susan Hayeri Yazdi Gas Turbine Power Plant Division, Monenco Iran Company, Tehran, Iran Sandra Magali Heberle Energy, Environment and Health Research Unit, FP-ENAS - UFP: University Fernando Pessoa, Porto, Portugal Kamal Hirech Faculty of Science, Department of Physics, Team: Materials, Electronics and Renewable Energies MERE, Laboratory Electromagnetism Signal Processing and Renewable Energy LESPRE, Mohamed Premier University, Oujda, Morocco Raza R. Hoque Department of Environmental Science, Tezpur University, Tezpur, Assam, India Md. Akhter Hossain Institute of Forestry and Environmental Sciences Chittagong University, University of Chittagong, Chittagong, Bangladesh A. M. M. Maruf Hossain School of Global, Urban and Social Studies, College of Design and Social Context, RMIT University, Melbourne, VIC, Australia Center for Integrated Knowledge Invention, Laverton, VIC, Australia Chaudhery Mustansar Hussain Department of Chemistry and Environmental Sciences, New Jersey Institute of Technology, Newark, NJ, USA Karishma Hussain Department of Bioengineering and Technology, Gauhati University, Guwahati, Assam, India

xxxii

Contributors

Farhaz Liaquat Hussain Research Scholar, Department of Chemistry, Dibrugarh University, Dibrugarh, Assam, India Saad Ibnsouda Koraichi Laboratoire de Biotechnologie Microbienne, Faculté des Sciences et Techniques, Université Sidi Mohamed Ben Abdellah, Fes, Morocco Centre Universitaire Régional d’Interface, Université Sidi Mohamed Ben Abdellah, Fes, Morocco Mohammad Ghaznavi Idris Department of Bioengineering and Technology (GUIST), Gauhati University, Guwahati, India P. Ilaiyaraja Department of Physics, Indian Institute of Technology Madras, Chennai, Tamil Nadu, India Syed Ali Hussain Jafri Department of Mechanical Engineering, Integral University, Lucknow, Uttar Pradesh, India Kwasu Jammeh Department of Parks and Wildlife Management, Kanifing, Gambia Saiko Janko KOMFORA, The Kombo/Foni Forestry Association, Kafuta, Gambia Ansumana K. Jarju NARI, Brikama, West Coast Region, Gambia Dipika Jaspal Symbiosis Institute of Technology, Symbiosis International (Deemed University), Lavale, Pune, India Daniel Jato-Espino GITECO Research Group, University of Cantabria, Santander, Cantabria, Spain M. K. Jhariya Department of Farm Forestry, Sarguja University, Ambikapur, Chhattisgarh, India Onkar Jogdand Department of Environmental Sciences, Dr. Babasaheb Ambedkar Marathwada University, Aurangabad, Maharashtra, India W. I. Jose Department of Chemical Engineering, University of the Philippines, Quezon City, Metro Manila, Philippines Md. Alamgir Kabir Department of Agroforestry, Patuakhali Science and Technology University, Dumki, Patuakhali, Bangladesh Ajay S. Kalamdhad Department of Civil Engineering, Indian Institute of Technology Guwahati, Guwahati, India Saliha Saadet Kalender Istanbul Technical University, Tirana, Albania S. Kar Centre for Interdisciplinary Research and Education, Kolkata, India Khalil Kassmi Faculty of Science, Department of Physics, Team: Materials, Electronics and Renewable Energies MERE, Laboratory Electromagnetism Signal Processing and Renewable Energy LESPRE, Mohamed Premier University, Oujda, Morocco

Contributors

xxxiii

Sampriti Kataki Energy Conservation Laboratory, Department of Energy, Tezpur University, Tezpur, Assam, India Anubha Kaushik University School of Environment Management, Guru Gobind Singh Indraprastha University, Dwarika, New Delhi, India Geetanjali Kaushik MGM’s Jawaharlal Nehru Engineering College, Mahatma Gandhi Mission, Aurangabad, Maharashtra, India Satoru Kawasaki Faculty of Engineering, Hokkaido University, Sapporo, Japan Rüstem Keçili Yunus Emre Vocational School of Health Services, Department of Medical Services and Techniques, Anadolu University, Eskişehir, Turkey Shraddha Khamparia Symbiosis Centre for Research and Innovation, Symbiosis International (Deemed University), Lavale, Pune, India Md. Nakibul Hasan Khan Department of Environmental Science and Engineering, Jatiya Kabi Kazi Nazrul Islam University, Mymensingh, Bangladesh Mohd Yawar Ali Khan Department of Earth Sciences, Indian Institute of Technology Roorkee, Roorkee, India Sunil Kumar Khare Department of Chemistry, Indian Institute of Technology, New Delhi, India Kamal Khodaei Department of Environmental Geology, Research Institute of Applied Sciences (ACECR), Shahid Beheshti University, Tehran, Iran Mohaddeseh Khorasanizadeh Department of Chemical Engineering, Sharif University of Technology, Tehran, Iran Mohammed Hussein Rafeq Khudhair Laboratory of Agro resources Polymers and Process engineering (LAPPE), Team of Macromolecular and Organic Chemistry, Faculty of sciences, Ibn Tofail University, Kenitra, Morocco Laboratory of Cement and Quality Control of Amran Cement Plant of Yemen, Amran, Yemen Laboratory of Chemistry of Solid State, Faculty of Science, Ibn Tofail University, Kenitra, Morocco A. K. Khurana Central Council for Research in Homeopathy, New Delhi, India Daniel Kibami Department of Chemistry, Kohima Science College (Autonomous) Jotsoma, Kohima, Nagaland, India Stamatia Kontogianni Laboratory of Heat Transfer and Environmental Engineering, Department of Mechanical Engineering, Aristotle University of Thessaloniki, Thessaloniki, Greece Raj Kumar Vallabhbhai Patel Chest Institute, University of Delhi, New Delhi, India

xxxiv

Contributors

S. Lakshmana Prabu Department of Pharmaceutical Technology, Bharathidasan Institute of Technology, Anna University, Tiruchirappalli, Tamil Nadu, India Sze-Mun Lam Department of Environmental Engineering, Faculty of Engineering and Green Technology, Universiti Tunku Abdul Rahman, Kampar, Perak, Malaysia Christof Lanzerstorfer University of Applied Sciences Upper Austria, Wels, Austria Ronald Kondo Lembang Faculty of Natural Science and Engineering Technology, University of Halmahera, Tobelo, North Mollucas, Indonesia Hayat Lionboui Department of Economics and Rural Sociology, National Institute of Agricultural Research, Research Center of Tadla, Beni Mellal, Morocco Laboratory of Environment and Valorization of Agro-Resources, Sultan Moulay Slimane University, Beni Mellal, Morocco Esther Lizasoain-Arteaga GITECO Research Group, Universidad de Cantabria, Santander, Spain Israel López Universidad Autónoma de Nuevo León, UANL, Facultad de Ciencias Químicas, Laboratorio de Materiales I, San Nicolás de los Garza, Nuevo León, Mexico Sandra Manoela Dias Macedo Departamento de Farmacociências, UFCSPA, Porto Alegre, Brazil Hisham A. Maddah Department of Chemical Engineering, King Abdulaziz University, Rabigh, Saudi Arabia Morteza Mahmoudi Energy Research Institute, Sharif University of Technology, Tehran, Iran Arunabha Majumder School of Water Resources Engineering, Jadavpur University, Kolkata, West Bengal, India All India Institute of Hygiene and Public Health, Govt. of India, Kolkata, India R. K. Manchanda Central Council for Research in Homeopathy, New Delhi, India Jubhar C. Mangimbulude Faculty of Natural Science and Engineering Technology, University of Halmahera, Tobelo, North Mollucas, Indonesia Meghna Manjunath R & D Centre, Department of Biotechnology Engineering, Dayananda Sagar Institutions, Bengaluru, India Visvesvaraya Technological University, Belagavi, Karnataka, India S. Manna Centre for Interdisciplinary Research and Education, Kolkata, India Asis Mazumdar School of Water Resources Engineering, Jadavpur University, Kolkata, West Bengal, India

Contributors

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Gordon McKay Division of Sustainable Development, College of Science and Engineering, Hamad Bin Khalifa University, Qatar Foundation, Doha, Qatar Subhash Medhi Department of Bioengineering and Technology (GUIST), Gauhati University, Guwahati, India Souad Mehalaine Department of Natural and Life Sciences, Faculty of Exact Sciences and Natural and Life Sciences, Larbi Tebessi University, Tebessa, Algeria Larissa Meincke Universidade Feevale, Novo Hamburgo, Brazil Mustapha Melhaoui Faculty of Science, Department of Physics, Team: Materials, Electronics and Renewable Energies MERE, Laboratory Electromagnetism Signal Processing and Renewable Energy LESPRE, Mohamed Premier University, Oujda, Morocco Bahia Meroufel Faculty of Technology, University Abou Bekr Belkaïd of Tlemcen, Tlemcen, Algeria Radin Mohamed Micro-Pollutant Research Centre (MPRC), Department of Water and Environment Engineering, Faculty of Civil and Environmental Engineering, University Tun Hussein Onn Malaysia (UTHM), Parit Raja, Johor, Malaysia Abdul Rahman Mohamed School of Chemical Engineering, Universiti Sains Malayisia, Engineering Campus, Nibong Tebal, Pulau Pinang, Malaysia Noor Mohammad Department of Law, Green University of Bangladesh, Dhaka, Bangladesh Alsharifa Hind Mohammad Water, Energy and Environment Center, The University of Jordan, Amman, Jordan Sathaporn Monprapussorn Department of Geography, Faculty of Social Sciences, Srinakharinwirot University, Bangkok, Thailand Jean-Hude E. Moudingo FAO Cameroon Representation Mangrove Project and Cameroon Wildlife Conservation Society (CWCS), CWCS Coastal, Forests and Mangrove Programme, Mouanko, Littoral Region, Cameroon Nicolas Moussiopoulos Laboratory of Heat Transfer and Environmental Engineering, Department of Mechanical Engineering, Aristotle University of Thessaloniki, Thessaloniki, Greece Jitendra Kumar Nagar Vallabhbhai Patel Chest Institute, University of Delhi, New Delhi, India Richa Dave Nagar Amity Institute of Environmental Sciences, Amity University, Noida, Uttar Pradesh, India Soumya Nair Microbial Biotechnology Laboratory, School of Biosciences and Technology, VIT University, Vellore, Tamil Nadu, India P. Nandy Centre for Interdisciplinary Research and Education, Kolkata, India

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Contributors

D. Nayak Central Council for Research in Homeopathy, New Delhi, India Modou Njei WABSA West Africa Study Birds Association, Banjul, Gambia Saeed Nojavan Department of Analytical Chemistry and Pollutants, Shahid Beheshti University, G. C., Evin, Tehran, Iran Efaq Ali Noman Faculty of Applied Sciences and Technology (FAST), Universiti Tun Hussein Onn Malaysia, Pagoh, Johor, Malaysia O. M. Olofinnade Covenant University, Ota, Ogun State, Nigeria Daniela Montanari Migliavacca Osório Programa de Pós-graduação em Qualidade Ambiental, Universidade Feevale, Novo Hamburgo, RS, Brazil A. Ouigmane Laboratory of Spectro-Chemistry Applied and Environment, University Sultan MoulaySlimane, BeniMellal, Morocco Laboratory of Environment and Valorization d of Agro-Resources, University Sultan MoulaySlimane, BeniMellal, Morocco Transdisciplinary Team of Analytical Sciences for Sustainable Development, University Sultan MoulaySlimane, BeniMellal, Morocco K. Ouzaouit REMINEX research center, Groupe Managem, Marrakech/Medina, Morocco Somnath Pal School of Water Resources Engineering, Jadavpur University, Kolkata, West Bengal, India Sudipta Pal Nutritional Biochemistry and Toxicology Laboratory, Department of Human Physiology, Tripura University (A Central University), Suryamaninagar, West Tripura, India Sukanchan Palit Department of Chemical Engineering, University of Petroleum and Energy Studies, Energy Acres, Dehradun, Uttarakhand, India Rashmi Paliwal Institute of Environmental Studies, Kurukshetra University, Kurukshetra, India Niraj Singh Parihar Department of Civil Engineering, Jaypee University of Information Technology, Waknaghat, Himachal Pradesh, India Musarrat Parween Water Programme, National Institute of Advanced Studies, Bengaluru, Karnataka, India Anju Patel Faculty of Civil Engineering, Shri Ramswaroop Memorial University, Barabanki, India Amulya A. Patil R & D Centre, Department of Biotechnology Engineering, Dayananda Sagar Institutions, Bengaluru, India Visvesvaraya Technological University, Belagavi, Karnataka, India Dayananda Sagar Institutions, Bengaluru, Karnataka, India

Contributors

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Satish Patil Department of Environmental Sciences, Dr. Babasaheb Ambedkar Marathwada University, Aurangabad, Maharashtra, India B. K. Paul Centre for Interdisciplinary Research and Education, Kolkata, India Hamid M. Pouran Faculty of Science and Engineering, University of Wolverhampton, Wolverhampton, UK N. Priyadarshini Department of Chemistry, SSN College of Engineering, Kalavakkam, Tamil Nadu, India Dunja Prokić Faculty of Environmental Protection, University Educons, Sremska Kamenica, Serbia R. C. Pullar Department of Materials and Ceramic Engineering, Department of Physics, CICECO – Aveiro Institute of Materials, University of Aveiro, Aveiro, Portugal P. N. Puniya Department of Environmental Sciences, Dr. Babasaheb Ambedkar Marathwada University, Aurangabad, Maharashtra, India Abdul Qadir College of Earth and Environmental Science (CEES), Faculty of Science, University of the Punjab, Lahore, Pakistan R. M. S. Radin Mohamed Micro-Pollution Research Centre (MPRC), Department of Water and Environmental Engineering, Faculty of Civil and Environmental Engineering, Universiti Tun Hussein Onn Malaysia, Parit Raja, Johor, Malaysia Mirzanur Rahman Department of Information Technology (GUIST), Gauhati University, Guwahati, India A. Raj Department of Forestry, College of Agriculture, I.G.K.V, Raipur, CG, India A. L. Ramanathan School of Environmental Sciences, Jawaharlal Nehru University, New Delhi, India Manviri Rani Department of Chemistry, Dr B R Ambedkar National Institute of Technology, Jalandhar, Punjab, India Nisha Rani Department of BioSciences, Himachal Pradesh University, Shimla, India T. E. Rasimphi Department of Ecology and Resource Management, University of Venda, Thohoyandou, South Africa Mohd Adib Mohammad Razi Micro-Pollutant Research Centre (MPRC), Department of Water and Environment Engineering, Faculty of Civil and Environmental Engineering, University Tun Hussein Onn Malaysia (UTHM), Parit Raja, Johor, Malaysia M. Rengasamy Department of Petrochemical Technology, Bharathidasan Institute of Technology, Anna University, Tiruchirappalli, Tamil Nadu, India

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Contributors

Yahia Rharrabti Polydisciplinary Faculty of Taza, Taza, Morocco Jan W. Rijstenbil AE3 Consultancy, Yerseke, The Netherlands Maria Romero-Gonzalez Department of Geography, University of Sheffield, Sheffield, UK Michele S. G. Rosa Pontifícia Universidade Católica do Rio Grande do Sul – PUCRS, Porto Alegre, Brazil C. Rosin Agence nationale de sécurité sanitaire de l’alimentation, de l’environnement et du travail (Anses) Direction de l’Evaluation des Risques, Unité évaluation des risques liés à l’eau, Maisons-Alfort Cedex, France Pankaj Kumar Roy School of Water Resources Engineering, Jadavpur University, Kolkata, West Bengal, India Malabika Biswas Roy Women’s College, Calcutta, Kolkata, India Anand Sagar Department of BioSciences, Himachal Pradesh University, Shimla, India Vipin K. Saini School of Environment and Natural Resources, Doon University, Dehradun, Uttrakhand, India Muhammad Shabery Sainudin Micro-Pollution Research Centre (MPRC), Department of Water and Environmental Engineering, Faculty of Civil and Environmental Engineering, Universiti Tun Hussein Onn Malaysia, Parit Raja, Johor, Malaysia Junaid Saleem Division of Sustainable Development, College of Science and Engineering, Hamad Bin Khalifa University, Qatar Foundation, Doha, Qatar Niloufar Salehi Energy Research Institute, Sharif University of Technology, Tehran, Iran P. Sankar Ganesh Department of Biological Sciences, BITS Pilani, Hyderabad Campus, Hyderabad, Telangana, India Debanjan Sannigrahi Advanced Management Institute for Training and Achievement (AMITA), Mumbai, Maharashtra, India A. K. Sannigrahi Proof & Experimental Establishment (PXE), Defence Research and Development Organization, Balasore, India Luis A. Sañudo-Fontaneda GICONSIME Research Group, University of Oviedo, Mieres, Asturias, Spain Abhishek Saxena Faculty of Civil Engineering, Sri Ramswaroop Memorial University, Barabanki, Uttar Pradesh, India S. Sendilvelan Department of Mechanical Engineering, Dr. M.G.R Educational and Research Institute, University, Chennai, India

Contributors

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Philippe Serp Laboratoire de Chimie de Coordination UPR CNRS 8241, composante ENSIACET, Université de Toulouse, UPS-INP-LCC, Toulouse Cedex 4, France Muafia Shafiq College of Earth and Environmental Science (CEES), Faculty of Science, University of the Punjab, Lahore, Pakistan Biotechnology and Food Research Centre, PCSIR Laboratories Complex, Lahore, Pakistan Usman Bin Shahid Division of Sustainable Development, College of Science and Engineering, Hamad Bin Khalifa University, Qatar Foundation, Doha, Qatar Apoorva Shankar R & D Centre, Department of Biotechnology Engineering, Dayananda Sagar Institutions, Bengaluru, India Visvesvaraya Technological University, Belagavi, Karnataka, India Aparajita Shankar School of Environment and Natural Resources, Doon University, Dehradun, Uttrakhand, India Uma Shanker Department of Chemistry, Dr B R Ambedkar National Institute of Technology, Jalandhar, Punjab, India Hardeep Rai Sharma Institute of Environmental Studies, Kurukshetra University, Kurukshetra, Haryana, India Kanu Shil Department of Human Physiology, Tripura University, Agartala, Tripura, India J. P. Shrivastava Centre of Advanced Studies and Department of Geology, University of Delhi, New Delhi, India Saurabh Shukla Faculty of Civil Engineering, Shri Ramswaroop Memorial University, Lucknow, India Zia Mahmood Siddiqi Jubail University College, Jubail Industrial City, Saudi Arabia Jin-Chung Sin Department of Petrochemical Engineering, Faculty of Engineering and Green Technology, Universiti Tunku Abdul Rahman, Kampar, Perak, Malaysia Poonam Singhal Centre for Rural Development and Technology, Indian Institute of Technology Delhi, New Delhi, India Indian Institute of Management, Calcutta, India Sinosh Skariyachan R & D Centre, Department of Biotechnology Engineering, Dayananda Sagar Institutions, Bengaluru, India Visvesvaraya Technological University, Belagavi, Karnataka, India Salah Souabi Laboratory of Process Engineering and Environment, Faculty of Sciences and Techniques, Mohammedia, University – Hassan II, Mohammedia, Morocco

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Contributors

Lekha Sridhar Consultant - United Nations Environment Programme (UNEP), New Delhi, India Nataša Stojić Faculty of Environmental Protection, University Educons, Sremska Kamenica, Serbia Snežana Štrbac Faculty of Environmental Protection, University Educons, Sremska Kamenica, Serbia Sumathi Suresh Centre for Environmental Science and Engineering, Indian Institute of Technology Bombay, Mumbai, Maharashtra, India Anushree Suresh Microbial Biotechnology Laboratory, School of Biosciences and Technology, VIT University, Vellore, Tamil Nadu, India Swati School of Environmental Sciences, Jawaharlal Nehru University, New Delhi, India Saul Taal Ministry of Agriculture/Research Institute, Banjul, Gambia Hadi Tabani Department of Environmental Geology, Research Institute of Applied Sciences (ACECR), Shahid Beheshti University, Tehran, Iran Vinod Tare Department of Civil Engineering, IIT Kanpur, Kanpur, India Indu Shekhar Thakur School of Environmental Sciences, Jawaharlal Nehru University, New Delhi, India D. Tinarwo Department of Physics, University of Venda, Thohoyandou, South Africa Lamin Mai Touray Department of Water Resources, Banjul, Gambia A. Umamaheswari Department of Pharmaceutical Technology, Bharathidasan Institute of Technology, Anna University, Tiruchirappalli, Tamil Nadu, India C. Veluchamy Department of Civil Engineering, Indian Institute of Technology Guwahati, Guwahati, India School of Environmental Science, University of Guelph Ridgetown Campus, Ridgetown, ON, Canada G. Venkatesan Department of Civil Engineering, Bharathidasan Institute of Technology, Anna University, Tiruchirappalli, Tamil Nadu, India Virendra Kumar Vijay Centre for Rural Development and Technology, Indian Institute of Technology, New Delhi, India Ali Vosoogh Department of Civil Engineering, Iran University of Science and Technology, Tehran, Iran Wenqin Wang School of Agriculture and Biology, Shanghai Jiao Tong University, Shanghai, China

Contributors

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A. N. Welfrad Micro-pollution Research Centre (MPRC), Department of Water and Environmental Engineering, Faculty of Civil and Environmental Engineering, Universiti Tun Hussein Onn Malaysia, Parit Raja, Johor, Malaysia D. K. Yadav Department of Farm Forestry, Sarguja University, Ambikapur, Chhattisgarh, India Kaan Yetilmezsoy Department of Environmental Engineering, Faculty of Civil Engineering, Yildiz Technical University, Istanbul, Esenler, Turkey Sherif A. Younis Analysis and Evaluation Department, Egyptian Petroleum Research Institute, Nasr City, Cairo, Egypt Laboratoire de Chimie de Coordination UPR CNRS 8241, composante ENSIACET, Université de Toulouse, UPS-INP-LCC, Toulouse Cedex 4, France A. Zeghnoun Agence nationale de santé publique 12 rue du Val d’Osne 94415, Saint Maurice Cedex, France Mohamed Amine Zenasni Faculty of Technology, University Abou Bekr Belkaïd of Tlemcen, Tlemcen, Algeria

Part I Major Environmental Issues and Problems

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Assessment of Some Aspects of Provisioning Sewerage Systems: A Case Study of Urban Agglomerations in Ganga River Basin Saurabh Shukla and Vinod Tare

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cost Estimates of Sewerage Systems: Conventional Approach . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Collection of Information . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Final Cost Estimation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cost Estimates of Sewerage Systems: Other Approaches . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sewerage Network . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sewage Pumping . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sewage Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Methodology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Estimation of Capex and Opex of Sewerage Network . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Estimation of Capex and Opex for Sewage Pumping . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Estimation of Capex and Opex of Sewage Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Findings . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sewerage Network . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cost of Sewage Pumping . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cost of Sewage Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cost Estimation of Complete Sewerage System . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Estimated Costs of Provisioning Sewerage Systems in Major Urban Agglomerations in Ganga River Basin . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Benefits of Provisioning Sewerage Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Discussion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

4 6 7 7 8 8 9 9 10 10 12 12 12 12 17 19 22 24 27 32 33 33

S. Shukla (*) Faculty of Civil Engineering, Shri Ramswaroop Memorial University, Lucknow, India e-mail: [email protected] V. Tare Department of Civil Engineering, IIT Kanpur, Kanpur, India e-mail: [email protected] © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_3

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Abstract

An appropriate techno-commercial framework is a prerequisite for sustainable sanitation solutions in urban centers. The first step toward developing such a framework is to have an assessment of provisioning sanitation systems in economic sense. The present study aims at estimating the per capita expenditure on sewerage system. The energy consumption and footprint are also important along with expenditure incurred and hence are also estimated separately. Results indicate that footprint for sewage treatment is approximately 0.1 m2 per person which is one tenth of the size of the toilet. The energy consumption in sewage pumping and treatment ranges from 0.03 to 0.1 KW-h (kilowatt-hour) which is equivalent to lighting 30 to 100 watt bulb for 1 h. The total per capita expenditure in availing sewerage infrastructure is estimated to be in the range INR (Indian Rupee Rates) 1.8–10.8 with an average of INR 3.93 and standard deviation 1.4. The higher values correspond to towns with very low population density, and the lower values correspond to very high population densities. Keywords

Sewerage systems · Sewerage network · Sewage pumping · Sewage treatment · Cost estimates · Capex · Opex · Ganga River Basin · Ganga River Basin Management Plan

Introduction Consortium of seven IITs (Indian Institute of Technology) has been engaged by the Government of India to prepare Ganga River Basin Management Plan (GRBMP). One of the most important challenges of the Consortium is to prepare an action plan for “unpolluted flow” or “Nirmal Dhara” in all rivers of the Ganga Basin. The main approach to achieve the ultimate objective of “Nirmal Dhara” has been to identify the type of polluting wastes, their sources of generation (point and nonpoint sources), and the techno-economic feasibility of collecting and treating them for their safe environmental discharge and/or possible recycle or reuse. Figure 1 illustrates the main identification results and the tasks. Among the point sources, urban and industrial wastewaters are the major sources of pollution (Luzio et al. 2003), needing immediate remediation. In consideration of the magnitude of domestic wastewater generation from different urban locales, urban settlements are divided into Class I towns (having a population over 100,000) and Class II towns (having a population between 50,000 and 100,000) (RBI 2003). The following main steps concerning sewerage infrastructure for medium to long term (over the next 25 years) are considered essential. 1. Complete stoppage of the discharge of sewage, either treated or untreated, from Class I and Class II towns to any river. 2. All sewage generated in Class I and Class II towns of GRB needs to be collected and treated up to the tertiary level with treated effluent standards of Biochemical

1

Assessment of Some Aspects of Provisioning Sewerage Systems: A Case. . .

5

Fig. 1 Types and sources of wastes and main identification tasks (IIT GRB 2013)

Oxygen Demand (BOD) C=O group; at 1000 cm1 represents C–Cl bonds; at 2750–3000 cm1 represents changes in alkane group (C–H) (Divyalakshmi and Subhashini 2016); at 1710–1750 cm1 represents formation of ketone or aldehyde C=O groups; at 1000 and 1700 cm1 represent oxidized fractions, such as moieties containing –OH groups (Esmaeili et al. 2013); and at 1465 cm1 represents CH2.

Gas Chromatography-Mass Spectrometry (GC-MS) This technique provides useful information about the molecular structure of the macromolecules. It also helps to study the gaseous products produced during

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degradation of plastic polymer. In recent years, the method is used in the structure study of the biodegradable synthetic polymer, comprising the following (Chakraborty and Das 2017): • Verification of chemical homogeneity of the polymer and to determine the chemical structure of end groups • Determination of the molar mass and weight distribution of degradation end products • Determination of the chemical composition and distribution of the sequence of units to individual macromolecules of copolymers One of the previous studies suggested the presence of compounds such as 2-butene, 2-methyl, acetone, ethane, alkanes, aromatic compounds, and fatty acids such as hexadecanoic acid and octanoic acid based on GC-MS analysis (Pramila and Ramesh 2015).

X-Ray Diffraction (XRD) The level of crystallinity of the plastic can be assessed using this technique. As the biodegradation progresses, the crystallinity of plastic is seen to decrease due to the breakage of bonds by the microbial consortia. Previous studies carried out with pretreatment of polyethylene with UV radiation have shown that the peaks of UV-irradiated films are higher than that of non-UV irradiated. This difference illustrates that oxidation of the plastic increases its crystallinity (Esmaeili et al. 2013).

Atomic Force Microscopy (AFM) AFM helps to analyze the surface morphology of the polyethylene films, degraded by the action of microbial consortia. AFM observations are based on interatomic forces. The plastic becomes rough due to the formation of cracks and grooves on the surface post degradation (Ojha et al. 2017; Chakraborty and Das 2017).

Scanning Electron Microscopy (SEM) SEM is an instrument employed to check the biofilm formation and to confirm degradation. The biodegradation of low-density polyethylene by microbial consortia is shown in Fig. 4. Dehydrated polyethylene is sputter-coated with a gold layer. The sputtering can be achieved after passing the pure and dry argon gas in the coating chamber, under vacuum. The plate voltage was 2000 V, and the current passed was 15 mA. A thickness of 2 nm of gold was achieved during a sputtering time of 10 s. The polyethylene was then examined under the scanning electron microscope. Few

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Application of Novel Microbial Consortia for Environmental Site. . .

a

100 μm

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b

EHT = 5.00 kV WD = 9.5 mm

Signal A = InLens Mag = 231 X

2 μm

EHT = 5.00 kV WD = 9.3 mm

Signal A = InLens Mag = 12.42 K X

c

2 μm

EHT = 5.00 kV WD = 9.3 mm

Signal A = SE2 Mag = 19.08 K X

Fig. 4 The biodegradation of low-density polyethylene by formulated microbial consortia characterized by scanning electron microscope: (a) control, (b) attachment of rod-shaped bacterial consortia on the surface of low-density polyethylene, (c) structural changes in LDPE observed during the biodegradation by microbial consortia

structural changes such as grooves, cracks, damaged layer, fragileness, pits, and roughening of the surface can be observed along with the biofilm formed (Ojha et al. 2017). Using BacLight Bacterial Viability Kit along with SEM helps determine the cell viability. The kit helps in monitoring the viability of bacterial populations as a function of the membrane integrity of the cell. The live cells take up the green dye and hence appear green in the SEM image, whereas the dead cells take up red dye and appear red (Sivan et al. 2006).

Infrared Methane Gas In Situ Analyzer It is an instrument employed to analyze the methane gas produced along with biogas. Usually, it is seen that as the days of incubation increase, the methane produced also increases (Wan et al. 2013).

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Nuclear Magnetic Resonance (NMR) Nuclear magnetic resonance (NMR) is the technique of determining physical and chemical properties of atom and molecules by exploiting their magnetic properties. It helps to study the detailed structural information, dynamics, reaction rate, and the chemical environment of molecules by exploiting the phenomenon of NMR. Molecules having nuclei possessing spin can be investigated by this technique; therefore, organic molecules are the primary compounds for analysis. The resonant frequency, the energy of the absorption, and the intensity of the signal is dictated by the strength of the magnetic field. NMR response of free induction decay is generated by exciting the sample with radio-frequency pulse. Thus, two spin states are generated in the presence of external magnetic field wherein one aligns with the magnetic field and the other aligns opposite to the field. This produces a difference in energy between the two spin states, thus increasing the field strength. Thus, it gives the chemical shift which is the resonant frequency of a nucleus relative to a standard in a magnetic field and determines the structure of the molecule (Chakraborty and Das 2017). NMR results are obtained in the form of a graph with peaks at different ppm. In one of our previous studies involving 1H-NMR analysis of LDPE and HDPE sample, it was observed that additional triplet peak between 1.0 and 1.5 ppm and a peak between 3.5 and 4.0 ppm specify the degradation of plastic, followed by the formation of a methyl group and an aldehyde moiety as end products post degradation (Skariyachan et al. 2017).

STURM Test Carbon dioxide evolved as a result of polymer degradation can be studied using this test. All degradation types produce carbon dioxide. The gas evolved is trapped using Durham’s tube and is measured using Steinfurth CO2 tester which is further compared with the standard. Previous studies have shown that all types of biodegradation by microorganisms have proven to produce carbon dioxide (Shah et al. 2013).

Recent Advances in Scale-Up of Plastic-Degrading Microbial Consortia The experiments carried out until date are all in small scale, and hence, for escalation to mid-scale or larger scale, techniques such as anaerobic digester and soil burial method are employed. The process can also be carried out in a reactor.

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Anaerobic Digester Plastic wastes are added to the digester in small scale along with the bacterial culture. The culture is retrieved. The content of carbon dioxide, methane, and biogas produced is then retrieved and used as a source of fuel. The study at mid-scale is carried out for 5–10 years. The rate of biogas produced increases at the initial part of the experiment and then comes to a steady state (Wan et al. 2013).

Scope of Enzyme Technology Enzyme-mediated processes have rapidly gained importance in detoxification of municipal solid waste, waste water, hazardous waste, and polluted soils (Le-RoesHill and Prins 2016; Shah 2017). Among biological approaches, microbial enzymes are highly efficient for the biodegradation of plastics due to their stability, targetspecific catalytic activity, and ease of production than plant and animal enzymes (Choi et al. 2015; Singh et al. 2016b). These enzymes are specific in their action on substrates; hence, multienzymatic processes can be designed to specifically target selected forms of plastic such as low- and high-density polyethylene that are detrimental to the environment (Brodhagen et al. 2015). The use of microbial consortia for environmental site remediation and hazardous waste management is extensive with a large number of enzymes participating in the biodegradation of toxic pollutants (Shah 2017). The process of bioremediation mainly depends on the enzymatic activity of microorganisms (Dash et al. 2013). The combination of microbial enzymes has been employed for the bioremediation of hazardous wastes containing toxic chemical compounds such as aromatic amines, nitriles, and phenols either by degradation or bioconversion to nontoxic products (H2O, CO2 or CH4) (Nikolaivits et al. 2017). Furthermore, with the recent biotechnological advances in the field of protein and genetic engineering, microorganisms can be genetically modified and cultured in large quantities for the production of enzymes at low cost to meet increasing demands (Dash et al. 2013). Only certain species of bacteria and fungi have proven their ability as potent bioremediation agents (Brodhagen et al. 2015; Paco et al. 2017). The process of microbial degradation depends on the type of plastic, the organisms, and the environment involved (Rayu et al. 2012). However, there will always be the involvement of extracellular and intracellular enzymes which help in the degradation of polymers at one stage or another in the process (Karigar and Rao 2011). Biodegradation of plastics is primarily due to the action of enzymes (hydrolysis and oxidation) that lead to the cleavage of the long-chain polymers into oligomers and monomers (Brodhagen et al. 2015). Microorganisms can survive in extreme environments, yet most of them prefer optimal conditions which are difficult to achieve outside the laboratory (Dash et al. 2013). The major limitations of bacterial growth

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which is under the influence of temperature, pH, moisture content, oxygen, the bioavailability of nutrients, contaminants, and the presence of other toxic compounds can be addressed by the immobilization of enzymes (Khan et al. 2013). Similarly, the gene expression of enzymes in microbial cells is regulated by specific inducers, repressors, and cofactors which may often be absent at the site of bioremediation (Quecholac-Pina et al. 2016). Hence, to overcome these drawbacks, investigations have been focused on developing methods to stabilize and utilize enzymes preferably by immobilization onto solid supports. Further, immobilized enzymes are not inhibited by the inhibitors of microbial metabolism. They can be used effectively at both low and high pollutant concentrations under extreme conditions. Immobilized enzymes specifically act against a given substrate, while microorganisms may prefer more easily degradable compounds than the pollutant (Rayu et al. 2012). In addition to being more mobile than microorganisms because of their smaller size, these features render enzymatic techniques, eco-friendly processes, and enzymes as biocatalysts (Rao et al. 2010; Choi et al. 2015). Additionally, the enzyme that has demonstrated potential degradation of a polymer in vitro does not guarantee similar in situ activity. There will be substantial variation during scaleup as a laboratory in vitro assay performed under optimized conditions does not completely represent the natural environment (Rao et al. 2010). Furthermore, if the native substrate for the microbial enzyme(s) is present in the environment, it may outcompete the plastic for enzyme binding (Quecholac-Pina et al. 2016). The degradation of foreign compounds in the environment is often performed by microbial consortia where a single species might initiate the process but may not be able to complete it independently (Khan et al. 2013; Shah 2017). Hence, development of techniques for the immobilization of multienzymes along with cofactor regeneration and retention systems is yet to be explored in the design of complex biochemical processes involving a series of chemical conversions. Immobilization methods can be divided into two general classes, namely, chemical and physical methods. Physical methods are usually characterized by weaker interactions such as hydrophobic interactions, van der Waals forces, hydrogen bonds, ionic binding and affinity binding of the enzyme with the support material, or mechanical containment of enzyme within the support (Es et al. 2015). In the chemical methods, the formation of covalent bonds is achieved either through thioether, ether, carbamate, or amide bonds between the enzyme and support material involved (Eibes et al. 2015). There are four principal techniques for immobilization of enzymes (Homaei et al. 2013), namely, adsorption, entrapment, crosslinking, and covalent bonding (Fig. 5). Physical adsorption can be defined as one of the simplest methods of reversible immobilization which involves the enzymes being physically adsorbed or attached to the nonreactive support material (Sakai et al. 2010; Es et al. 2015). It is easy to perform, and moreover, it preserves the catalytic activity of the enzyme (Mohamad et al. 2015). Entrapment can be defined as the physical restriction of an enzyme within a network or a confined space. It is an irreversible method of immobilization wherein the enzymes are entrapped in a lattice structure of fibers or in polymer membranes which allow the movement of substrate and products across the

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Fig. 5 The major techniques used for the immobilization of microbial enzyme for degradation studies: (a) physical adsorption, (b) encapsulation, (c) covalent bonding, (d) cross-linking, (e) entrapment

membrane but retain the enzyme (Wang and Zhang 2015; Beloqui et al. 2016). Cross-linking is an irreversible method of enzyme immobilization that does not require a support as the enzyme acts as its own substrate. It is also known as carrierfree immobilization and copolymerization (Sirisha et al. 2016). In this method, virtually pure enzymes are obtained by eliminating the advantages and disadvantages associated with carriers (Datta et al. 2013). Covalent bonding is one of the most widely used methods for irreversible enzyme immobilization, and it mainly depends on the formation of covalent bond between the enzyme and the support material (Sirisha et al. 2016). Covalent bond formation between the enzyme and the matrix occurred through the side-chain amino acids like cysteine (thiol group), lysine (ε-amino group), and aspartic and glutamic acids (carboxylic groups) (Mohamad et al. 2015). Covalent bonds provide powerful linkages between the lipase and its carrier matrix allowing its reuse more often than with other available immobilization methods and prevent enzyme leakage into the environment. In a study conducted by Datta et al. (2013), immobilization of microbial lipase from Yarrowia lipolytica onto octyl-agarose and octadecyl-sepabeads supports by physical adsorption resulted in higher stability (tenfold) and better yield than that

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of free lipase. Similarly, lipase from Candida rugosa physically adsorbed onto biodegradable poly(3-hydroxybutyrate-co-hydroxyvalerate) exhibited 94% residual activity at 50  C after 4 h and reusability till 12 cycles (Datta et al. 2013). Likewise, in a dual-enzyme reaction system composed of a polyester hydrolase and an immobilized carboxyl esterase TfCa from Thermobifida fusca KW3, a twofold higher yield of biodegradation end products was reported by Barth et al. (2016). Similarly, the use of immobilized cells for a mechanical size reduction in the polyurethane foam has been suggested by Cregut et al. (2013). However, an efficient biocatalytic degradation approach using immobilized cells for polyethylene has not yet been demonstrated.

Scope of Bioinformatics and Computational Biology The bioremediation of xenobiotic compounds either ex situ or in situ by individual isolates does not represent the actual behavior of microorganisms; however, it depends on cooperative metabolic activities of microbial consortia (Khan et al. 2013). Hence, in silico analysis helps in predicting the fate of a compound by all microorganisms for either partial or complete degradation to nontoxic compounds (Arora and Bae 2014). Similarly, several computational tools and databases have been developed to assist the design and implementation of microbial consortia for environmental site remediation and hazardous waste management (Fulekar and Sharma 2008). These may include biodegradation databases, chemical toxicity prediction systems, and biodegradation pathway prediction systems (Arora and Bae 2014). It must be noted that the huge amount of data generated by computational analysis needs validation to better understand the biodegradation potential of a compound by microorganisms in the environment. Further, the pathway prediction systems and phylogenetic analysis will help in minimizing the number of possible combination required for the development of novel microbial consortia (Khan et al. 2013).

Biodegradation Databases Biodegradation databases store information related to biodegradation of chemicals including xenobiotic-degrading bacteria, metabolic degradation pathways of toxic chemicals, and enzymes and genes involved in the biodegradation (Arora and Bae 2014). These databases include the biodegradation network-molecular biology database (Bionemo), University of Minnesota Biocatalysis/Biodegradation Database (UM-BBD), BioCyc, and MetaCyc (Arora and Bae 2014).

Biodegradation Network-Molecular Biology Biodegradation network-molecular biology database (Bionemo) is a manually curated database that provides information pertaining to proteins and genes involved

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in the biodegradation metabolism (Carbajosa et al. 2009). It was developed by the Structural Computational Biology Group at the Spanish National Cancer Research Centre, and it can be freely accessed at https://bionemo.bioinfo.cnio.es. Bionemo links sequence database entries to biodegradation reactions by the information retrieved from published research articles. The database comprises of 234 microbial species, 1107 proteins, 212 transcription units, 90 transcription factors, 945 reactions, 537 enzymatic complexes, and 145 biochemical pathways (Carbajosa et al. 2009; Shah 2017). Like other databases, Bionemo is also cross-linked to GenBank for DNA sequences, UniProt for protein sequences, NCBI Taxonomy for microbial species, UM-BBD for metabolic reactions, and PubMed for references (Gao et al. 2010). The information available in the Bionemo database may aid in primer design, for cloning experiments, and directed evolution experiments. The complete database can also be freely downloaded from PostgreSQL (https://www.postgresql.org/).

University of Minnesota Biocatalysis/Biodegradation Database The University of Minnesota Biocatalysis/Biodegradation Database (UM-BBD) is a widely used database in the field of biodegradation, and it is freely available at http:// eawag-bbd.ethz.ch/. This database provides information regarding microbial biocatalytic reactions, enzymes, genes, and biodegradation pathways involved in bioremediation (Gao et al. 2010). The metabolic pathways of various xenobiotic compounds are available in text and graphical formats. These manually annotated metabolic pathways display multistep enzymatic reactions in a series initiating from the starting compound. Currently, the UM-BBD database includes 543 microbes, 993 enzymes, 1503 chemical reactions, and 219 microbial degradation pathways which are cross-linked to several other databases including NCBI (https://www.ncbi. nlm.nih.gov/), BRENDA (http://www.brenda-enzymes.org/), ExPASy (http://www. expasy.org/), and ENZYME (http://enzyme.expasy.org/) to provide vital information describing genes and enzymes involved in the biodegradation of xenobiotic compounds (Shah 2017). Additionally, this database includes a Biochemical Periodic Table (UM-BPT) and a rule-based Pathway Prediction System (UM-PPS) that predicts plausible pathways for microbial degradation of organic compounds (Gao et al. 2010). UM-BBD compound data are also linked to public chemical databases such as PubChem and ChemSpider.

MetaCyc and BioCyc MetaCyc database is a collection of experimentally elucidated metabolic pathways and enzymes highly curated from primary scientific literature from all domains of life (Caspi et al. 2016). MetaCyc is currently the largest collection of metabolic pathways with more than 2400 pathways derived from more than 46,000 publications. It comprises of 2526 experimentally determined metabolic pathways from more than 2844 different organisms (Shah 2017). This database hosts metabolic

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pathways involved in both primary and secondary metabolisms in addition to enzymes, reactions, metabolites, and genes involved for metabolism (Caspi et al. 2016). This database is freely accessible at http://metacyc.org/. In addition to being an online repository of metabolism, other scientific applications of MetaCyc include computational prediction and comparison of the metabolic pathways in sequenced genomes, information pertaining to metabolic engineering through enzyme database, and aid in metabolomics research (Caspi et al. 2016). BioCyc database is a collection of 5700 organism-specific Pathway/Genome Databases (PGDBs). The complete genome and predicted metabolic network of one organism, including metabolites, enzymes, reactions, metabolic pathways, predicted operons, transport systems, and pathway-hole fillers, can be visualized (Caspi et al. 2012). The Pathway Tools software in BioCyc predicts metabolic pathways using MetaCyc as a reference database (Caspi et al. 2016). Further, this database offers multiple tools for querying and analyzing PGDBs for the comparative analysis of gene expression and metabolomics data using Omics Viewers and genome browsers (Karp et al. 2015). This database is freely accessible at https://biocyc. org/. Recent developments of the database include addition of Gibbs free energy values for compounds and reactions, the addition of a tool for creating diagrams containing multiple-linked pathways, new search capabilities for searching genes based on organism’s phenotypes, sequence patterns, and a metabolite identifier translation service (Caspi et al. 2012). Recent PGDB additions to BioCyc database over the past 2 years include YeastCyc (Saccharomyces cerevisiae), EcoCyc (Escherichia coli) (Karp et al. 2014), HumanCyc (Homo sapiens) (Romero et al. 2004), and BsubCyc (Bacillus subtilis). MetaCyc and BioCyc are curated and maintained by researchers at SRI International. These databases aid in the selection of a particular microorganism for the biodegradation of xenobiotics.

Computational Biology Resources for Pathway Prediction PathPred PathPred (Moriya et al. 2010) is a knowledge-based web server that predicts possible enzyme-catalyzed reaction pathways of a query compound by information retrieved from the KEGG REACTION (Muto et al. 2013) and KEGG RPAIR databases (Oh et al. 2007). The KEGG REACTION database contains all the reactions taken from the KEGG metabolic pathways in addition to known enzymatic reactions retrieved from the IUBMB enzyme nomenclature. Similarly, KEGG RPAIR is a collection of biochemical structure transformation patterns (RDM patterns) for substrate-product pairs (reactant pairs) in KEGG REACTION (Muto et al. 2013). This web server provides plausible reactions and transformed compounds and displays all possible predicted reaction pathways in a tree-shaped graph (Moriya et al. 2010). An end product/compound can be queried in three different ways, either in the SMILES representation or the MDL mol file format or directly by the help of KEGG compound identifier (Moriya et al. 2010). The pathway prediction results are

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also linked to the genomic information. This web server displays new and alternative reactions irrespective of whether enzymes for these reactions are present or not. If the enzyme is nonexistent, users can use the enzyme tool (http://www.genome.jp/tools/ e-zyme/) to assign a possible EC number. PathPred web server can be freely accessed at http://www.genome.jp/tools/pathpred/. It is currently maintained by Kyoto University Bioinformatics Center.

Biochemical Network Integrated Computational Explorer Biochemical Network Integrated Computational Explorer (BNICE) is a computational approach for the development of novel pathways based on the reaction rules of the Enzyme Commission classification system (Finley et al. 2009). Initially, this approach checks for all possible metabolic pathways from a given user-specific target or starting molecule. Further, it filters out all possible pathways for thermodynamic feasibility based on the Gibbs free energies of the reaction and then selects a feasible unique thermodynamic pathway (Finley et al. 2009). These generated pathways can be further assessed using existing pathway analysis systems, such as thermodynamics-based flux balance analysis (FBA) that allows investigation of the overall effects of these novel pathways on metabolic network performance in host organisms (Soh and Hatzimanikatis 2010). FBA aids in the prediction of phenotypic changes, changes in bioenergetics of the system for metabolic engineering, effects of gene knockouts, and biodegradation of xenobiotics. Major applications of BNICE include discovery and analysis of novel metabolic pathways, evolutionary analysis between metabolic pathways of various organisms, and analysis of degradation pathways of xenobiotic compounds (Finley et al. 2009).

From Metabolite to Metabolite From Metabolite to Metabolite (FMM) predicts metabolic pathways form one metabolite to another metabolite among different species mainly based on the Kyoto Encyclopedia of Genes and Genomes (KEGG) database and other integrated biological databases (Chou et al. 2009). It is a freely available web server that can be accessed at http://FMM.mbc.nctu.edu.tw/. FMM generates combined metabolic pathways by combining the information available in KEGG PATHWAY and KEGG LIGAND databases (Muto et al. 2013). This web server also provides information regarding the corresponding enzymes, genes, and organisms for comparative analysis, in which metabolic pathways can be compared between numerous species. FMM is an efficient tool in the field of synthetic biology and metabolic engineering. For biodegradation studies, metabolic pathways (both local and global graphical maps) of xenobiotic compounds can be searched and visualized (Chou et al. 2009). FMM provides a highly effective way to elucidate the genes (from which species) should be cloned into the microorganisms for increased metabolite production based on comparative analysis.

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Metabolic Tinker A recently developed web tool, Metabolic tinker (McClymont and Soyer 2013), can be employed to design synthetic metabolic pathways between an end product (user-defined target) and source compounds. Metabolic tinker uses a tailored heuristic search approach to return a large number of thermodynamically feasible paths in the entire known metabolic universe owing to its broad search base. Metabolic tinker is a freely available web service which can be accessed at http:// osslab.ex.ac.uk/tinker.aspx. The same website also provides the source code, allowing it to be developed further or run on personal machines for specific applications. The program contains a directed graph known as Universal Reaction Network (URN), which represents the entire set of known compounds and reactions from the ChEBI and Rhea databases (McClymont and Soyer 2013). Edges and nodes on this graph represent reactions and metabolites, respectively. Metabolic tinker searches possible biochemical pathways between two compounds within this URN using standard search algorithms developed in computer science and graph theory. Existing tools do not incorporate the search for compound similarity and thermodynamic stability in searching natural pathways defined in a specific species context; hence, Metabolic tinker is an important tool for the prediction of metabolic pathways.

MetaRouter MetaRouter is a system that maintains heterogeneous information related to bioremediation and biodegradation pathways. The core of this system is a relational database where the information on reactions, chemical compounds, enzymes, and organisms is stored in an integrated framework (Pazos et al. 2005). The current set of data includes 740 compounds, 502 enzymes, 820 reactions, and 253 organisms (Pazos et al. 2005; Shah 2017). It is currently linked to major public databases such as ENZYME (http://enzyme.expasy.org/), UM-BBD (http://eawag-bbd.ethz. ch/), Swiss-Prot (http://www.uniprot.org/), and Protein Data Bank (http://www. rcsb.org/pdb/). Although, MetaRouter system does not provide information regarding kinetics or thermodynamics of the proposed pathways, it aids in assessing the environmental fate of harmful compounds and in designing biodegradative strategies for the same (Hatzimanikatis et al. 2005; Khan et al. 2013). It is freely available at http://pdg.cnb.uam.es/biodeg_net/MetaRouter. Other pathway prediction systems include Metabolic Route Explorer (MRE) (http://www.cbrc.kaust.edu.sa/mre/) (Kuwahara et al. 2016), Metabolic Route Search and Design (MRSD) (http://bioinfo.ustc.edu.cn/softwares/mrsd/) (Xia et al. 2011), eXTended Metabolic Space (XTMS) (http://xtms.issb.genopole.fr) (Carbonell et al. 2014), CarbonSearch (http://www.kavrakilab.org/atommetanet/) (Heath et al. 2010), GEM-Path (Campodonico et al. 2014), MetaRoute (http://

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www-bs.informatik.uni-tuebingen.de/Services/MetaRoute) (Blum and Kohlbacher 2008), and AGPathFinder (http://210.36.16.170:8080/AGPathFinderWeb/login.jsp) (Huang et al. 2017).

Protein Structure Prediction Protein structure prediction programs predict three-dimensional structures of the proteins that have not been experimentally elucidated and deposited in structural databases. Three-dimensional structure of an enzyme is essential for computational analysis such as molecular docking and molecular dynamic simulations. In some cases, when a three-dimensional structure of the enzyme of interest is unavailable in its native form, the 3D structure of the protein can be predicted by computational biology tools. There are three approaches for protein structure prediction, namely, homology modeling, fold recognition, and ab initio approaches.

Homology Modeling Homology modeling is predicting an atomic resolution model of the “target” protein (protein of interest) from its amino acid sequence and an experimental 3D structure of a related homologous protein of the “template” (Lee et al. 2009). It is also known as comparative modeling. It has been reported that the protein structures are highly conserved than protein sequences amongst homologues, but sequences falling below a 30% sequence identity can have different structures. Similarly, naturally occurring homologous proteins have similar protein structures, and evolutionarily related proteins have similar sequences. The quality of the predicted homology model is dependent on the quality of the sequence alignment and template structure. The major steps involved in homology modeling include target retrieval, template identification, target-template alignment, model prediction, loop modeling, side-chain refinement, model refinement, and validation (Lee et al. 2009).

Fold Recognition Fold recognition, also known as protein threading, is a method of structure prediction which involves proteins that have the same fold as proteins deposited in the structural databases but are not homologous in nature (Kelley et al. 2015). It differs from the homology modeling method of structure prediction as it is used for proteins which do not have their homologous protein structures deposited in any of the structural databases. The prediction is carried out by aligning each amino acid in the target sequence to a position in the template structure and evaluating how well the target fits the template. Homology modeling and protein threading are both template-based methods (Lee et al. 2009).

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Ab Initio Method Ab initio or de novo methods predict a 3D structure directly from the primary sequence, without the need for a template. De novo techniques are much more computationally intensive than template-based methods and usually limited to smaller proteins (90 . A surface would be considered hydrophilic when θ is 5 shows extremely polluted (Rastegari Mehr et al. 2017). The mutual pollution effect at different stations by different metals can be calculated using pollution load index. PLI is the geometric mean of the CF values for the n metals: PLI ¼

p ffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi n CF1  CF2  . . .  CFn

• A PLI value close to 1 indicates heavy metals load near the background level, while values >1 indicate pollution (Rastegari Mehr et al. 2017).

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Potential Ecological Risk Load Index (PERI) Potential ecological risk load index (PERI) is used to assess the degree of metal contamination in soils. The equations for PERI are as follows. Ci f ¼ C1 =Ci n , Cd ¼ Σn i¼1 Ci f , Ei r ¼ Ti r x Ci f , PER ¼ Σn i¼1 Ei r where Cif is the contamination factor of a single metal, Ci is the content of metal in samples, and Cin is the background value of the heavy metal. The background values (preindustrial samples of the study area) of Cr, Ni, Cu, As, Cd, and Pb in soils were 29, 32, 27, 6.5, 0.82, and 23 mg kg1, respectively. The sum of contamination factor (Cif) for all metals represents the integrated pollution degree (Cd). Eir is the potential ecological risk index and Tir is the biological toxic factor of an individual metal. The toxic-response factors for Cr, Ni, Cu, As, Cd, and Pb were 2, 6, 5, 10, 30, and 5, respectively (Islam et al. 2017). PER is the comprehensive potential ecological risk index, the sum of Eir.

Human Health Risk Assessment Human health risk assessment of metals in soils is related to hazard discrimination, exposure evaluation, and risk characterization. In consideration of potential toxicity and carcinogenicity of most elements, carcinogenic and noncarcinogenic effects of metals can be studied. Three exposure ways of metals to human body can be considered: direct oral ingestion of soil particles (ADDing), dermal absorption of elements in soils adhered to exposed skin (ADDdermal), and inhalation of resuspended soil particulates by nose or mouth (ADDinh). This methodology was developed by the United States Environmental Protection Agency for health risk assessment (USEPA 1989). The population is divided into two parts: adults and children. One population group that could potentially be more highly exposed to inhalation exposures at a site is children. The definition and reference values for all parameters are listed in Table 5. The equations for exposure dose are as follows: Direct oral ingestion of soil particles (ADDing):     ADDing mg kg1 d1 ¼ C mg kg1  ððIngR  EF  EDÞ=ðBW  ATÞÞ  106 Dermal absorption of elements in soils adhered to exposed skin (ADDdermal):

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Table 5 Definition and reference value of some parameters for health risk assessment of metal in soils (Sun et al. 2017) Symbol (units) IngR (mg day1) InhR (mg day1) EF (d a1) ED (a) BW (kg) SA (cm2)

Parameter Soil ingestion rate Soil inhalation rate Exposure frequency Exposure duration Body weight Exposed skin surface area Adherence factor AF

Value 200 (children); 100 (adults) 7.6 (children); 20 (adults) 180 6 year (children); 24 year (adults) 15 (children); 70 (adults) 1150 (children); 2145 (adults)

0.001

AT (day)

Dermal absorption factor Average time

PEF (m3 kg1)

Emission factor

SL mg cm2 day1) ABS (unitless)

0.2 (children); 0.07 (adults)

ED  365 (noncarcinogenic); 70  365 (carcinogenic) 1.36  109

    ADDinh mg kg1 d1 ¼ C mg kg1  ððIngR  EF  EDÞ=ðPEF  AT  BWÞÞ Inhalation of resuspended soil particulates by nose or mouth (ADDinh):     ADDdermal mg kg1 d1 ¼ C mg kg1  ððSA  SL  ABS  EF  EDÞ=ðBW  ATÞÞ  106 where ADDing, ADDdermal, and ADDinh are the average daily intake from soil ingestion, dermal, and inhalation absorption, respectively (mg kg1 d1) and C is the concentration of metal in soil (mg kg1). The doses calculated for each element and exposure pathway are subsequently divided by the toxicity threshold value which is referred to as the reference dose (RfD, mg kg1 d1) of a specific chemical to yield a noncarcinogenic hazard quotient (HQ), whereas for carcinogens, the dose is multiplied by the corresponding slope factor (SF) to produce a level of cancer risk (Sun et al. 2017). To assess the overall potential for noncarcinogenic effects posed by all exposure pathways, a hazard index (HI) which is the total noncarcinogenic risk of exposure to a variety of pollutants has been employed as described by Sun et al. (2017). If HQ < 1 or HI < 1, the exposed individual is unlikely to experience obvious adverse health effect. On the contrary, if HQ > 1 or HI > 1, there is a chance that noncarcinogenic effect may occur with a probability which tends to increase as HI increases (Sun et al. 2017). HI ¼

X

HQ ¼

X

ADDi =RfDi

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Generally, carcinogenic risk is regarded as the probability of an individual developing any type of cancer in the whole lifetime due to exposure to carcinogenic hazards (Sun et al. 2017). Similarly, the aggregate carcinogenic risk (Risk) is calculated by summing the individual cancer risk across all exposure pathways: Risk ¼ ADD  SF

Organic Pollutants Persistent organic pollutants (POPs) are, by definition, organic compounds highly resistant to photolytic, biological, and chemical degradation. These compounds are also moderately volatile, allowing their long-distances atmospheric transport. POPs compounds are very poorly dissolved in water, and very good in fats, and easily pass through phospholipid structures of biological membranes, after which they are deposited in fatty tissue and other high lipid-containing tissues. All these properties ensure the widespread distribution of these compounds in the environment, even in those regions where they have never been used. Because of this, POPs chemicals are characterized as a ubiquitous class of compounds. Although there are natural sources of organochlorine compounds, most POPs chemicals originate from anthropogenic sources, which are related to the production, application, and disposal of these chemicals. The general population is most often exposed to POPs chemicals through food. Often these substances are present in the workplace and in the environment. Exposure to POPs compounds can cause harmful effects on human health. These effects most often include neurological disorders, disorders of liver function, reproductive system, behavioral, immune, and endocrine disorders, as well as carcinogenic effects. Due to the potential hazard posed by POPs properties, these days, special attention is paid to the system of control and management of toxic substances and waste. There are international conventions, regulations, and protocols, which regulate certain details related directly or indirectly to POPs chemicals, that have been adopted and applied. One of them is Stockholm Convention (Stockholm Convention 2001) for POPs adopted in 2001 in Stockholm. According to the Convention, there are 12 initial and 16 new POPs compounds which can be placed in two categories: intentionally and unintentionally POPs (Fig. 1).

Intentionally POPs Group of intentionally POPs includes agrochemicals like organochlorine pesticides (OCP) and industrial chemicals like PCBs, some polybrominated diphenyl ethers (PBDE) etc.

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POPs

Unintentionally

Intentionally

Agrochemicals

Industrial chemicals

PAHs

Furans

Dioxins

Fig. 1 POPs categories

Agrochemicals Pesticides POPs pesticides originate almost entirely from anthropogenic sources and are associated largely with the manufacture, use, and disposition of certain organic chemicals. Despite the efforts to use biodegradable and less persistent pesticides in developing countries, organochlorine pesticides like HCH, DDT, heptachlor, and chlordane are still being used. Due to their long-distance transport capability, this is not only a problem of countries that produce and use them but also is a problem for developed countries and for areas where they have never been used such as the Antarctic (Kang et al. 2012). The presence of pesticides on the Antarctic has been confirmed already 50 years ago. A group of scientists found residues of pesticides in vertebrates: fish species, Adelie penguins, Weddell seals, and Skuas (George and Frear 1966). Despite the prohibition of use, some OCPs are still present on these nontarget areas. As part of the 2007–2008 International Trans Antarctic Scientific Expedition (ITASE) program, surface snow samples were obtained from the coastal areas to the interior regions of East Antarctica. The samples were analyzed for OCPs and trace inorganic elements. Of the 22 OCPs, α-hexachlorocyclohexane (HCH), γ-HCH, and hexachlorobenzene (HCB) were frequently detected in the snow with concentration ranges of 17.5–83.2, 33–137, and ND–182 pg L1, respectively. These results indicate that the OCPs were subjected to long-range atmospheric transport and were deposited in the surface snow. Hexachlorocyclohexane (HCH) is one of the polyhalogenated organic compounds: α-HCH, β-HCH, or γ-HCH (lindane). Contrary to lindane, alpha-HCH and beta-HCH are not intentionally produced. They are produced as a part of technical HCH which is used as an insecticide or as an intermediate chemical in the manufacturing of lindane. Linadane was used as an insecticide on fruit, vegetables, and forest crops, and animals and animal premises. Although technical-grade HCH

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is no longer used as an insecticide in many countries, α-, β-, and γ-HCH have been found in the soil and surface water at hazardous waste sites because they persist in the environment. In humans, breathing toxic amounts of γ-HCH and/or α-, β-HCH can result in blood disorders, dizziness, headaches, and possible changes in the levels of sex hormones in the blood (ATSDR 2005a). HCH (all isomers) as possibly carcinogenic to humans. The EPA has determined that there is suggestive evidence that lindane (γ-HCH) is carcinogenic, but the evidence is not sufficient to assess its human carcinogenic potential. The EPA has additionally classified technical HCH and α-HCH as probable human carcinogens, β-HCH as a possible human carcinogen (US EPA 2005). Chlordane is a more highly chlorinated analogue of heptachlor. The technical mixture contains chlordane, heptachlor, nonachlor, and related componds. Chlordane is a man-made chemical and popular pesticide that was used in the United States from 1948 to 1988. Due to concerns about environmental damage and human health risks, the US Environmental Protection Agency banned the use of chlordane in 1983, except for controlling termites (US EPA 1997). It was banned for that use, and all uses, in 1988 but still its effect is felt across the world. Chlordane is readily absorbed in both animals and man through the skin, ingestion, and probably also inhalation. Some accumulation occurs in the body on repeated exposure – mainly in adipose tissue. Elimination from the body is fairly slow. Food is the major source of exposure of the general population to chlordane. Significant exposure to chlordane can occur in buildings where chlordane has been used for termite or other insect control. Chlordane is rapidly metabolized in organisms into oxychlordane and γ-chlordane or into impurities such as trans-nonachlor or cis-nonachlor. These are metabolic products that persist in the tissue of fish (Joseph et al. 1987), birds (Fiska et al. 2001), and mammals and that are found in breast milk. In the 1980s, when there was a current problem with chlordane, were done a number of studies examining chlordane metabolites in breast milk. Almost all studies have confirmed the presence of chlordane in the mother’s milk of women living in areas where chlordane was used as an insecticide in agriculture and for the home destruction of termites (Solomon and Weiss 2002). Chlordane can affect the digestive and nervous system. High exposure can cause convulsions and death. Additionally, exposure to herbicides and insecticides is especially harmful to the thyroid; chlordane is one of the most toxic chemicals in this regard (Goldner et al. 2011). DDT was widely used during World War II to protect soldiers and civilians from malaria, typhus, and other diseases spread by insects. After the war, DDT continued to be used to control disease, and it was sprayed on a variety of agricultural crops, especially cotton. DDT continues to be applied against mosquitoes in several countries to control malaria. The best-known toxic effect of DDT is described as Silent Spring – book which documented the death of birds resulting from the aerial spraying of DDT to kill mosquitoes. Researchers discovered that earthworms were accumulating the persistent pesticide and that the birds eating them were being poisoned. These evidences led to a nationwide ban on DDT for agricultural uses and inspired an environmental movement that led to the creation of the US Environmental Protection Agency (Paull 2013).

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Heptachlor is a nonsystemic stomach and contact insecticide, used primarily against soil insects and termites. It was used extensively in the 1960s and 1970s in household and agricultural settings. Studies link heptachlor to cancer, endocrine disruption, and developmental toxicity. Heptachlor is readily converted to more potent heptachlor epoxide (HE) once it enters the environment or the body. A group of scientists (Cassidy et al. 2005) from USA have worked on determination of the levels of HE, oxychlordane (OC), and dichlorodiphenyldichloroethylene (DDE) in adipose tissue within breast biopsies in a series of 34 women evaluated for breast abnormality. They have shown that HE increases oxidant levels that damage DNA and cell membranes by a 17b–E2 receptor-mediated process and they proposed that HE-induced DNA damage may contribute to breast cancer in exposed women. Heptachlor began to be phased out in the late 1970s but its use was not banned until. Its only legal use today is to kill fire ants in power transformers and in underground cable television and telephone cable boxes (ATSDR 2005b).

Industrial Chemicals Polychlorinated Biphenyls Polychlorinated biphenyls (PCBs) are persistent organic pollutants (POP) that once released into the environment remain there a long time (many years) because they are resistant to photolytic, biological and chemical degradation. PCBs are found in the ground, groundwater, and surface water, and mostly in the air. Because of their nonsolubility in water, they accumulate in fatty tissues of living organisms. In large concentrations are found in the higher levels of the food chain (birds, fish, mammals, people). As a consequence of their bioaccumulative characteristics, they often occur in organisms of humans and animals who live far from the source of PCBs. These organic compounds commonly enter into living organisms through the food and the consumption of fish, poultry meat, and dairy products. Often comes to occupational exposure to PCBs, and there are also accidental poisonings by relatively high doses of these compounds. PCBs are extremely dangerous for workers who work on maintenance and repair of transformers because PCB concentrations there are much larger compared to other areas (Stojić et al. 2014), such as the air in buildings that have electrical devices containing PCB or the air in the external environment, including the air in hazardous and toxic waste landfills. Exposure to individuals to PCBs can lead to certain harmful effects to the health of humans and animals. These effects usually include neurological disorders, disorders of liver function and reproductive system, change in behavior, disorder of the immune and endocrine systems, and carcinogenic effects. Especially sensitive to the presence of PCBs are babies and fetals that are in the phase of development since these compounds can pass through the placenta and are excreted in breast milk (Shea et al. 2007). Fortunately, application and production of PCBs has been banned for several decades, and so their levels in the environment are continuously decreasing. PCBs have been released in the environment only as a result of human activity. Their atmospheric transport depends on the number of Cl atoms. Biphenyls with 0–1

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chlorine atom remain in the atmosphere, those with 1–4 chlorines gradually migrate toward polar latitudes in a series of volatilization/deposition cycles, those with 4–8 chlorines remain in midlatitudes, and those with 8–9 chlorines remain close to the source of contamination (Wania and Mackay 1996). PCBs in the air usually come from evaporation from the ground or water surface. Once they come in the environment, further they can be transformed and degraded depending on their structure. The different structure of PCBs and therefore the different degree of toxicity imposed the development of risk assessment regulations. The toxicity of any PCB congener is dependent upon both the number and position of each chlorines on the biphenyl ring. PCB congeners with a higher number of Cl atoms also have higher toxicity. However, PCBs with chlorines in the ortho positions of each ring (positions 2, 20 , 6, and 60 ) are less toxic than non-ortho or mono-ortho PCBs. Non-ortho PCBs, also known as the coplanar PCBs, bind the aryl hydrocarbon receptor (AhR) and are capable of producing dioxin-like effects within biological systems. In accordance with that, two groups of PCBs are separately observed: dioxin-like and nondioxinlike PCBs. There are 12 dioxin-like PCBs with similar structure as TCDD (2,3,7,8tetrachlorodibenzo-p-dioxin) (Fig. 2). Indicator of toxicity for these compounds is toxic equivalency factor (TEF). The reference congener is the most toxic TCDD which per definition has a TEF of 1 (Birnbaum et al. 2006). TEF takes into account chemical structure and behavior of chemical compound. For each chemical, the model uses comparative measures from individual toxicity assays, known as relative effect potency (REP), to assign a single Fig. 2 Structural formula of TCDD

Table 6 TEF values for PCB congeners

Congeners Non-ortho PCB PCB 77 PCB 81 PCB 126 PCB 169 Mono-ortho PCB PCB 105 PCB 114 PCB 118 PCB 123 PCB 156 PCB 157 PCB 167 PCB 189

TEF values 0.0001 0.0003 0.1 0.03 0.00003 0.00003 0.00003 0.00003 0.00003 0.00003 0.00003 0.00003

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scaling factor known as the TEF. TEFs are determined using a database of REPs that meet WHO (World Health Organization) established criteria, using different biological models or endpoints and are considered estimates with an order of magnitude of uncertainty (Van den Berg et al. 1998). By applying the appropriate TEF to each dioxin congener, the expert toxicologist can calculate a toxic equivalent (TEQ) which expresses the total toxicity of a mixture of dioxins. TEF values for PCB congeners are given in Table 6 (Van den Berg et al. 2006).

Polybrominated Diphenyl Ethers (PBDE) Tetrabromodiphenyl ether, pentabromodiphenyl ether, hexabromodiphenyl ether, and heptabromodiphenyl ether are PBDEs added to the list of Stockholm Convention for POPs in 2009. PBDE are industrial chemicals that are used more than 40 years mostly as flame retardants. In the environment are usually released during their production process and use as well as during the process of servicing appliances containing PBDE, their improper recycling and disposal in landfills. Another source of PBDEs can be the old, discarded furniture that falls apart under the influence of atmospheric conditions (sun, wind, rain, extreme temperatures) in tiny particles that become part of the soil, air, and groundwater. Human uptake is thought to be through inhalation, dermal absorption, and consumption of contaminated food. These chemicals are among the most toxic chemicals ever synthesized. There are evidences that they cause cancer (Wenning 2002), and also they cause damage to the central and peripheral nervous systems (Banasik and Suchecka 2011), diseases of the immune system (Fernie et al. 2005), reproductive function disorders (Kodavanti et al. 2010), functional disorders of the endocrine system (Goldner et al. 2011), and most importantly they affect neurological development and growth (Dingemans et al. 2011).

Unintentionally POPs They are an unwanted product of combustion or some chemical process. Unintentionally, POPs are divided into three groups: polycyclic aromatic hydrocarbons (PAHs), dioxins, and furans.

Polycyclic Aromatic Hydrocarbons (PAHs) PAH (Fig. 3) are organic substances which are widespread in the environment. The EPA (US EPA 2014) listed 16 PAHs on a list of priority pollutants since they are considered either possible or probable human carcinogens. Fig. 3 Structural formula of PAHs (benzo[a]pyrene)

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The most frequent sources are the anthropogenic activities such as: production of coal and coal tar, asphalt, aluminum, petroleum refineries, and motor exhaust. But, there are also natural sources of PAHs like open burning, volcanic activities, and natural losses or seepage of petroleum or coal deposits. PAHs from the air are disintegrated by the action of sunlight and from the water and soil by the action of microorganisms over a period of 1 week to a month. However, some of them remain for years in the soil, underground waters, and sediments of rivers, lakes, and seas. Investigations on experimental animals have shown that PAHs cause mutagenic and carcinogenic changes and therefore assume that they are potentially mutagenic and carcinogenic to humans. Based on the available evidence, both the International Agency for Research on Cancer (IARC 1987) and the US EPA (US EPA 1984) classified a number of PAHs as carcinogenic to animals and some PAH-rich mixtures as carcinogenic to humans. PAHs are associated with lung cancer, so smokers who inhale smoke containing more PAHs are particularly vulnerable. A study conducted in the United States (McCarty et al. 2009) showed that daily exposure to PAHs at smokers is 30 times higher than for others.

Dioxins and Furans Dioxins are the most widespread toxic chemicals in the environment, which are the result of technological activities of man. Dioxins are not water-soluble but are highly soluble in fats and tend to accumulate in higher animal species, including humans. They are volatile and easily transmitted over long distances, and so they are transnational environmental pollutants belonging to a group of persistent organic pollutants. The name dioxin refers to a large group of 210 different dioxin compounds and furans, with 17 of them being extremely toxic. The most prominent and most toxic dioxin compound is 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). They are unintentional nus-products of various processes (industry – manufacture of other chemicals, such as pesticides and dyes, steel production, paper bleaching, household – wood burning, natural disasters, and incidents – forest fires, volcanic eruptions, and uncontrolled incineration of hazardous waste). Therefore, dioxins and furans are most commonly found in the air. One of the most famous poisoning with dioxins is the use of a chemical substance, the so-called “Agent Orange,” which the US military used as a herbicide for the destruction of the rainforest in which the enemy was hiding from 1961 to 1971 in the Vietnam War. This herbicide is a mixture of 2,4-dichlorophenoxyacetic acid (2,4-D) and 2,4,5-trichlorophenoxyacetic acid (2,4,5-T), which contains the most toxic dioxin TCDD, which is the cause of poisoning in the Vietnamese population and war veterans of the United States. Dioxins and furans are carcinogenic and can cause problems in reproduction and development (Yonemoto 2000). They destroy the immune system (Radenkova-Saeva 2009) and interfere with the hormonal system. The biggest problem for humans is that they are very slowly decomposed and accumulated in the body; therefore, this is why chronic exposure to dioxins is particularly dangerous for human health.

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Analytical Methods The choice of analytical methodology depends largely on the sample matrix (sample type) and the chemical structure of the target analytes. Collection of a sample and its preparation for analysis is a very important step in the analytical procedure. Most POPs in water are sampled by the use of a conventional bottle sampling technique, whereby a precleaned sampling bottle is immersed into a target water body, usually with the help of a water pump. Passive dosimetry technique is also used, with which analytes can be isolated and preconcentrated at the same time. A review of some parameters characterizing some of the passive dosimeters from water and most commonly used over the last 15 years sample preparation techniques has been published by a group of authors (Tankiewicz et al. 2011). According to them, the most commonly used passive dosimeters are polar organic chemical integrative sampler, trimethylpentane-containing passive sampler, and membrane-enclosed sorptive coating. With the use of passive dosimeters, there are fewer steps in the analytical procedure, which means results are more reliable and reproducible. After sampling, the sample should be isolated and preconcentrated, which is also an important step in the analysis due to their low concentration. Liquid–liquid extraction (LLE) and solid-phase extraction (SPE) are the oldest and at the same time the most frequently applied sample-handling techniques in the determination of POPs in water. LLE is perhaps the simplest but not the best technique. LLE uses a large amount of toxic solvents like dichloromethane, mixtures of petroleum ether and dichloromethane or hexane and dichloromethane. In order to avoid this disadvantage, some changes were made as it is liquid–liquid microextraction (LLME), dispersive liquid–liquid microextraction (DLLME) (Berijania et al. 2006), and ionic liquids DLLME (IL-DLLME) (Lijun et al. 2009; Diao et al. 2016). These techniques are rapid, simple, and eliminate the need to cleanup of the extract and use only few microliters of solvents. Despite the advantages of these microextraction techniques, SPE is still widely accepted as the best technique for isolating POPs in water samples, because it is fast, precise, cheap, and uses a small amount of solvents (Pucarevic et al. 2017). For soil samples, traditional sample preparation methods (liquid-liquid extraction, Soxhlet extraction, etc.) are still used. But they are laborious, time consuming, expensive, and require large amounts of organic solvents and usually involve many steps, leading to loss of some analyte quantity. As a result, modern sample preparation procedures, such as accelerated solvent extraction (ASE) have been developed to overcome the drawbacks of the traditional approaches (Đurović and Đorđević 2011). ASE technique is equivalent to US EPA Methods (US EPA 3540, 3550, 8150) for the extraction of organochlorine pesticides (OCPs), organophosphorous pesticides (OPPs), semivolatiles or base neutral acids (BNAs), chlorinated herbicides, polycyclic aromatic hydrocarbons (PAHs), and polychlorinated biphenyls (PCBs). The accelerated solvent extraction technique complies with US EPA Method (US EPA 3545A) for these compounds and is an extraction technique that

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significantly streamlines sample preparation. This technique uses extraction solvents at elevated temperatures and pressures to increase the kinetics of the extraction process. The high pressure allows the solvent to be used above its boiling point, keeping it in a liquid state, and thus decreases the amount of time and solvent required to extract the desired analyte from the sample matrix. Accelerated solvent extraction technique replaces extraction techniques such as Soxhlet, sonication, and wrist-shaker with equivalent or better results. Other, but not less important, techniques that used are: supercritical fluid extraction (SFE), microwave-assisted extraction (MAE), solid-phase extraction (SPE), solid-phase microextraction (SPME), matrix solid-phase dispersion (MSPD) extraction, and QuEChERS (quick, easy, cheap, effective, rugged, and safe). It should be indicated that SFE, ASE, and MAE are instrumental techniques and often use SPE and SPME for purification of obtained extracts, and also its concentration in case of SPME. The separation and detection of most POPs are usually accomplished by GC coupled with several types of detectors, such as electron capture detector (ECD) and MS detector or high pressure liquid chromatography (HPLC) (Pandit et al. 2002). Identification based only on chromatographic analysis (retention time) without the use of spectrometric detection is not suitable as confirmatory method, so MS detection has found to be indispensable for high sensitivity and unambiguous detection, confirmation, and determination of such residues in different matrices (EU Commission Decision 2002/657, 2002). In summary, with the improvement of the sensitivity of analytical instruments, the POPs list is increasing, sample volume/mass used for extraction/ analysis and the volume of solvents used is decreasing. All of this contributes to the protection of the environment and to a better quality of life.

Remediation In many courtiers around the world, including the most developed countries, there are a large number of contaminated sites on which confirmed the presence of hazardous and harmful substances in concentrations that pose a significant risk to human health and the environment. While it is difficult to quantify, it has been estimated that the number of contaminated sites, for example, in Australia is approximately 100.000 with an estimated remediation cost of $5–8 billion. These values are small compared to the number of contaminated sites and estimated remediation costs in Europe and USA (Naidu et al. 2010). Soil contamination requiring cleanup is present at approximately 250,000 sites in the EEA member countries, according to the recent estimates. And this number is expected to grow. Potentially polluting activities are estimated to have occurred at nearly three million sites and investigation is needed to establish whether remediation is required. If current investigation trends continue, the number of sites needing remediation will increase by 50% by 2025 (EEA 2007). Soil contamination is the result of mixing solid and liquid hazardous substances with the substances that are naturally present in soil. Landfills are the simplest and

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most commonly used method of disposing municipal solid waste (MSW). Landfills pose a potential threat, because they can endanger the quality of surrounding land, surface, and groundwater. Landfill leachate is one of the main sources of groundwater and surface water pollution if it is not properly collected, treated, and safely disposed (Abd El-Salam and Abu-Zuid 2015). Leachate usually contains pollutants and often toxic substances from waste. Many locations contaminated with certain pollutants require remediation. There are different definitions of remediation which differ from country to country but in general include the implementation of appropriate measures to prevent further pollution of all environmental factors and restoring contaminated sites in a condition suitable for use, with no negative impact on the environment and human health. According to the US EPA (United States Environmental Protection Agency), the procedure of application specific remediation technology certainly must precede the identification of present pollutants and assessment of their impact on human health and the environment (US EPA 2008). After performing the process of identification and assessment of the effects of pollutants, it is followed by a method of treatment or removal of the contaminated medium. For the implementation of all these procedures, the US EPA uses a wide range of technologies, in order to present pollutants or completely removed from the site or be treated to the point where no longer pose a threat to human health and the environment (US EPA 2000). In this context, remediation technologies are referred to as technologies for the treatment or complete removal of pollutants from contaminated media. In Part 2A of the Environmental Protection Act, which came into force in 2000 in England (EA 2009), remediation is defined as a process involving the breakdown of links between pollutants and receptors, as well as disabling the routes of transporting contaminants to the receptor. Often, this involves the removal of pollutants from the location but also preventing the spread of polluting substances to the receptor, as well as changes are made in the receptor (e.g., to change the purpose of land, to limit access to challenge, etc.). In the last two cases, it is not necessary to change the level of contamination in the soil (Defra 2008). Changes in legislation on waste in the European Union at the beginning of the twenty-first century have a significant impact on the classification and methods of waste treatment but also on the remediation of contaminated soil. The most significant changes are related to the revision of the European Waste Catalog and the introduction of the Landfill Directive (Bone et al. 2004). The European Waste Catalog from 2002 identifies construction waste and demolition waste, including excavated soil from contaminated sites, as a special waste class (Commission Regulation (EC) No 574/ 2004). The Landfill Directive primarily plays an important role in preventing the formation of contaminated sites from waste disposal, because it (Council Directive 1999): – Prohibits the disposal of certain waste types (liquid, combustible, explosive, infectious hospital or slaughter waste, old tires, and other types of waste that do not meet the criteria set out in Annex II) – Reducing biodegradable waste disposal

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– Prohibits the deposit of untreated waste and the joint disposal of inert, hazardous, and municipal waste – Introduces the classification of landfills according to the type of waste that is deposited on it, etc. In the Netherlands, for example, the management of contaminated sites policy is based on two key laws, the Soil Protection Act (Wesselink et al. 2006) and the Environmental Protection Act (Soil Remediation Circular 2009). Among other things, the Soil Protection Act is prescribed to take a soil remediation in such a manner that, after its implementation, soil can be used for the above specified purpose, whereby the risk for people, animals, and plants from exposure to contamination must be minimized. The law also stipulates that remediation must be performed in such a way that as much as possible reduce the possibility of the spread of contamination in the surrounding area as well as to reduce the necessity of taking accompanying measures and the limited use of land after treatment (Soil Remediation Circular 2009). There are a number of technologies for the treatment of contaminated soil. In selecting appropriate technology for the treatment or elimination of pollutants from the soil should be taken into account a number of factors. The most important are the type of pollutants at the site, the level or degree of contamination, the possible consequences for the living world, the surface of the contaminated medium, the location, etc. One of the most important factors is also the level of financial investment required for the implementation of certain technologies, which also should not be decisive for the choice of appropriate technology, especially because any delay and inadequate choice can have significant and lasting effects (Jaksic and Ilic 2000).

Soil Pollution Caused by the Disposal of Solid Waste to Unsanitary Landfills Soil pollution is defined as the presence of contaminants in soil, in high enough concentrations to pose a risk to human health and/or the ecosystem. In the case of contaminants which occur naturally in soil, even when their levels are not high enough to pose a risk, soil pollution is still said to occur if the levels of the contaminants in soil exceed the levels that should naturally be present (Regulation 2010). According to EU regulations (Commission of the European Communities 2006), soil contamination is discussed when in the soil identifies the presence of pollutants above the prescribed level, causing deterioration or loss of one or more soil functions. The most common causes of soil contamination are associated with human activities, which result in the emission of artificial chemicals into naturally occurring soil components, which disturbs the natural balance and causes adverse effects on human health and the environment. This type of contamination most often derived from the leaching of hazardous and harmful substances from the above-ground tanks, the use of pesticides, the penetration of contaminated surface water in the lower layers of the soil and groundwater, leaching of pollutants from the unsanitary landfills, etc. The most

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common causes of soil contamination are petroleum hydrocarbons, solvents, heavy metals, etc. The occurrence of these phenomena is correlated with the degree of industrialization and the intensity of use of chemicals. The link between soil contamination and waste management is more than obvious. Inadequate waste management leads to the formation of a large number of contaminated sites. Advanced waste management systems which include the recycling of waste in construction products or fertilizers can also be positively or negatively affect the quality of the soil. Modern landfills, having elements of the sanitary protection, are designed to prevent contamination of the surrounding soil, surface water, and groundwater (Van Camp et al. 2004). In case of unsanitary disposal of municipal and other waste streams, there is an uncontrolled distribution of pollutants from the disposed waste, which indicates the possibility of contamination of the soil and other environmental media. Lack of the protective synthetic liner and system for collection of leachate at landfills for inert, nonhazardous or hazardous waste is a problem due to the migration of pollutants into surrounding soil, underground, and surface waters (Ismail et al. 2003).

Sources of Soil Pollution In the literature, there are many different classifications of soil pollution sources. Caliman et al. (2011) claim that the following activities mostly lead to pollution of soil and groundwater: – – – – –

Leaching from unsanitary landfills. Uncontrolled dumps. Accidental discharge of chemical and waste materials. Inadequate storage of liquid waste. Setting up a system for collection of sewage wastewater in hydrologically and geologically inappropriate locations. – Inadequate application of fertilizers and pesticides in agriculture, etc. In terms of soil contamination, the major problem is technology that gives a large amount of waste. It is estimated that around 25 billion tonnes of raw materials are processed annually in the world, from 1 to 1.5 billion tonnes in finished products, while the rest is waste that requires adequate treatment and disposal (Kastori et al. 2006). The European Environment Agency (EEA) has estimated that about three million sites in the EU countries potentially threatened by anthropogenic activities that take place on them, of which 250,000 locations requires cleaning (EEA 2007). Considering the current trends, the EEA estimated that the number of contaminated sites would increase by 50% by 2025 (Caliman et al. 2011). On the basis of the conducted investigations on contaminated sites in Europe, the most important sources of contamination are industrial activities, the treatment and disposal of waste (Fig. 4).

Pollutants in Soil In order to successfully protect soil from pollution, it is necessary to know the sources of pollution but also the quantity and characteristics of pollutants and their

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Military Mining 0.9% 1.4%

Others 8.2%

Transport spills on land 2.1% Power plants 3.9% Industrial production and commercial service 41.4%

Storage 5.4%

Industrial waste treatment and disposal 7.3%

Oil industry 7.3%

Municipal waste treatment and disposal 15.2%

Fig. 4 Overview of activities causing soil contamination in Europe (EEA 2007)

harmful effects. The number and type of pollutants are unlimited, as they constantly change and complement depending on the applied technologies, the degree of urbanization, etc. (Kastori et al. 2006). Generally, the pollutants or contaminants are substances that can damage or pollution of the environment, regardless of whether they are naturally present in the environment or have been released from an industrial process or other human activities. Kostic (2007) under the pollutant means any physical, chemical, biological or radioactive, gaseous, liquid or solid matter or substance that reduces the natural quality of water, soil, or air. Although there are many classifications of pollutants, which are used in the examination of the contaminated areas, in the broadest sense they can be divided into organic and inorganic. Gradel (1978) argues that organic substances constitute a large group of over 1600 chemicals of natural and anthropogenic origin that are present in a natural and polluted environment. Among the most important polluting substances of organic origin are: products from the oil refining process, such as petroleum hydrocarbons, organochlorine compounds, such as pesticides or polychlorinated biphenyls, dioxins, and furans, etc. In the waste, organic pollutants may occur in the form of soluble compounds or as complex mixtures with other compounds, including inorganic pollutants. Many organic pollutants possess hazardous characteristics. US EPA (1992) defined hazardous substances as having at least one of the following characteristics of hazardous substances: toxicity, corrosivity, self-inflammability, and chemical

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reactivity. Kastori et al. (2006) define hazardous substances as toxic substances (carcinogenic, teratogenic, and mutagenic), persistent or susceptible to bioaccumulation. One of the most important groups of hazardous pollutants are persistent organic compounds (POPs) that are not changing in the environment or their disappearance is so slow that they are accumulating in parts of the ecosystem. Most of the polluting substances of organic origin are toxic at very low concentrations. Typical inorganic pollutants in soil are heavy metals and metalloids. In addition to theme, other cations and anions that pose a significant risk to the receptors in the environment can be found in the soil. Heavy metals in the environment may be naturally present or may be of anthropogenic origin. Unlike metals that come into the environment by human action, natural metals in the environment are present at relatively low concentrations. In recent years, the number of metal anthropogenic sources increases which causes a significant increase in their concentration in the environment. The most important sources of metal emissions into the environment are dumps, transportation means, mines and molten metals, organic mineral fertilizers, sludge from a wastewater treatment plant, etc. (Granero and Domingo 2002). Heavy metals represent pollutants that are unchangeable, nonbiodegradable, and persistent in the soil. Although the soil has a natural capacity to mitigate bioavailability and the movement of metals by means of different mechanisms (precipitation, adsorption, and redox reactions), when heavy metal concentrations become so large that they overcome the natural regulatory capacity of the soil, their mobilization results in contamination of agricultural products and groundwater. In these cases, it is necessary to take all necessary measures for the remediation of contaminated soil (Shi et al. 2009). Knowing the physical and chemical characteristics of pollutants present in the soil allows predicting their mobility and the success of the remediation process. According to the literature data (Bone et al. 2004), chemical interactions of pollutants with soil primarily depend on the phase in which the pollutant is located (solid, liquid or gas), its solubility, soil mineralogy, pH values, and the amount of organic matter present in the soil. Williams et al. (2002) point out that naturally occurring organic matter in the soil most often react and form complex compounds with metals. Depending on the solubility of newly formed complex compounds, there is a decrease or increase in the mobility of metals in the soil. Thus, monitoring of the reactions between soil and pollutants and the parameters of the environment in which they occur is very important for determining the fate of pollutants and the risks that their presence in the soil represents for human health and the environment. The main phenomena that describe the reaction between the pollutants, soil, and water in the pores of the soil are: – – – – –

Sorption (physical or chemical adsorption). Oxidation-reduction reactions. precipitation. Hydrolysis. Biological degradation.

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The Criteria for the Classification of Contaminated Soil Soil classification depending on the presence of pollutants is very important because it determines the extent to which the soil is contaminated, whether it poses a risk to human health and the environment, or whether its remediation is required. In many countries around the world, different soil classification is applied, depending on the degree of presence of pollutants. In USA, each state shall prescribe permissible content of pollutants in the soil and accordingly determines the classification of contaminated soils. In USA, soil is generally categorized as follows (Kurisko et al. 1998): – Soils containing hazardous wastes – require “on-site” remediation or “off-site” management of soil as hazardous waste. – Nonhazardous soils. – Soils containing pollutants below the values that require cleaning. In the European Union, the soil protection framework is the COM (2006) 232 final amending Directive 2004/35/EC (Environmental Liability Directive, ELD) which establishes a framework to prevent and remedy environmental damage. Although all member states are obliged to apply the provisions of Directive COM (2006) 232 final, it did not define the limit values for remediation of hazardous and harmful substances in soil and each member state defines the aforementioned criteria. In the Netherlands, there are precisely defined maximum permissible concentrations for the content of pollutants (metals, other inorganic substances, aromatic compounds, polycyclic aromatic hydrocarbons, chlorinated hydrocarbons, etc.) in soil and groundwater. In this sense, the contaminated soil is classified as soil where it is not required or requires remediation. Soil remediation projects are performed if the average concentration of any hazardous or harmful substances of more than 25 m3 volume area exceeds the prescribed value (Soil Remediation Circular 2009). In many countries of the world, there are clearly defined criteria for classifying contaminated soil as waste. European Waste Catalogue, adopted in 1994, introduced a new method of classification and treatment of waste, including the remediation of contaminated soil. The legal basis for the European Waste Catalog makes Directive 75/442/EEC on waste, as amended by Directive 26/12/EC. According to the requirements of Directive 91/689/ EC on hazardous waste, the Council of Europe has made Decision 94/904/EC defining a list of wastes from the Waste Catalogue that have hazardous characteristics, known as the list of hazardous wastes (Council Decision 1994). Contaminated soil is not identified on the hazardous waste list but is categorized as special waste, unless it has one of the hazardous characteristics of the H list (from H3 to H8), which presents a list of characteristics of waste that make it hazardous. The European Waste Catalog has been periodically revised and updated. In the latest version of the European Waste Catalog, there are several waste groups and a separate group of wastes consists of group 17 which includes construction and demolition waste (including excavated soil from contaminated sites). The Catalog within the group 17 as a subgroup of 05 states soil (including soil excavated from

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contaminated sites) and rock excavation. Within this subgroup, the index numbers of 03–08 different categories of soil referred to as waste, including hazardous waste.

The Criteria for the Remediation and the Remediation of Contaminated Sites Under the remediation of polluted sites, in the broadest sense, it means the implementation of all measures and activities for eliminating the danger to humans and the environment caused by the release of pollutants. With the aim of reducing and preventing soil contamination in numerous studies and research, various technologies for the remediation of soil contaminated with pollutants of organic and inorganic origin have been developed and applied. The choice of suitable technology is complex but a very important step in the successful implementation of the remediation (Khan et al. 2004). After conducting an analysis of the presence and concentration of pollutants in the soil, as well as assessments of their impact on human health and the environment, making a decision on the choice of remediation method is a key step in remediation of contaminated soil. When choosing technology, it is also important to conduct an analysis of the effectiveness of alternative remediation methods in removing contaminated pollutants from the contaminated medium, as well as economic analysis, or determining the amount of financial investments needed to perform remediation.

The Process of Risk Management in the Context of the Remediation of Contaminated Sites For the adoption of appropriate decisions on the treatment of contaminated sites is crucial to assess the degree of risk that contamination has on human health and the environment. The risk is most often estimated on the basis of the total concentration of pollutants in the soil and on the basis of other factors such as: land use (agricultural, land in residential or recreational zones, industrial, etc.), including limit values of pollutant content for certain uses soil and level of contamination (Sahuquillo et al. 2003). Risk management, which may occur as a result of soil contamination or the management of waste streams, includes procedures that include identification, assessment, and final risk assessment, taking into account procedures for their prevention or mitigation, monitoring, and follow-up monitoring. Therefore, the risk management process in the context of the remediation of contaminated sites is a wider concept of the risk assessment and coverage process (Marjanović et al. 2009): – Risk assessment: determining whether there is an unacceptable risk and, in case there is, identification of further activities to be undertaken at the site. – Evaluation of options: evaluation of feasible options for remediation and determining the most appropriate strategy for the remediation site.

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– Implementation: implementation of remediation strategy and demonstration of its efficiency in the present and in the future. Risk assessment is used as a formal instrument of environmental policy by which it determines the way of dealing with contaminated media. US National Research Council (NRC) recognizes four levels of risk assessment (Ferguson et al. 1998): – Hazard identification: identifying agents that can cause side effects. – Dose–response ratio: estimation of the quantitative relationship between exposure and adverse effects confirmed by laboratory experiments or epidemiological tests. – Exposure analysis: estimation of intensity, frequency of occurrence and duration of exposure to hazardous agents (most often involves transport and fate of pollutants in underground and surface waters). – Risk characterization: evaluation and conclusions based on the results of previously conducted steps.

Methods and Techniques of Performing Remediation Remediation can be conducted in a way that implies (Kostic 2007): – Complete removal of pollutants – restoring the polluted area to the state before pollution, what is possible, but technically difficult to implement and financially very demanding solution. – Reduction of pollutants in the contaminated soil to an acceptable level. – Immobilization and blocking pollutants in contaminated media, which does not involve the removal of pollutants, but their immobilization or preventing their movement. For each of the above-mentioned modes, there are different technical possibilities. For example, it is possible to remove or treat pollutants using various physical, chemical, and biological agents. All technologies for remediation of various environmental media can be classified into two main groups, namely (Jaksic and Ilic 2000): – In situ technologies – technologies used on-site mainly for the treatment of large areas of contaminated soil, sediment, large amounts of water, and for milder forms of contamination. – Ex situ technologies – technologies in which the removal of contaminated media (soil, sediment, water) is first performed and its further processing is carried out at another location. This kind of processing is most often used when it comes to extremely contaminated media, small surfaces and quantities, as well as treatments that require special conditions (high temperature, extraction, etc.).

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Table 7 Technologies for soil remediation to the place of application (Kostic 2007)

In situ technologies Biological processes - Bioventilation - Biostimulation - Phytoremediation -Natural remediation Abiotic processes Physical-chemical processes - Steam extraction (vacuum extraction) - Stabilization / solidification - Vitrification - Chemical reduction-oxidation - Washing (a) Other processes - Surface cover

611 Ex situ technologies - Biodegradation in liquid state - Biodegradation in soil state

- Stabilization / solidification - Vitrification - Washing - Chemical reductionoxidation - Chemical extraction (b) Thermal processesa (c) Other processes - Excavation of contaminated land

Thermal processes - desorption - burning - pyrolysis - open burning a

In the in situ process, heat is supplied as part of another process

Technology for the remediation of soil using biotic or abiotic (physical, chemical, or thermal) processes. They may be applied at the site where pollution is found (in situ remediation) or outside the site of pollution (ex situ remediation), when the land pre-excavated and treated in a different location. The following table presents the division of soil remediation technologies to the site of application (Table 7).

Conclusion In the modern world, the major problem is the large area of contaminated sites requiring remediation and restores a state suitable for use. A large share of contaminated sites is derived from waste materials that are inadequate and often improperly disposed directly on the soil. As a result of the mixing of solid and liquid waste materials with substances that are naturally present in the soil, sites whose future use are disabled and which, without recovery, represent a lost natural resource. There are a number of technologies for the treatment of contaminated soil. In selecting appropriate technology for the treatment or elimination of pollutants from soil should be taken into account a number of factors. The most important are the type of pollutants at the site, the level or degree of contamination, the possible consequences for the living world, the surface of the contaminated medium, the location, etc.

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During remediation, it is possible to simultaneously apply several different methods. An individual remediation method may support numerous activities or operations for the reduction or complete elimination of pollutants. However, in some cases it is not enough to apply one method to solve all the problems that exist at the site. It is possible that one method is suitable for the treatment of certain pollutants, while it is not suitable for the treatment of others. In order to ensure the efficiency of such surge remediation strategies and ensure its practical applicability, it is necessary to carefully and detailed planning and design. The aim of this chapter is to contribute to faster and more efficient decisionmaking and to expand current knowledge about how to perform effective remediation. It should also help the participants in the process of remediation to understand and apply the necessary remediation measures and use the end product of remediation in the form of useful products. It is extremely important to help and understand the concept of sustainable waste management as a measure of prevention for the recovery and remediation process.

Cross-References ▶ Monitoring and Risk Analysis of PAHs in the Environment

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Part III Hazardous Waste Management

Development of an Environmentally Sustainable Approach for Safe Disposal of Arsenic-Rich Sludge

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Pankaj Kumar Roy, Arunabha Majumder, Somnath Pal, Gourab Banerjee, Malabika Biswas Roy, Jayanta Debbarma, and Asis Mazumdar

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Treatment Methods for Arsenic-Rich Sludge . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Preparation of Arsenic-Rich Sludge by Employing Oxidation and Precipitation Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Arsenic Concentration in Filtered Supernatant Water After Coprecipitation . . . . . . . . . . . . . . . Preparation of Concrete (M15-1:2:4) with Arsenic-Rich Sludge . . . . . . . . . . . . . . . . . . . . . . . . . . . Use of Arsenic-Rich Sludge in Brick Manufacturing . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Disposal of Arsenic-Rich Sludge in Anaerobic Bioreactor . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Results and Discussion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Arsenic-Rich Sludge Used for Concrete Cubes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Arsenic-Rich Sludge Used in Brick Preparation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Arsenic-Rich Sludge Stabilization in Anaerobic Bioreactor . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion and Recommendations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

Arsenic contamination in groundwater has been reported by many researchers from different parts of the world. In West Bengal, arsenic contamination in P. K. Roy (*) · S. Pal · G. Banerjee · A. Mazumdar School of Water Resources Engineering, Jadavpur University, Kolkata, West Bengal, India e-mail: [email protected] A. Majumder School of Water Resources Engineering, Jadavpur University, Kolkata, West Bengal, India All India Institute of Hygiene and Public Health, Govt. of India, Kolkata, India M. B. Roy Women’s College, Calcutta, Kolkata, India J. Debbarma PWD (Water Resources), Agartala, Tripura, India © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_14

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groundwater was first detected in 1983 when a few patients with arsenicosis reported at the School of Tropical Medicine, Kolkata. Today, arsenic contamination in groundwater is found to be affecting 82 blocks of eight districts of West Bengal, namely, Maldah, Murshidabad, Nadia, North 24 Parganas, South 24 Parganas, Burdwan, Hooghly, and Howrah, and also in 11 municipal areas and 18 non-municipal outgrowth areas. Providing safe drinking water to people in rural community is a major challenge in arsenic-affected areas in and around the world. One of the options for supplying arsenic-free and potable water is to remove arsenic from contaminated groundwater, and the second option is to provide potable water filtered from surface water system by various mechanical filtration technologies. Both these options involve a huge amount of cost and manpower and adaptation of suitable engineering methods and economically feasible solutions. On the other hand, arsenic removal processes generate arsenic-rich sludge which requires safe disposal as because the sludge becomes hazardous. Disposal of arsenic-rich sludge generated from contaminated water by the method of coprecipitation and adsorption is a major environmental concern. Qualitatively arsenic-rich sludge is hazardous, and uncontrolled disposal may lead to environmental degradation. In order to stabilize arsenic-rich sludge, it was mixed in different proportions with cement concrete and clay soil. In the first phase of the experiment, the compressive strength of the concrete cubes and bricks was analyzed, and toxicity characteristic leaching test was conducted to determine the quantity of arsenic in the leachate. In the second phase, injection of the sludge into bench-scale anaerobic bioreactors was carried out to monitor the stabilization of arsenic-rich sludge. The toxicity characteristic leaching test for all the concrete cubes and bricks indicated the presence of arsenic concentration in leachate within the permissible limit. The study showed that arsenic-rich sludge could be potentially disposed through environmentally friendly manner by mixing with cement concrete and bricks. Keywords

Adsorption · Compressive strength test · Concrete · Coprecipitation · Leachate · TCLP · Groundwater

Introduction In nature, arsenic (As), in its various forms, is widely presenting both surface and subsurface of the earth’s crust. It gets introduced in water through dissolution of minerals and ores and through erosion from natural sources. Arsenic is of major environmental concern because of its toxicity to plants, animals, and human beings, although the necessary intake may be as low as 0.01 mg/day (http://www.lenntech. com/periodic/elements/as.htm). Its presence in natural water may originate from geogenic formations, geochemical reactions, industrial waste discharge, burning of arsenic-containing fossil fuels, volcanic eruptions, and arsenic pesticides apart from

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its existence in soil. Soil is an important natural resource for mankind, but it also serves as a medium for accumulation, transformation with subsequent disintegration, and migration of toxicants. The occurrence of arsenic in groundwater is mostly due to the leaching of geological material (Purkait and Roy 2006), mineral precipitation, dissolution of unstable arsenic minerals, adsorption-desorption, chemical transformations within formation, input from geothermal sources, etc. (Purkait and Roy 2006). The arsenic-contaminated groundwater poses serious risk to the people who depend on groundwater. Consumption of arsenic-contaminated water for a prolonged period may cause arsenicosis with serious health disorders. Arsenic contamination in groundwater has been reported by many researchers from different parts of the world. In West Bengal, arsenic contamination in groundwater was first detected in 1983 when a few patients with arsenicosis reported at the School of Tropical Medicine, Kolkata (Nath and Mazumdar 1999). Today, arsenic contamination in groundwater is found to be affecting 79 blocks of eight districts of West Bengal, namely, Maldah, Murshidabad, Nadia, North 24 Parganas, South 24 Parganas, Burdwan, Hooghly, and Howrah, and also in 11 municipal areas and 18 non-municipal outgrowth areas (AIIH & PH 2007). About 42.7 million people have arsenic levels in groundwater above the World Health Organization (WHO) maximum permissible limit of 50 μg/L (Chowdhury et al. 2000). Arsenic is basically found in the eastern part of the river Ganga due to deposition of Late Quaternary sediments. In order to investigate the mechanism of As released to anoxic groundwater in alluvial aquifers, the authors sampled groundwaters from 3 piezometric wells, 79 shallow (80 m) wells, in an area of 750 m by 450 m, just north of Barasat, near Kolkata (Calcutta), in southern West Bengal (McArthur et al. 2004). However surface water sources and dug wells were found free from arsenic contamination. The provision of availability of safe drinking water is one of the prior conditions necessary for the overall socioeconomic development. Arsenic in drinking water is a global problem where as many as 170 million people in over 70 countries have been affected (IWA 2007). The sinking of tube wells for drinking water in Bangladesh has in the past been promoted as a means of avoiding waterborne diseases associated with the consumption of untreated surface water. Unfortunately, the recent discovery of arsenic in groundwater in Bangladesh has reached up to 2 mg/L, which has led to describe the situation as a major public health emergency (Allan et al. 2000).The recent innovation of Sono 3-Kalshi filter (Hussam et al. 2008) could be a landmark for mitigation of arsenic in the drinking water. Solid waste management in developing countries is often unsustainable, recycling on uncontrolled disposal in waste dumps. A particular problem arises from the disposal of treatment residues generated during removal of arsenic (As) from drinking water because As can be highly mobile and has the potential to leach back to ground surface waters (Sullivan et al. 2010). Providing safe drinking water to people in rural community is a major challenge in arsenic-affected areas in and around the world. One of the options for supplying arsenic-free and potable water is to remove arsenic from contaminated groundwater, and the second option is to provide potable water filtered from surface water system by various mechanical filtration technologies. Both these options involve a huge

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amount of cost, manpower, and adaptation of suitable engineering methods and economically feasible solutions. On the other hand, arsenic removal processes generate arsenic-rich sludge which requires safe disposal as the sludge is most hazardous for environment. Appropriate sustainable methods are needed to be developed for safe disposal of arsenic-rich sludge and engineered media and the same need to be practiced when arsenic removal units are installed. Whatever may be the sizes of arsenic removal units, arsenic-rich sludge is considered to be hazardous waste, and uncontrolled disposal may cause adverse impact on the environment. It is therefore necessary to provide eco-friendly treatment to arsenic-rich sludge so that their disposal will not cause any adverse impact to the environment. In the present study, an appropriate eco-friendly method is designed for safe disposal of arsenicrich sludge in order to (a) examine the extent of arsenic richness of sludge, (b) assess the quantum of sludge generated after treatment, (c) study different methods for stabilizing sludge and the leach ability of stabilized arsenic-rich sludge through TCLP testing, and (d), finally, recommend appropriate treatment and safe disposal system for arsenic-rich sludge. An environmentally sustainable development framework is needed and is conceptualized through experimental design in the present study considering three subsystems of economy, society/community, and environmental interaction. Mostly researchers have been focused on awareness building and the development of water treatment system removing arsenic from drinking water. The disposal of arsenic-rich sludge generated from the different techno-treatment processes is one of the issues that have received little attention from the sponsors of the technologies and the users (Eriksen-Hamel and Zinia 2001; Kameswari et al. 2001; Dutre and Vandeeasteele 1995; Akhtar et al. 2000). The solidification or stabilization process would be the best practical technology to treat the arsenic waste (Artiola et al. 1990; Voigt et al. 1996; Vandeeasteele et al. 2002; Leist et al. 2003). It is found appropriate by many investigators in treating arsenic-contaminated wastewater (Fuessle and Taylor 2000; Sanchez et al. 2003; Shih and Lin 2003; Pal 2001). It is further found in one study that the recommended proportion of contaminated sludge in brick making is up to 15–25% by weight (Rouf and Hossain 2003). A similar study was carried out by Banerjee and Chakraborty (2005) where they used arsenic-contaminated sludge for the preparation of concrete cubes, briquette production, and cement mortar preparation. The study depicted that arsenic sludge can safely be utilized for briquette production along with clay up to a proportion of 10% (v/v) with respect to the total mixed ingredients; the leachate formed by TCLP contains As concentration below its safe discharge limit of 0.2 g/m3 set by the Central Pollution Control Board (CPCB) under the Ministry of Environment and Forest (MoEF), Government of India (GOI). Arsenic sludge can be stabilized by standard cement concrete ingredients up to a proportion of 40% (v/v) with respect to the total volume of the mixture, and the leachate formed in the TCLP test contains arsenic below the desired safe discharge limit. However concrete or clay mortar that is richer in arsenic-laden sludge showed an increase in the amount of leachate generated. In the preparation of cement mortar, the arsenic-laden sludge could be

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blended up to a proportion of around 10% (v/v) with respect to cement content. Banerjee and Chakraborty (2005) had also shown that replacement of cement by fly ash up to a proportion of 50% (v/v) of cement results in 45–50% reduction in the compressive strength of cement mortar at an arsenic sludge content of 10%. Studies also showed that arsenic-contaminated sludge could be used safely up to 4% for making ornamental bricks (Mahzuz et al. 2009). While considering option for waste disposal, it is essential that appropriate leaching tests are applied and results are correctly interpreted. Leist et al. (2003) undertook an evaluation of leaching test and commented that they did not model the conditions that the waste would experience when placed in landfill. Badruzzaman (2003) reported that the application of TCLP may not be suitable for assessment of long-term leaching of As from As-rich waste, as such leaching may be kinetically restricted. Eriksen-Hamel and Zinia (2001) carried out a study on the leaching characteristics of arsenic-contaminated sludge in some selected part in Bangladesh like Pabna, Rajshahi, Kachua, Nawabganj, Maijdi, etc. Their results show that the arsenic concentration in TCLP extraction fluid of different samples taken from different places toxicity is well below the level of “hazardous waste.” Visoottiviseth and Ahmed (2008) carried out different technology option for remediation and disposal of arsenic. In their study, percentage removal of arsenic using different technologies, namely, conventional filtration, phytoremediation using arsenic hyperaccumulating fern, bioremediation using Chlorella vulgaris algae, combined sand and nano-filtration, reverse osmosis, low pressure nano-filtration and reverse osmosis, ion exchange, cartridge filter, indigenous filter, adsorption by activated carbon, activated alumina, lime treatment using calcium oxide or calcium hydroxide, coagulation and filtration, solar oxidation, chemical oxidation, in situ oxidation, passive sedimentation, oxidation, etc., has been analyzed, and an alternate water supply option (arsenic-free) is emphasized. The alternative sources that they pointed out were deep tube-well dug/ring well, rain water harvesting (RWH), etc. As far as our study is concerned, different methods have been applied for assimilation of arsenic sludge such as using arsenic sludge in making concrete cubes, in preparation of bricks with certain proportion replacement, and using the same in an anaerobic bioreactor where arsenic is consumed by the anaerobic bacteria resulting in a decrease in overall concentration of arsenic sludge. The overall objective of this present research is to explain the management procedure of the arsenic rich sludge which generated from after treatment of hand pump attached arsenic removal units (ARUs). Since the units are installed in the rural areas where fly ash is not abundantly available due to the absence of any thermal power plant in and around the investigation area, so the research work has done with OPC cement concrete mixture. Despite the successful development in the laboratory of technologies for arsenic remediation, few have been successful in the field. A sustainable arsenic remediation technology should be robust, composed of local resources, and user-friendly as well as must attach special consideration to the social, economic, cultural, traditional, and

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environmental aspects of the target community. One such technology is in operation on the Indian subcontinent. Wide-scale replication of this technology with adequate improvisation can solve the arsenic crisis prevalent in the developing world.

Treatment Methods for Arsenic-Rich Sludge The arsenic-rich sludge is prepared by employing oxidation and precipitation processes and in filtered supernatant water after coprecipitation. The following preparation methods are employed to prepare the arsenic-rich sludge: (i) (ii) (iii) (iv) (v) (vi)

Preparation of concrete (M15-1:2:4) with arsenic-rich sludge Use of arsenic-rich sludge in brick manufacturing Disposal of arsenic-rich sludge in anaerobic bioreactor Arsenic-rich sludge used for concrete cubes Arsenic-rich sludge used in brick preparation Arsenic-rich sludge stabilization in anaerobic bioreactor

Preparation of Arsenic-Rich Sludge by Employing Oxidation and Precipitation Processes Arsenic stock solution (1000 mg/L) was added in requisite proportion to prepare spike solution of arsenic having an arsenic concentration of 0.5 mg/L. Arsenic-rich sludge in bulk quantum was generated at the laboratory of the School of Water Resources Engineering, Jadavpur University. Bleaching powder and alum were added in arsenic-contaminated water at the rate of 5 mg/L and 70 mg/L, respectively. Both the chemicals were rapidly mixed for 60 s to generate the effect of coagulation. It was then followed by slow mixing for 10 min for development of aluminum hydroxide flocs. The water in the container was then allowed to settle for 2.5 h. The concentrated arsenic-rich sludge was collected as bottom sludge after draining out the supernatant. The quantum of sludge could be estimated as follows: Al2 ðSO4 Þ3 2AlðOHÞ3 :18 H2 O þ 3 CaðHCO3 Þ2 ! þ 3 CaSO4 þ 3 CO2 þ 18 H2 O ð666Þ ð156Þ In a 15 L bucket filled with arsenic-contaminated water (As-0.5 mg/L), 1.05 g of alum was added. In the process aluminum hydroxide of 0.246 g was formed. The supernatant water after coprecipitation process was found to contain 0.03 mg/L of arsenic. The arsenic content in raw was analyzed as 0.53 mg/L as depicted in Table 1. So arsenic adsorbed by aluminum hydroxide was estimated to be around 7.05 mg. The total dry weight of arsenic-rich sludge formed by coprecipitation process in a

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Table 1 Standardization for generation of arsenic-rich sludge Volume of raw water (L) 15 15 15

Arsenic concentration in raw water (mg/L) 0.51 0.53 0.52

Arsenic concentration in supernatant water (mg/L) 0.02 0.03 0.03

Bleaching powder dose (mg/L) 5.0 5.0 5.0

Alum dose (mg/L) 70 70 70

Dry weight of sludge (g) 0.25335 0.25305 0.25335

Table 2 Arsenic concentration in raw water and filtered supernatant water after coprecipitation

Sample name ARS-1 ARS-2 ARS-3 ARS-4 ARS-5 ARS-6 ARS-7 ARS-8 ARS-9 ARS-10

Arsenic concentration in raw water (mg/L) 0.68 0.64 0.55 0.50 0.58 0.64 0.68 0.76 0.52 0.54

Arsenic concentration in supernatant water after coprecipitation and filtration (mg/L) (in %) 0.045 (93.39) 0.055 (91.40) 0.064 (88.36) 0.044 (91.20) 0.058 (90.0) 0.052 (91.88) 0.075 (88.97) 0.078 (89.74) 0.040 (92.31) 0.048 (91.11)

Average removal % of arsenic by the method of coprecipitation 90.84

15 L bucket water was (0.246 + 0.00705) g, i.e., 0.25305 g. The above process was repeated several times to generate the arsenic-rich sludge.

Arsenic Concentration in Filtered Supernatant Water After Coprecipitation Arsenic-rich sludge was prepared by the process of coprecipitation, utilizing alum and bleaching powder. The sludge was stored for carrying out study for its safe and eco-friendly disposal. The arsenic concentrations in raw water as well as filtered supernatant water for different batch operations are presented in Table 2. Arsenic was measured by the model AAnalyst 200 Atomic Absorption Spectrophotometer (AAS) with the model MHS-15 Mercury/Hydride Generation System in the laboratory of the School of Water Resources Engineering, Jadavpur University, according to the standard methods (APHA 2005). The water quality analysis indicated that 90.84% arsenic removal could be possible by the method of coprecipitation. The percentage of arsenic concentration in supernatant water after coprecipitation decreased with the decrease in the arsenic concentration in raw water.

626 Table 3 Arsenic-rich sludge analysis

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Sample name ARS-1 ARS-2 ARS-3 ARS-4 ARS-5 ARS-6 ARS-7 ARS-8 ARS-9 ARS-10

Arsenic concentration in sludge (mg/kg) 2775 2550 2100 1930 2290 2680 3130 3250 2275 2380

Iron concentration in sludge (mg/kg) 32,250 31,150 36,700 44,474 37,250 35,154 32,197 34,850 35,190 37,198

The arsenic-rich sludge was analyzed for arsenic and iron concentration for different batch processes. The arsenic concentration in the sludge varied from 1930 to 3250 mg/kg. The iron concentration in the sludge was found to be in the range 31,150–44,474 mg/kg as presented in Table 3. The arsenic-rich sludge was digested to solubilize particulate forms to measure total arsenic and oxidize reduced forms of arsenic and to convert any organic compounds to inorganic ones (APHA 2005). Arsenic and iron were analyzed by using the model Analyst 200 Atomic Absorption Spectrophotometer (AAS) with model MHS-15 Mercury/Hydride Generation System in the laboratory.

Preparation of Concrete (M15-1:2:4) with Arsenic-Rich Sludge Concrete cubes were prepared by using cement, sand and stone chips, and water. In order to prepare concrete cube, three different sizes of molds, namely, 50 mm, 75 mm, and 100 mm, were used. The specifications of different ingredients of concrete are furnished below: • Cement: Ordinary Portland Cement (OPC) of 53 grades conforming to the Indian Standard Code of practice manufactured by ACC was used in the study. • Sand: Effective size, 0.28 mm; uniformity coefficient, 2.2; type of sand, river sand. • Stone – chips: 18 mm downgrade. • Water: collected from the Kolkata Municipal Corporation (KMC) tap. The sludge was added (by weight) in 1%, 0.5%, and 0.1% proportion in M15 concrete preparation. Each of the concrete mixture was placed in three same sizes of mold, and accordingly nine molds were used for concrete cube preparation. The concrete cubes were cured for 28 days keeping in water chamber, and compressive

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Table 4 Sludge addition in concrete preparation Sample name ARS-1 ARS-2 ARS-3 ARS-4 ARS-5 ARS-6 ARS-7 ARS-8 ARS-9

Sludge conc. (%) 1.00 1.00 1.00 0.50 0.50 0.50 0.10 0.10 0.10

Sludge added (g) 6.65 6.65 6.65 5.50 5.50 5.50 5.65 5.65 5.65

Cement added (g) 95 95 95 158 158 158 808 808 808

Sand added (g) 190 190 190 314 314 314 1614 1614 1614

Stone chips added (g) 380 380 380 628 628 628 3228 3228 3228

Water added (mL) 48 48 48 80 80 80 411 411 411

strengths were detected for each concrete cube. The details of aggregate and sludge added for concrete preparation are furnished in Table 4. TCLP testing was carried out for all concrete cubes.

Use of Arsenic-Rich Sludge in Brick Manufacturing Nine small bricks were manufactured by using soil and varying proportion of sludge. The brick dimension was 10.8  5.7  3.2 cm. The soil was collected from Indira Gandhi Water Treatment Plant, Palta, constructed for laying water pipeline from Palta to Kolkata city to get treated water. The dried soil in each brick was measured approximately as 428 g. Dry sludge having arsenic concentration between 1% and 3% was added with clay to manufacture bricks. Distilled water was used in brick manufacturing. The brick was sun-dried before burning in muffle furnace at a controlled temperature and duration. TCLP testing of all the bricks were carried out in the laboratory.

Disposal of Arsenic-Rich Sludge in Anaerobic Bioreactor Sewage sludge was collected from the activated sludge plant (ASP) located near Howrah Station (Eastern Railway) for feeding the bioreactor. The sewage sludge was analyzed in the laboratory, and the characteristics of such sludge have been presented in Table 5. The physicochemical, chemical, and bacteriological parameters were analyzed according to the standard methods (APHA 2005). The pH value was measured by a rugged field kit. Three bench-scale anaerobic bioreactors made of glass, having a capacity of 5.5 L each, were installed at the laboratory of the School of Water Resources Engineering, Jadavpur University, to monitor the stabilization of arsenic-rich sludge by anaerobic bacteria. Initially all the bioreactors were inoculated by anaerobic bacteria collected from a septic tank. In each bench-scale anaerobic bioreactor, 5 L of sewage sludge was taken as feed.

628 Table 5 Test results for sewage sample used in anaerobic bioreactor

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Sl. no. 1 2 3 4 5 6 7 8 9 10 11 12

Parameters pH Total solids (mg/L) Volatile solids (mg/L) Volatile acids (as CH3COOH), mg/L Organic solids (mg/L) Biochemical oxygen demand (mg/L) Chemical oxygen demand (mg/L) Sulfate (as SO4), mg/L Nitrate as (NO3), mg/L Iron (as Fe), mg/L Total coliform (MPN/100 ml) Fecal coliform (MPN/100 ml)

Results 6.7 42,660 10,320 27.3 10,186 11,442 20,373 12.5 0.62 6.4 8  107 2.2  107

Results and Discussion Arsenic-Rich Sludge Used for Concrete Cubes Compressive strength for all the concrete cubes was tested in the laboratory, and the results of such test are furnished in Table 6. The comparative study for compressive strength indicated except for two samples all the other samples conform to the standard compressive strength for M15 concrete as per IS 456, 1978. The arsenicrich sludge and concrete mixing ratio of 1.0:99, 0.5:99.5, and 0.1:99.9 developed compressive strength of 15 N/mm2. So, qualitatively the concrete is acceptable for use. Table 6 indicated that sludge proportion with size of concrete was the key factor for determining the quality of concrete cubes and bricks. The compressive strengths of concrete cubes decreased with the increase in sludge proportion. All the concrete cubes were used for carrying out TCLP test in the laboratory. The leachate of TCLP tests for the concrete samples was analyzed, and results are furnished in Table 7. The arsenic content in leachate after TCLP test indicated that the concentration of arsenic was much below the maximum permissible value as prescribed by EPA (1992), i.e., 5 mg/L. Leachates analyzed for As indicated presence of arsenic in leachate of concrete cubes ranged between 0.012 and 0.064 mg/L.

Arsenic-Rich Sludge Used in Brick Preparation The TCLP results highlighted that the arsenic concentrations in leachates of all the bricks were found to be below the permissible limit (5 mg/L) as per the US-EPA Standard (see Table 8). Now, as per IS 3102 (1971) and IS 1077 (1992), factorymade bricks in India have the strength of 17 N/mm2 in dry condition. But common

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Table 6 Compressive Sample strength of concrete cubes name ARS-1 ARS-2 ARS-3 ARS-4 ARS-5 ARS-6 ARS-7 ARS-8 ARS-9

Table 7 Arsenic concentration in leachate of concrete cubes after TCLP test

Sample name ARS-1 ARS-2 ARS-3 ARS-4 ARS-5 ARS-6 ARS-7 ARS-8 ARS-9

Sludge in mixture (%) 1.00 1.00 1.00 0.50 0.50 0.50 0.10 0.10 0.10

Size of concrete cubes (mm) 75.0 75.0 75.0 75.0 50.0 50.0 100.0 100.0 75.0

Mixing of sludge (%) in concrete 1.00 1.00 1.00 0.50 0.50 0.50 0.10 0.10 0.10

629 Compressive strength (N/mm2) 12.99 12.68 12.24 15.30 14.56 15.20 18.93 18.47 16.22

Arsenic concentration (mg/L) in leachate 0.029 0.054 0.045 0.064 0.055 0.039 0.014 0.012 0.016

Table 8 TCLP results for bricks made by using soil and arsenic-rich sludge

Sample name ARS-1 ARS-2 ARS-3 ARS-4 ARS-5 ARS-6 ARS-7 ARS-8 ARS-9

Dry weight of soil (g) 428 420 425 427 424 425 428 424 426

Dry weight of sludge (g) 4.28 (1%) 5.04 (1.2%) 4.25 (1%) 8.54 (2%) 8.48 (2%) 10.20 (2.4%) 12.84 (3%) 10.6 (2.5%) 12.78 (3%)

Comprehensive strength (N/mm2) 3.4 3.8 3.4 4.1 4.1 4.9 5.2 5.1 5.2

TCLP results of leachate Arsenic, Iron, mg/L mg/L 0.0010 0.044 0.0012 0.090 0.0010 0.090 0.0034 0.183 0.0025 0.160 0.0020 0.150 0.0015 0.150 0.0018 0.160 0.0020 Nil

hand-made bricks can have their strength between 3 and 5 N/mm2. Brick samples were tested, and those brick samples showed strength ranging between 3.4 and 5.2 N/mm2.

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Arsenic-Rich Sludge Stabilization in Anaerobic Bioreactor Initially 370 mL arsenic sludge was added in each of the three anaerobic bioreactors to mix with 5 L of sewage sludge. The anaerobic bioreactors were sealed properly keeping an outlet for the release of methane and carbon dioxide gas. The anaerobic reactor was kept at a warm place (average temperature- 30  C) for 75 days. The digested sludge samples collected from the bioreactors were analyzed, and the results are represented in Table 9. Similarly, supernatant liquid of all the three bioreactors was also analyzed for different parameters which are listed in Table 10. The anaerobic bioreactor was operated in batch process. Kinetics of sludge digestion involves two groups of bacteria, namely, acid formers and methane formers. The digester operated as single-stage reactor and without recirculation. The hydraulic retention time (HRT) was the same as the solids retention time (SRT). The volatile solid in the influent sludge of the reactor was measured as 10,320 mg/L (Table 5) indicating volatile solids loading as 10.32 kg/m3/day. It indicated that thickened sludge was added in the reactor. The volatile acid in the influent sludge was detected as 27.3 mg/L, and the same was found to be in the range of 149.37 mg/L and 158.93 mg/L as depicted in Table 10 in the reactors indicating sludge digestion at lower rate. The volatile solids reductions in the reactors were observed between 23% and 37% in the reactors observed from Table 9. Normally in Table 9 Test result for the sludge sample Sl. no 1 2 3 4 5 6 7

Parameter pH Total solid (g) Volatile solid (%) Volatile acid (as acetic acid) (%) Organic solid (%) Arsenic (as As) (%) Aluminum (as Al) (%)

Results Bioreactor-1 7.20 157.28 37.8 0.437 15.05 0.008 0.545

Bioreactor-2 6.92 146.31 22.24 0.394 14.74 0.007 0.502

Bioreactor-3 7.10 162.36 24.50 0.512 15.28 0.009 0.647

Table 10 Test results for the supernatant sample Sl. no 1 2 3 4 5 6 7 8 9

Parameters 3 days BOD at 27  C, mg/L COD, mg/L pH Volatile solids, mg/L Volatile acids (as CH3COOH), mg/L Total solids, mg/L Organic solids, mg/L Arsenic (as As), mg/L Aluminum (as Al), mg/L

Results Bioreactor-1 8.53 54.33 5.81 51.33 158.93 790 49.33 0.04 1/103 >1/104 >1/105 >1/106

Characterization Very probable Probable Often probable Often but probable Slightly probable Not very probable Completely not probable

Level of reported hazarda High Significant Average (4) (3) (2) S S S S S S S S M S M M M M L M L L A A A

Low (1) M L L A A A A

Nonexistent (0) A A A A A A A

a

Hazard characterization: None/negligible (Α) = 0, low (L) = 1, medium (M) = 2, significant (S) = 3

to happen, and the hazard level is significant when labors do not apply or are ignorant of the first aid measures within the facility (0.45% as depicted in Fig. 5). In the set of scenarios that were run for all the hazard levels examined, several levels of impact probability occurred. This, in combination with the results of Fig. 4, suggests that the possibility of a significant level of risk (4) occurring is negligible. In general, HHW has a negative impact on the MRF’s workforce, and the risk is moderate for each level of risk control. This is reinforced by the fact that eventual risk level 1 cases have a very low chance of occurrence. The most important outcomes of this assessment are summarized as follows: • • • •

83% of risk level 1 is very unlikely to occur. 64.03%v of risk level 2 has medium probability. 79.8% of risk level 3 has medium probability. 86.73% of risk level 4 has medium probability.

For the case of transfer stations and sanitary landfills, 927 scenarios were investigated; the average risk is characterized as medium. The medium probability of risk is the outcome of 57.61% of the scenarios. On the other hand, significant probability of level 4 hazard occurs when the parameter of “no protection measure taken” is taken into consideration. Overall, the probability of a level 4 hazard is low but certainly higher than the one calculated in the case of MRF; high percentages of risk occurrence were calculated in the case of medium hazard (M) of levels 2, 3, and 4, being ~59%, ~78%, and ~69%, respectively. For the case of HHW disposal in municipal bins, 305 scenarios were investigated. On average, the risk is low and occurs in 45.74% of the scenarios. Overall, the scenarios proved that in 50.87% of the cases investigated, the probability of any level of risk to the citizens is negligible; in 16.03% of the cases, it is low; and in 33.10% of the cases, it is medium mainly due to the parameter of “animal’s access in bins.” On the other hand, the risk due to chemical compound, which bears biological hazards (DNA corruption, introduction into the food chain, etc.), is high. Similarly, for the compounds which are characterized as being hazardous for the environment (N),

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0.45%

100% 22.32%

80%

14.73%

18.75%

18.75%

18.75%

10.27%

6.25%

6.25%

7.59%

25.00% 46.88%

11.61%

37.95%

19.53%

60%

37.05%

40% 66.07%

75.00%

74.55%

75.00%

73.66%

37.50% 55.47%

20% 15.63%

25.00%

0% Flammable Harmful (F/F+) (Xn) M

Toxic (T) Dangerous Corrosive Οxidising Biological (C) for the (O) (B) environment (N) A S

Irritating (Xi) L

Fig. 5 Probability characterization per hazard level category and property of the chemical compounds in hazardous household waste in MRF

it has 50.91% probability of having medium impacts to city environment and citizens. Mainly based on the later two compounds’ properties, high percentage of scenarios where the probability of occurrence is medium for hazard levels 2, 3, and 4 was observed (being ~32%, ~37%, and ~56%, respectively).

Dental Waste Fraction The development of scenarios for the investigation of DW hazard was affected by the negligible probability of them being treated in MRF; the scenario was considered nonachievable, so their existence was considered a fact for transfer stations and sanitary landfills, and 927 scenarios were developed. On average, the hazard risk is medium (58.04%). The significant hazard risk (level 4) for those facilities’ labors occurs when parameters such as “the use of protective equipment” and the “presence of safety officer” are not applied. The “presence of litter in the perimeter” increases the introduced toxicity in the flora of the nearby area. Out of the total scenarios which were run for the determination of hazard risk in the case of DW presence in sanitary landfills and transfer stations, 28 of them show a significant probability of high hazard risk. This disturbing fact is attributed mainly to the chemical nature of the waste fraction compounds. Despite the fact that level 1 hazard risk occurs most often (>80%), they concern negligible hazards, while the significant level hazards 3 and 4 are assessed to have medium probability of occurrence (~75%) as depicted in Fig. 6. The higher probabilities for hazard risk are summarized as follows: • • • •

Low probability of level 1 hazard risk (~82%) Medium probability of level 3 hazard risk (~78%) Medium probability of level 4 hazard risk (~74%), και Medium probability of level 2 hazard risk (~52%)

Percentage of hazard risk (%)

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L

M

S

Level of hazard LEVEL 1

LEVEL 2

LEVEL 3

LEVEL 4

AVERAGE

Fig. 6 Percentage of hazard risk for every level of hazard risk (the average hazard risk is indicated as well) 100% 80%

41.67%

37.50%

40.63%

40.63%

25.00%

25.00%

60%

24.62%

20.83%

40%

32.31%

33.33%

20%

31.25%

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Toxic (T)

43.08%

41.67% 25.00%

50.00%

0% Flammable (F/F+)

Harmful (Xn)

M

L

A

Dangerous for the environment (N)

Corrosive (C)

Fig. 7 Hazard risk characterization per hazard category and per property of the included chemical compounds of DW fraction in urban areas

Five hundred sixty-four scenarios were investigated for the case of DW presence in municipal bins, and the average hazard risk is calculated to be as low as 43.25%. A significant probability of hazard risk was not found (Fig. 7), while the higher probabilities were calculated when the parameters “existence of litter nearby” and “existence of animals in bins” were applied. Those parameters are linked and common within the urban areas. The medium probability of hazard risk is on average 36%, and the low one is 27.71%. It is important to mention that the average is exceeded only for the case of chemical compounds with the characteristic of being hazardous for the environment (N) and rises up to 43%. Similarly, it happens with those that bear corrosive properties (~50%). Under the average hazard risk calculation, there are no concerns for the introduced risk by the chemical compounds of the DW fraction. However, the individual

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study of their properties may be the starting point for a discussion/consultation to change the composition of some products. Besides it came up that, in level 4 hazard risk, the probability is medium in 40 of those compounds and negligible for the rest of them in case the parameter “potential fire in bin” is applied.

Conclusions Presented in this research paper was the first phase out of an extended (four phases) research during which researchers aimed to identify the hazardousness of the two selected fractions (HHW and DW) which currently in Greece are not managed separately from the MSW (Kontogianni 2013). Any researcher or stakeholder aiming to design or implement a holistic waste management system to avoid treatment of hazardous fractions among the MSW and thus prevent occupational accidents by developing enhanced (integrated) safety procedures can use these approaches and outcomes since they have a global relevance. The major outcome of the overall research would be the development of separate HHW collection and management programs as in other countries worldwide and the incentive for the actual organization of DW program for all private, public, and university dental units since the legislation already exists for more than a year in Greece. The findings regarding the chemical content of the investigated fractions presented in this paper present strong arguments toward the need of proper management and implementation of waste hierarchy at higher levels. The uncertainty about the environmental and safety hazards due to the presence of these currents throughout the MSW management course rises as it continues its course from the bin to the final disposal or treatment. The newly introduced HHW stream and the unobstructed rejection of OA in mixed bins do not favor the availability of data on the chemical reactions that take place or the exact composition of the compounds in each product. The General Chemical State Laboratory, which is a national agency, should adapt strict measures toward the investigation and frequent recording of the investigated commercial products’ chemical content to minimize the hazard primarily arising from their disposal in municipal waste bins (mE I). The introduction of an environmental certificate for such product producers is the means to do so, but the overall procedure should be overseen by an independent agency to achieve maximum efficiency. The aforementioned will assist in minimizing the occupational hazard in SWM facilities, which is due to end-of-life product mismanagement. Due to the presence of HHW, MRF workers face a 68.37% chance of having serious (level 4) health impacts, while 68.8% of workers in transfer stations and sanitary landfills have a significant impact (level 3). For the first case mentioned, 14 substances and for the second one, 10 substances were tested for their toxicological effect. Three and four of these substances were considered the most toxic with certain toxic effects in the body and for which their critical concentration in the air of the workplace was examined. It is necessary and urgent that the strengthening of health and safety measures and intensive training of occupational health and safety officers and personnel in SWM

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facilities be a prerequisite to minimize the impacts of the chemicals’ properties mostly in mE II. Besides the existence of a structured plan for every facility, the frequent medical examinations (under management supervision and not after personnel’s initiative and charge) will seriously act toward the prevention of health risks related to this working activity.

Cross-References ▶ A Review on Treatment of Pharmaceuticals and Personal Care Products (PPCPs) in Water and Wastewater ▶ Household Hazardous Waste Quantification, Characterization, and Management in Developing Countries’ Cities: A Case Study ▶ Sustainable Biomedical Waste Management

References EEA (2010) European Environmental Agency, Waste and material resources. http://www.eea. europa.eu/themes/waste. Accessed Nov 2010 Ferraris D, Kievit-Kylar D, Moreau T, Moroukian M, Shaner H, Taylor P (1999) Responsible dental office: a guide to proper waste management in dental offices. Northeast Natural Resource Center of the National Wildlife Federation and The Vermont State Dental Society. http://www.mercvt.org/ Gendebien A, Leavens A, Blackmore K, Godley A, Lewin K, Franke B, Franke A (2002) Study on hazardous household waste with a main emphasis on hazardous household chemicals. European Commission – Directorate general Environment Final report. Belgium (WRc Ref: CO 5089-2) Gladding TL, Thorn J, Smith R (2003) Air quality and worker health effects in materials recovery facilities (MRFs) in England and Wales, Sardinia 2003: proceedings of the 9th international landfill symposium, 6–10 Oct, Cagliari Gu B, Zhu W, Wang H, Zhang R, Liu M, Chen Y, Wu Y, Yang X, He SH, Cheng R, Yang J, Bi J (2014) Household hazardous waste quantification, characterization and management in China’s cities: a case study of Suzhou. Waste Manag 34(11):2414–2423 Gurski NP (1995) Household hazardous waste: an analysis of two approaches. Bachelor of Science. Faculty of the Graduate College, Oklahoma State University. University of Scranton, Pennsylvania, USA Hagiwara M, Watanabe E, Barrett JC, Tsutsui T (2006) Assessment of genotoxicity of 14 chemical agents used in dental practice: ability to induce chromosome aberrations in Syrian hamster embryo cells. Mutat Res 603:111–120 Hamer G (2003) Solid waste treatment and disposal: effects on public health and environmental safety. Biotechnol Adv 22(1–2):71–79 Hoffmann FO, Hammonds JS (1994) Propagation of uncertainty in risk assessments: the need to distinguish between uncertainty due to lack of knowledge and uncertainty due to variability. Risk Anal 14(5):707–712 Inglezakis VJ, Moustakas K (2015) Household hazardous waste management: a review. J Environ Manag 150:310–321 Karpinski G, Glaub J (1994) Screening hazardous wastes in solid waste landfills. Waste Age 25(8):91–98 Kim K-H, Kabir E, Jahan SA (2016) A review on the distribution of Hg in the environment and its human health impacts. J Hazard Mater 306:376–385

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Kontogianni ST (2013) Hazard analysis of specific waste streams included in municipal solid waste. PhD in Mechanical Engineering. Aristotle University of Thessaloniki Kontogianni ST, Karagiannidis A (2009) Assessing occupational health and safety factors in Hellenic waste management facilities. In: Cossu R, Diaz LF, Stegmann R (eds) Proceedings of the 12th international waste management and landfill symposium. S. Margherita di Pula (Cagliari), Sardinia, 5–9 Oct, CD-ROM edition Kontogianni S, Xirogiannopoulou A, Karagiannidis A (2008) Investigating solid waste production and associated management practices in private dental units. Waste Manag J 28(8):1441–1448 Kontogianni ST, Karagiannidis A, Logothetis D, Alexiou G (2013) Investigation of the household hazardous waste stream in a Hellenic municipality. Int J Environ Waste Manag. (Corrected proof). Incomplete Kontogianni ST, Moussiopoulos N (2017) Investigation of the occupational health and safety conditions in Hellenic solid waste management facilities and assessment of the in-situ hazard level. Saf Sci 96:192–197 Lohbauer U (2010) Dental glass ionomer cements as permanent filling materials? – Properties, limitations and future trends. Materials 3:76–96 MED Programme (2012) Zero waste project (1G-MED08-533). http://www.med-zerowaste.eu/ needsupdate Miller EG, Washington VH, Bowles WH, Zimmermann ER (1986) Mutagenic potential of some chemical components of dental materials. Dent Mater 2:163–165 Official Journal of the European Union (2000) Commission decision of 3 May 2000 replacing decision 94/3/EC establishing a list of wastes pursuant to Article 1(a) of Council Directive 75/442/EEC on waste and Council Decision 94/904/EC establishing a list of hazardous waste pursuant to Article 1(4) of Council Directive 91/689/EEC on hazardous waste (notified under document number C (2000) 1147). No 2000/532. L226: 3–24 Official Journal of the European Union (2008a) Directive 2008/98/EC on waste and repealing certain directives. L 312; p 3–30 Official Journal of the European Union (2008b) Regulation (EC) No 1272/2008 of the European Parliament and of the Council of 16 December 2008 on classification, labelling and packaging of substances and mixtures, amending and repealing Directives 67/548/EEC and 1999/45/EC, and amending Regulation (EC) No 1907/2006. L353: 1–1355 Poulsen C, Dandanell G (1995) Identification and characterization of genes (xapA, xapB, and xapR) involved in xanthosine catabolism in Escherichia coli. J Bacteriol 177:5506–5516 Sanida G, Kallergis K, Karagiannidis A, Rakimbei P, Xirogiannopoulou A (2004) Infectious hospital waste management at the 2nd regional health and welfare authority of Central Macedonia. In: Proceedings on protection and restoration of the environment VII, Myconos, 28 June1 July, CD-ROM edition Schrab GE, Brown KW, Donnelly KC (1993) Acute and genetic toxicity of municipal landfill leachate. Water Air Soil Pollut 69:99–112 Slack RJ, Gronow JR, Voulvoulis N (2005) Household hazardous waste in municipal landfills: contaminants in leachate. Sci Total Environ 337(1–3):119–137 Slack RJ, Gronow JR, Voulvoulis N (2009) The management of household hazardous waste in the United Kingdom. J Environ Manag 90:36–42 Soltani A, Sadiq R, Hewage K (2017) The impacts of decision uncertainty on municipal solid waste management. J Environ Manag 197:305–315 Turner GPA (1980) Introduction to paint chemistry and principles of paint technology, 2nd edn. Chapman and Hall LtD, New York Van Landuyt KL, Snauwaert J, De Munck J, Peumans M, Yoshida Y, Poitevin A, Coutinho E, Suzuki K, Lambrechts P, Van Meerbeek B (2007) Systematic review of the chemical composition of contemporary dental adhesives. Biomaterials 28:3757–3785 Washington State Dental Association (WSDA) (2004) Best management practices for dental office wastes. Department of Ecology, State of Washington, HWTR Publication, Publication No. 0604-007, Washington, DC Wolf AA, Kettler LE, Leahy JF, Spitz AH (1997) Surveying household hazardous waste generation and collection. Environ Health 59(7):6–11

Management of Hazardous Paper Mill Wastes for Sustainable Agriculture

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Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Raw Materials Used in Indian Paper Industry . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Generation of Wastes in Paper Mill . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Quantity and Quality of Paper Mill Solid Wastes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Harmful Impact of Paper Mill Wastes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Reuse of Paper Mill Solid Wastes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Mass Utilization of Paper Mill Waste Through Low-Cost Composting Technique . . . . . . . . . . . Advantages of Vermicomposting for Sustainable Agriculture . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Social Benefits of Vermicomposting of Paper Mill Waste . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cross-References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

The papermaking has far-reaching environmental impacts due to the production of wastes in the form of sludge since that is creating major problems to soil, crop, and human health due to the presence of different toxic chemicals like arsenic, cadmium, chromium, lead, mercury, nickel, etc. India has about 700 pulp and paper mills, which are generally considered as most polluting industries. Different paper mills use different raw materials like bamboo, wood, paddy straw, waste papers, etc. The quality of effluents also varies with quality of inputs/raw materials as well as technical procedures followed in papermaking. Paper mill wastes are the residuals constituted with fibers produced from the pulping of raw materials but unsuitable for papermaking, paper sludge, lime sludge, inks, fly ash, clays, and other fillers developed in de-inking process. These solid wastes in some cases are A. K. Sannigrahi (*) Proof & Experimental Establishment (PXE), Defence Research and Development Organization, Balasore, India © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_17

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utilized for energy generation, but in many places those are disposed of by dumping on land by the side of effluent channel. Management of these solid wastes for beneficial utilization is a great challenge not only to the management authority of paper industry but also the scientific community for finding out some amicable solutions. Since this solid waste contains a considerable quantity of organic matter, this must be utilized for some beneficial purposes. Vermicomposting is the rapid as well as cost-effective treatment method for recycling the organic solid waste through the action of earthworms. The worms act as an aerator, a grinder, a crusher, a chemical degrader, and a biological stimulator. Worms decompose the organic fraction in the sewage sludge, mineralize the nutrients, ingest the heavy metals, and devour the pathogens (bacteria, fungi, nematodes, and protozoa). Paper mill wastes are not easy material to compost and normally need both structural and nitrogen amendments using different organic matters like sawdust, fruit and vegetable wastes, food processing industry’s wastes, leaf litters, cow dung, pig wastes, poultry wastes, water hyacinth, municipal solid wastes, etc. before composting. Vermicomposting of paper mill wastes is not only a value addition to problematic wastes but also its beneficial utilization for sustaining organic agriculture. It helps to reduce the problem of environment pollution and creates an opportunity for employment generation. Keywords

Earthworm · Energy recovery · Hazardous waste · Heavy metals · Landfill · Land spreading · Paper mill wastes · Pulp and paper mill sludge · Social benefits · Sustainable agriculture · Vermicomposting · Waste quality

Introduction The knowledge on paper manufacturing process, though developed in China in 105 AD, spread to other countries through silk and trade routes and reached to India around 605 AD. The evolution of papermaking and creation of paper products occurred over centuries, not in overnight. The manufacturing process got momentum in global arena in 1282 AD with the invention of paper mill. British literature revealed about the early establishment of different categories of paper mills in India during the sixteenth to nineteenth century (Ramaseshan 1989). The century-old paper industry occupies an important position in the Indian economy for its extended role of early industrialization and social sector development. At present Indian paper industry is among the top 15 global players with total installed capacity more than 15 million tons. According to IPMA (2016), this industry has grown from just 17 mills in 1951 with a capacity of 0.14 million tons to 825 mills now with a capacity of 15.3 million tons. The total installed capacity of the paper industry has grown at a compounded annual growth rate (CAGR) of 6% over the past decade. The changing policy focus of Indian government, like economic liberalization in 1991, de-licensing of pulp and paper sector in 1997, allowing 100% foreign direct investment (FDI), etc., has changed the growth performance of India’s paper industry, resulting sharp rise in the

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share of India to the world’s total production of paper and paper products from 0.68% in 1981 to 2.6% in 2016. The number and capacity of industry producing paper in India during 2001 and 2011 are mentioned in Table 1. Five percent large-scale mills contribute to 28% of the installed capacity, while medium- and small-scale industries contribute only 63% and 9% of the total installed capacity (CSE 2013). In 2012, around 88 paper mills with capacity higher than 50,000 tons per annum (TPA) contributed to around 60% of the total paper production in India. About 70% of the total paper production in India is accounted by paper mills situated at Gujarat, West Bengal, Orissa, Andhra Pradesh, Karnataka, and Maharashtra, while 25% of paper is produced by mills situated at Uttar Pradesh, Tamil Nadu, Haryana, Kerala, Bihar, and Assam. The estimated turnover of the industry is approximately INR 50,000 crore (USD 8 billion), and its contribution to the exchequer is around INR 4500 crore. These paper industries provide employment to more than 0.5 million people directly and 1.5 million people indirectly (IPMA 2016). The paper industry is classified into two main segments – paper and paperboard (comprises of writing, printing, packaging, specialty, and tissue paper) and newsprints (comprises of newspapers, flyers, and other printed materials intended for mass distribution). Similar to the production of paper, the domestic consumption of paper has also been increased from mere 0.83% in 1981 to 2.47% of the world’s total consumption in 2010. During the last 26 years (1990–2016), domestic consumptions of newsprint, printing and writing papers, and packaging papers and board have been increased from 3.95 to 25.4 lakh tons, 9.15 to 46.38 lakh tons, and 9.22 to 75.02 lakh tons, respectively (Fig. 1). According to the Indian Paper Manufacturers Association Table 1 Distribution of paper mills in terms of number and capacity Year 2001 2011

Small scale (100,000 TPA) 11 (1.9%) 32 (5%)

Average capacity (Tons/annum) 14,200 22,100

Source: CSE Survey Report (2013)

75.02

Printing + writing papers

News print Wrapping + Packaging paper + Board

49.4

46.38

41.15 25.4 16.31 18.16 9.15 9.22 3.95

1990

16.52

7.5

2000

2010

2016

Fig. 1 Trend in India’s domestic consumption of paper (Lakh tons) (Source: IPMA survey)

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Table 2 Indian paper industry market (production, consumption, export, and import of paper) (million tons) Group Domestic production Exports Imports Domestic consumption

2011–2012 10.1 0.5 2.1 11.7

2012–2013 10.5 0.5 2.0 12.0

2013–2014 11.4 0.5 2.2 13.1

2014–2015 12.2 0.6 2.3 13.9

Source: IPMA

(IPMA), the major factors for India’s robust paper demand are its economic growth, increasing literacy rate, rapid urbanization, improvement in living standards, quick change in lifestyles, better-quality consciousness in publication systems (magazines, multicolor printings, advertising), etc. As estimated by IPMA, the demand for domestic consumption of paper in India will go up to 20 and 23.5 million tons by 2020 and 2024, respectively. Table 2 showed that the demand for domestic product is increased from 11.7 to 13.9 million tons within 2011–2012 to 2014–2015, which will further increase to 23.5 million tons by 2024–2025. The per capita consumption of paper has also been increased from 9.18 kg in 2009–2010 to around 11 kg at present which is still lower than the present global average of 56 kg and the Asian average of 40 kg. Increase of per capita paper consumption by 1 kg will increase the demand by about 1.25 million tons per annum. Since the domestic demand is on the rise, the fate of Indian paper industry is more promising in the future.

Raw Materials Used in Indian Paper Industry Indian paper industry utilizes a wide variety of cellulosic raw materials like wood, bamboo, bagasse, wheat straw, rice straw, jute sticks, grasses, waste papers, etc., and paper mills can be categorized into three groups on the basis of used raw materials: wood-/forest-based mills, agriculture residue-based mills, and waste paper-based mills. Initially woods and bamboos, the conventional raw materials, were used in large quantities (84% in 1970) for manufacturing of paper. But the share of these forestbased raw materials to the total raw material consumption declined over the years in India due to their fast depletion, gradual price hike, and increasing prohibitory regulations (Fig. 2). Agricultural residues like bagasse, wheat and rice straw, jute sticks, and grasses gained importance, and small- and medium-sized agro-based paper mills were encouraged by the government for production during the 1980s and 1990s. Use of agricultural residues instead of forest-based raw materials was not only sustainable for paper manufacturing but also helped to reduce the stress of paper production considerably on endangered forests. Pulp formation from agricultural residues takes less time to cook than wood pulps, resulting in the requirement of less energy, water, as well as chemicals. But the quick decline in share of agricultural residues on paper production after 2000 might be due to (a) shorter fiber length, fiber width, excessive amount of primary fines, and low drainability of agricultural residues like bagasse and straw as

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Management of Hazardous Paper Mill Wastes for Sustainable Agriculture 100% 90%

7 9

80%

30 47

70%

70

60% 31

50% 40%

84

22

30% 20%

39

10% 0%

683

1970 Forest based

2000 Agri residue based

10 31 2011

20 2016

Waste paper based

Fig. 2 Raw material consumption pattern in India (Source: DIPP Annual report 2012 and 2016)

compared to wood fibers (Mathur et al. 2009); (b) own constraints of agricultural residue like its seasonal availability, high transportation costs, and more investment in pollution control equipments; and (c) more emphasis on environmental protection by recycling wastes and latest development of technologies to use waste paper in place of woodbased raw materials and encouragement from Indian government in the establishment of a large number of small-sized recycled fiber-based pulp and paper mills for meeting the rising paper demand of domestic consumers. Pulp and paper industry is the third largest water-consuming industrial sector in India. Freshwater consumption in waste paper mills (75–100 m3/ton of paper) is also comparatively lesser than that in wood-based (125–200 m3/ton of paper) and agro-based paper mills (125–225 m3/ton of paper) (Chakrabarti 2006). All these factors have helped waste paper in meteoric rise of its share in utilization in Indian paper industry from mere 7% in 1970 to 70% in 2016 as compared to other two groups. In terms of present share in total paper production, approximately 24% are based on wood, 11% on agro-residues, and 65% on recycled fiber (IPMA 2016). India’s domestic collection of recycled paper is insufficient to meet the needs, since a considerable quantity of old newspapers are used for various domestic purposes like wrapping of materials; supply of grocery items in packets; packaging of foods, fruits, snacks, etc.; and also glassmaking. India imports almost half of its recycled paper mainly from the USA, Latin America, and Indonesia. India imports nearly 400,000 tons of pulp, 20,000 tons of recycled paper, and about 2500 tons of finished paper annually (Mathur et al. 2009).

Generation of Wastes in Paper Mill Generally, the paper industry is considered as a major consumer of natural resources (forest products, water) and energy (fossil fuels) and a significant contributor of pollutant discharges in the form of gaseous, liquid, and solid wastes to the environment (Hossain and Ismail 2015). The paper industry typically generates large quantities of

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wastewater which are a collection from five different unit processes like debarking or raw material preparation, pulping, bleaching, washing of pulps, and papermaking. Table 3 summarizes the main pollutants produced during those five processes. The dark brown-colored effluent is heavily loaded with organic compounds depending on the nature of the raw material, process, and chemicals used, finished product, and extent of water reused. It has high biochemical oxygen demand (BOD), chemical oxygen demand (COD), chlorinated compounds as adsorbable organic halides (AOX), suspended solids mainly fibers, dissolved lignin and its degradation products, hemicelluloses, fatty acids, resin acids, tannins, phenols and sulfur compounds, etc. (Ali and Sreekrishnan 2001). Various chemical additives are used in paper industry either as “product aids” for improving the product properties or as “process aids” for enhancing the productivity of paper machine. Majority of the “product aids” do not end up in effluent (barring dissolution losses), while most of the “process aids” reappear in effluent (barring retention aids). Overall pollution effects of paper mill effluent are (a) oxygen depletion in the receiving water body; (b) the presence of undesirable color, odor, and taste in the water; (c) reduced photosynthesis; (d) formation of blanket of suspended solids settling at the bottom of the receiving body of water; (e) death of fish; and (f) toxicity added to the aquatic life due to the formation of mercaptans, pentachlorophenol, sodium pentachlorophenate, etc. (Kakahi et al. 2011). More and more public awareness about the fate of these pollutants and stringent regulations made by various authorities are forcing the industry to treat effluents to the required compliance level before discharging into the environment (D’Souza et al. 2006). In many modern mills, reduced inputs of toxic chemicals and improved wastewater treatment have resulted in significant reduction of effluent toxicity (Heuval and Ellis 2002) and of environmental impacts (Sandstrom and Neuman 2003). The treatment of effluent from paper mill is, however, always mill-specific and complex. Wastewater treatment methods followed by pulp and paper mills are chemical coagulation, flocculation, sedimentation, flotation, filtration, and different biological methods like stabilization pond, trickling filters, activated sludge process, anaerobic digestion, etc. Table 3 Major pollutants released from paper- and pulp-making process Process stages Raw material preparation Pulping

Wastewater (v) Low

Pollution load Low

Low

High

Bleaching

High

High

Papermaking

Depends on recycling effluents

Low

Source: Hossain and Ismail (2015)

Effluent contents Suspended solids including bark particles, fiber pigments, dirt, grit, BOD, and COD Color, bark particles, soluble wood materials, resin acids, fatty acids, AOX, VOCs, BOD, COD, and dissolved inorganics Dissolved lignin, color, COD, carbohydrate, inorganic chlorines, AOX, EOX, VOCs, chlorophenols, and halogenated hydrocarbons Particulate wastes, organic and inorganic compounds, COD, and BOD

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The gaseous emissions from paper mills mainly originate from energy generation (steam and electricity). The irritating odor caused by lower organic acids and reduced sulfur compounds like H2S, methylmercaptan, dimethyl sulfide, etc. developed from effluent treatment plants, from coal storage yard, and from in-plant process water circulation system creates the air pollution problem in the neighborhood of paper mills. Annoying odors may also be found in paper mills that have less water circuit closure.

Quantity and Quality of Paper Mill Solid Wastes In pulp and paper industry, different types of solid wastes and sludges are generated. The types and amounts of wastes vary greatly among paper mills depending on the raw materials used, the paper production process followed in the mills, type and grade of paper produced, and wastewater treatment technologies followed. These wastes include bark wood residues, screening rejects, different types of wastewater treatment sludge, lime sludge, de-inking sludge, fly ash from boiler operations, etc. (Monte et al. 2009; Yadav and Madan 2013; Bajpai 2015). The rejects consist of sand, bark, portion of raw materials undesirable for papermaking, impurities like staples, metal from ring binders, plastic and paper constituents as fillers, sizing agents, and other chemicals. Green liquor sludge or dregs or lime sludge is inorganic sludge separated from the chemical recovery cycle. De-inking sludge contains mainly short fibers or fines, coatings, fillers, ink particles, extractive substances, and de-inking additives. Wastewater treatment sludge is the combination of primary and secondary or biological sludge. Primary sludge, mostly fines and fillers, is generated in the clarification of process water by air flotation technique in primary tanks. Secondary sludge is generated in the clarifier of the biological units of wastewater treatment. The primary sludge can be dewatered relatively easier, while secondary sludge is very difficult to dewater. Kraft pulp process produces about 100 kg waste/air dry ton (Adt) of pulp product, while semichemical and mechanical processes produce about 60 kg waste/Adt (IPPC 2001). CEPI (2006) reported that 11 million tons of waste was generated during the production of 99.3 million tons of paper in Europe in 2005 indicating 11% waste generation during paper production. Kay (2002) mentioned about variation of waste generation quantity with production of different paper grades from recycled fiber (Table 4). Joyce et al. (1979) reported about the generation of 40–50 kg dry sludge during the production of 1 ton paper at a paper mill in North America, of which approximately 70% was primary sludge and 30% was secondary sludge (Elliot and Mahmood 2005). Generation of de-inking sludge also depends on types of paper produced in the mill from recycled fiber, 20% in a newsprint mill to 40% in a tissue mill (Bajpai 2015). Table 5 presents the approximate amount of solid waste generation at different sources in paper mill. Although the solid waste generation in paper mill is unavoidable, its quantity can be reduced by adopting better in-plant control measures, good housekeeping, and creation of general consciousness at all levels.

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Table 4 Waste generated through the production of different paper grades from recycled fiber Paper grade Packaging paper Newsprint Lightweight coated paper/super calendered paper Tissue and market pulp

Solid waste (dry basis, kg/Adt) 50–100 170–190 450–550 500–600

Source: Kay (2002) Table 5 Solid waste generation in paper mill Solid wastes Wood- and bamboo-cutting wastes Bark residue Bagasse pith Straw dust, knots, fines Cyclean rejects Lime sludge Sludge from effluent treatment plant Coal fly ash

Approximate quantity 40% of raw materials 8–15% 12–15% of raw materials 6.5% 30–50 kg/ton of paper 0.45–0.65 ton/ton of paper 40–50 kg/ton of paper 0.32–0.37 ton/ton of coal burnt

Harmful Impact of Paper Mill Wastes The papermaking process is complex with the use of many toxic chemicals as both “product aids” and “process aids” and has far-reaching environmental impacts due to the production of wastes that are creating major health problems and environmental degradation. Presence of high organic content not only creates the problem of high BOD and COD but also creates immobilization of nitrogen in soil due to high C/N ratio showing deleterious effect on plant growth. Since conventional bleaching using chlorine and its compound released large amount of carcinogenic dioxins to the environment, totally chlorine-free bleaching has been started in modern paper mills where peracetic acid, ozone, hydrogen peroxide, and oxygen are used for avoiding this chlorine-related problem. Modern paper mills are also applying some new technologies like “organic solvent pulping” using ethanol, methanol, etc.; “acetic acid pulping”; “biopulping” using microbial enzymes like xylanases, pectinases, cellulases, hemicellulases, ligninases, etc.; and “biobleaching” using white rod fungi, etc. instead of conventional pulping and bleaching process for reducing the chemical load in wastes. Paper mill wastes contain a considerable quantity of heavy metals like cadmium, lead, mercury, vanadium, chromium, nickel, copper, zinc, and arsenic causing a great concern to the environmentalists. Cadmium, lead, and mercury got most attention due to their relative toxicity. Some of the heavy metals, especially cadmium, can deposit in the soil and be absorbed by plants. Mercury can be transformed into methyl mercury in sediments and be accumulated in the food chain, especially through water. Excessive levels of heavy metals can provoke a number of health diseases. Excessive amounts of lead and mercury are especially

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dangerous with regard to the damage to the nervous system and fatal life. Mercury exposure at high levels can harm the brain, heart, lungs, and immune system of people of all ages. Lead can also give cardiovascular diseases and anemia. Excessive amounts of cadmium can damage the kidney after long-term exposure. Zinc is also a heavy metal that damages the human body. Luqman et al. (2013) reported the higher values of various heavy metals in the paper mill effluent. Abdullah et al. (2015) found that the concentration of heavy metals in the paper mill sludge varied among different paper mills in Malaysia, but except chromium the concentrations of other heavy metals like copper, lead, and cadmium were considerably higher than their availability in agricultural soils of Malaysia. The higher content of copper in paper mill sludge might be the result of the addition of cyan ink as additives during papermaking. Tucker and Douglas (2001) also recorded elevated copper levels in all de-inking sludges due to the selective concentration of copper from the inks into the sludge on de-inking. The presence of toxic concentration of mercury, cadmium, and lead was reported by Ahirwar et al. (2015) in the effluent of Orient Paper Mills, Amlai (Madhya Pradesh, India). Kumar and Chopra (2016) also recorded higher values of different heavy metals in the effluent of Uttaranchal Pulp & Paper Mill, Roorkee (Uttarakhand, India), which might be due to the application of various dyes in the manufacturing of color papers.

Reuse of Paper Mill Solid Wastes The modern sustainable management of any production process is highly concerned about reduction in generation of wastes, decreasing its hazardous nature by adopting clean technologies and effective disposal of its waste. As mentioned above paper industry is generating enormous quantities of solid waste, over 70% of which is presently disposed of in landfills (Yadav and Madan 2013). Even today in many paper mills situated in different parts of the world, solid waste is generally disposed of in open dumps or poorly designed landfills causing health hazards by groundwater contamination. The problems associated with its landfilling are the large volumes involved, the production of considerable quantity of greenhouse gases in landfill, and the possibility of hazardous substances leaking into the environment (Monte et al. 2009). Legislation for protection of environment, decreasing landfill space, rising landfill costs, and increasing public awareness on the negative effects of disposing organic material to landfill have drawn more attention toward alternative strategies for disposal of paper mill wastes. Many innovative approaches have been made for conversion of paper mill solid waste into useful resource material, but due to very small market demands, those could not become successful in diverting this waste from landfill disposal. Efforts have been made to use this waste for energy recovery, for soil quality enhancement through land spreading, for utilization in brick and cement industry, and for converting that to some beneficial materials like oil adsorbent material, heat insulation material, landfill capping material, etc. Energy recovery from paper mill waste (both rejects and different sludges) using different thermal processes is very common in Europe. Those processes are efficient

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incineration at 850  C or more in fluidized bed and circulating fluid bed combustion chambers for power and steam generation, sophisticated pyrolysis in the absence of oxygen to produce a mixture of gaseous and liquid fuel, indirect steam gasification following pulse combustion technology, wet oxidation at temperatures of 150–330  C and pressure 1–22 MPa for production of carboxylic acids, super critical water oxidation at 400–600  C for converting wastewater to reusable water and CO2, and gasification at 1000–1400  C to produce inflammable gas. High moisture (30–50%) in paper mill sludge is the limiting factor for energy recovery. Dewatering of sludge is carried out through centrifugation, band filters, filter presses, and screw presses. Sludge containing a high level of lignin is easier to dewater than other types of sludge. Monte et al. (2009) described about advantages and disadvantages of above thermal process for wastes of paper industry (Table 6). Kunzler (2001) showed that secondary Table 6 Advantages and disadvantages of various thermal processes for energy recovery from paper mill wastes Thermal process Incineration

Advantages Reduction of the amount of residues to be landfilled Nearly complete elimination of the organic materials Possible applications for the ashes obtained

Pyrolysis

Steam reforming

Non-burning process

Production of mixture of gaseous and liquid fuels and a solid inert residue Can be sited at most existing plants Minimization of air, land, and water pollution Conversion of all sludge biomass fractions into useful energy Volume reduction by as much as 90% and production of sterile carbon char Higher heat transfer rate Low NOX emission Low operating and maintenance costs Production inhibition of dioxins and furans Vaporization minimization of toxic metals

Disadvantages Incineration process of high moisture content paper mill wastes can be energy deficient Air pollution problems (NOX and SO2 emissions) Toxic metals in residue ashes Source of chlorinated compounds High cost due to increasing demand on the flue gas cleaning A consistent waste stream such as tires and plastics is required to produce a usable fuel product This technology based on the recovery of liquid fuel requires low moisture in the sludge (10%; adequate P, K, and trace minerals; pH 6.5–8; moisture 50–60%; and free air space of around 30%. C/N ratio lower than 20:1 may create undesirable odors due to release of ammonia gas, while C/N ratio higher than 40:1 will provide insufficient nitrogen for optimal growth of microbial populations, thus slowing the rate of composting. Kunzler (2001) reported that about 300,000 tons of paper mill wastes were composted by US paper mills in 1999. Thyagarajan et al. (2010) successfully converted the secondary treatment sludge of a pulp and paper industry at Karur, Tamil Nadu, into value-added compost in the period of 90 days by aerobic co-composting with sawdust in the ratio of 3:1 and by mixing with cow dung + effective microorganisms as inoculums. Champagne and Westman (2012) also recorded effective composting within 21 days by mixing three parts biosolids of St Marys Papers Ltd. (Canada) with bulking agents like one part wood chips and one part sawdust and bringing C/N ratio from 65 to 30 using nitrogen amendments like ammonium nitrate and chicken manure. Thermophilic conditions sustained for 12 days reaching a maximum temperature of 52  C and final C/N ratio of 21. Aerobic composting is always preferable as it emits only CO2, while anaerobic decomposition helps in emission of greenhouse gases like methane and nitrogen dioxide. Vermicomposting is a rapid method of composting where complex mechanical and biochemical transformation of organic waste takes place through the symbiotic action of both epigeic earthworms and decomposer microorganisms. Earthworm plays the role of an aerator, a mixer (by mixing calcium with soft organic materials at esophagus), a grinder (by grinding soft neutralized organic matter in muscular gizzard), a chemical degrader (by decomposing ground organic particles with the help of proteases, lipases, amylases, cellulases, and chitinases secreted in the

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intestine), and a biological stimulator (by accelerating the population of beneficial microbes grown in its gut). Singleton et al. (2003) found bacteria species as efficient decomposer like Pseudomonas, Mucor, Paenibacillus, Azoarcus, Burkholderia, Spiroplasma, Alcaligenes, and Acidobacterium associated with the intestine and vermicasts of Lumbricus rubellus earthworms. Similar to other organic wastes, earthworms decompose the organic fraction of pulp and paper mill wastes, mineralize the nutrients, ingest the heavy metals, and devour the pathogens (harmful bacteria, fungi, nematodes, and protozoa). Earthworms release coelomic fluids having antibacterial properties and destroy all pathogens in the waste biomass (Pierre et al. 1982). Cardoso and Remirez (2002) reported a 90% removal of fecal coliform and 100% removal of helminthes from sewage sludge by earthworms after vermicomposting. Earthworm species like Eisenia fetida, Eisenia andrei, Eudrilus eugeniae, Perionyx excavatus, Lumbricus rubellus, etc. are best suited for vermicomposting of sludge. Promising results recorded by different researchers on vermicomposting of pulp and paper mill wastes using various earthworms and structural amendments are discussed in details below. Idea of vermicomposting of paper mill sludge was started by an initial experiment of Butt (1993) in which he investigated two species of earthworms, Lumbricus terrestris L and Octolasion cyaneum, fed on a paper mill sludge mixed with yeast extract as a nitrogen source and found that earthworm response to a particular feed was highly species dependent and C/N ratio around 25 was the optimum. Elvira and Dominguez (1995) observed that Eisenia andrei earthworms suffered weight loss when those were fed with only paper mill sludge collected from the National Cellulose Company Ltd. in Spain but they grew dramatically when raw sludge was supplemented by rabbit manure, sewage sludge, or pig or hen slurries, indicating the possible impact of improved nutrient balance and better microbial populations on their growth. In fact raw paper mill sludge is highly nutrient deficient for sustaining vermiculture, and hence its blending with other nutrient-rich materials is a prerequisite. During vermicomposting of paper mill sludge mixed three parts to one with primary sewage sludge using Eisenia andrei earthworms, Elvira et al. (1996) recorded significant loss of CO2 from substrate, increase in total nitrogen, lowering of C/N ratio, and mineralization of metals in the worm-worked product than in the control (composted without worms). Vermicomposting is a low-temperature process, but pathogen reduction requires good thermophilic phase with rising of temperatures about >55  C for more than 3 days. Ndegwa and Thompson (2001) found that composting followed by vermicomposting of sewage sludge blended with paper mulch was a better treatment option than vermicomposting alone in raising the required temperature of thermophilic phase. Sannigrahi and Sannigrahi (2006) also confirmed that two-stage composting technique, i.e., initial aerobic composting for about a month in open space followed by vermicomposting under shade, was the best technique for rapid vermicomposting of different organic wastes using Eisenia fetida. Raising the height of aerobic composting bed to >1 m and covering materials with black polythene sheet helped to raise high temperature (>55  C) inside the bed during thermophilic

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phase. In an effort to utilize environmental problematic both paper mill waste and tomato plant debris in vermicompost preparation, Fernandez-Gomez et al. (2013) found that 2:1 mixture ratio of tomato plant debris and paper mill sludge was the most suitable feed ratio for optimum growth and reproduction of Eisenia fetida during vermicomposting process. Yadav and Madan (2013) also prepared goodquality vermicompost from paper mill sludge of Star Paper Mills, Saharanpur (Uttar Pradesh, India), after mixing with combined wastes (mixture of agricultural waste, municipal waste, and poultry waste) in 1:3 ratio and using Eisenia fetida earthworms. Total nitrogen, available phosphorus, and total potassium increased, while percentage of organic carbon decreased with progress of vermicomposting up to 60 days. During investigation on the feasibility of using Perionyx excavatus to biotransform solid paper mill sludge of Nagaon Paper Mill (Assam) after mixing with cow dung and food processing waste (breakfast item making waste) in equal ratio, Sonowal et al. (2013) found that total nitrogen and phosphorus were increased by 58.7% and 76.1%, respectively, while total organic carbon was decreased by 74.5%. In similar type of another experiment, the same researchers observed that mixture of solid paper mill sludge with cow dung and food processing waste in 1.0:0.46:0.04 ratio was the best feed combination for vermicomposting by Eudrilus eugeniae earthworms (Sonowal et al. 2014). Quintern (2014) recommended addition of nutrient-rich municipal biosolids with pulp mill solids for highly cost-effective disposal of both wastes through successful vermicomposting. In a 90-day study to convert mixed liquor-suspended solids (MLSS) collected from Tamil Nadu Newsprint and Papers Limited (TNPL) into vermicompost using Eudrilus eugeniae earthworms, Ponmani et al. (2014) recorded best performance in a mixture of cow dung, MLSS, and leaf litter (dried and chopped leaves of mango and eucalyptus) (1:1:2 combination) with respect to higher nutrient contents and better earthworm growth. In a similar experiment, Natarajan and Gajendran (2014) also prepared vermicompost successfully by mixing solid sludge of same paper mill with bed materials (cow dung, pig waste, and water hyacinth) in the ratio of either 50:50 or 25:75 and observed that more addition of paper mill sludge in bedding material reduced its suitability for earthworm cocoon formation. Quintern et al. (2016) reported that MyNOKE in New Zealand produced every year about 20,000 tons of vermicompost by blending 30,000 tons per year dairy waste (co-decanter dewatered waste activated sludge and dissolved air flotation sludge) from Fonterra milk processing plants and 50,000 tons/year pulp and paper mill sludge from Oji Fibre Solutions Paper Mills. The Noke vermicomposting process developed by Quintern has grown from vermicomposting 2000 tons of sludge in 2008 to 200,000 tons in 2015. During vermicomposting of solid wastes of waste paper-based Emami Paper Mill at Balasore (Odisha, India) after structural amendment with different organic materials, Mohapatra et al. (2017a) found that paper mill waste (PMW) alone was not palatable to Eisenia fetida; many of them either died or moved away from the tray. Amendment of PMW with cabbage leaves was also not promising. However, the amendment of PMW with cow dung and sawdust in 1:0.5:0.5 produced good

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quantity vermicompost in comparatively less time with better survival of earthworms (Table 7). Vermicompost of this treatment was also recorded comparatively higher nutrient contents and performed better growth of paddy seedlings. In another experiment carried out for studying the role of bed preparation techniques on vermicomposting of waste paper-based Emami Paper Mill waste amended with cow dung and sawdust, Mohapatra et al. (2017b) observed that mixing together all substrates and initial turning twice at 7-day interval gave better performance in relation to vermicompost production (Table 8), growth of Eisenia fetida earthworms, and growth of paddy seedlings as compared to other bed preparation techniques by placing different substrate layer wise following six different arrangements during vermicomposting (Fig. 3).

Table 7 Vermicomposting of Emami Paper Mill wastes amended with different substrates Quantity of mixed raw materials in beds 2 kg paper mill waste (PMW) 1.5 kg (PMW) + 0.5 kg cow dung (CD) 1 kg PMW + 1 kg CD 0.5 kg PMW + 1.5 kg CD 1.5 kg PMW + 0.5 kg cabbage leaves (CL) 1 kg PMW + 1 kg CL 0.5 kg PMW + 1.5 kg CL 1 kg PMW + 0.5 kg CL + 0.5 kg CD 1.5 kg PMW + 0.5 kg paddy straw (PS) 1 kg PMW + 0.5 kg PS + 0.5 kg CD 1.5 kg PMW + 0.5 kg water hyacinth (WH) 1 kg PMW + 1 kg WH 0.5 kg PMW + 1.5 kg WH 1 kg PMW + 0.5 kg WH + 0.5 kg CD 1 kg PMW + 1 kg saw dust (SD) 1 kg PMW + 0.5 kg SD + 0.5 kg CD

Period of vermicomposting (days) Not completed

Vermicompost produced (kg) 0.35  0.10

Earthworms at harvesting Nos. (big + Total small) weight (g) 5+0 4

50

0.67  0.05

10 + 30

15

45 40

0.48  0.04 0.57  0.07

20 + 200 15 + 120

45 40

50

0.33  0.05

8+6

15

55 49

0.16  0.04 0.30  0.01

10 + 0 2 + 10

20 3

52

0.34  0.14

5 + 80

10

80

0.66  0.16

25 + 10

52

75

0.82  0.05

20 + 80

40

50

0.38  0.08

10 + 5

19

47 45

0.27  0.03 0.12  0.02

8 + 15 5 + 10

12 8

40

0.41  0.12

18 + 20

32

45

1.84  0.08

20 + 100

64

30

1.15  0.08

25 + 300

78

Source: Mohapatra et al. (2017a)

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Table 8 Vermicomposting of Emami Paper Mill wastes under different bed preparation techniques

Treatment no. T1 T2 T3 T4 T5 T6 T7

Pattern of layer of raw material [PMW (1 kg), SD (0.5 kg) and CD (0.5 kg)] from bottom to top Mixture of PMW, CD, and SD PMW – CD – SD PMW – SD – CD CD – PMW – SD CD – SD – PMW SD – CD – PMW SD – PMW – CD

Earthworms harvested Period of vermicomposting (days) 40

Vermicompost harvested (kg) 0.79  0.04

65 70 55 65 70 53

0.26 0.29 0.37 0.19 0.14 0.43

     

0.18 0.17 0.15 0.08 0.06 0.14

Nos. (big+ small) 22 + 150

Total weight (g) 50

15 10 10 7 5 10

22 15 20 8 5 12

+ + + + + +

10 8 20 5 0 15

Source: Mohapatra et al. (2017b)

Fig. 3 Glimpses of some bed preparation (Source: Mohapatra et al. (2017b))

Advantages of Vermicomposting for Sustainable Agriculture Vermicomposting of organic wastes is many times faster than other conventional composting techniques due to joint action of earthworms and microorganisms (Sannigrahi and Chakrabortty 2000). The process of composting becomes faster with time as the earthworms multiply at very rapid rate doubling their population in every 60–70 days producing new battalions of degrader worms. Different research results as discussed above have shown that large quantity of good-quality vermicompost can easily be prepared from never-ending waste materials of paper

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mills after its desired structural amendment with nutrient-rich organic wastes with the help of locally available epigeic earthworms like Eisenia fetida, Eudrilus eugeniae, or Perionyx excavatus. The harmful impact of the indiscriminate use of both chemical fertilizers and chemical pesticides to soil, water, crops, and animals, even the human being, is an open secret now. Every country is gradually adopting organic farming technique for their agriculture-horticulture cultivation and even in pasture development. There is a large requirement of manures, compost, and vermicompost for making organic farming successful. Use of vermicompost is beneficial to soil, plant, and environment. Vermicompost in soil acts as store house of nutrients and reduces various losses of vital nutrients. Application of vermicompost in soil improves different physicochemical properties of soil as well as nutrient status in soil and releases nutrients to crop roots in a controlled manner. Vermicomposting being aerobic in nature avoids development of anaerobic condition and reduces possibilities of production of different greenhouse gases from different organic wastes, which is the cause of major concern to environmentalists for landfills. Vermicomposting also reduces the load of pathogens in environment as earthworms kill or feed pathogens during vermicomposting of various organic wastes. Eisenia fetida earthworms can readily bioaccumulate high concentrations of metals including heavy metals such as cadmium, mercury, lead, copper, manganese, calcium, iron, zinc, etc. in their tissues without affecting their physiology during vermicomposting, thus beneficial to environment. Due to the presence of humic substances like humic acids, humates, gibberellins, auxins, 3-indole acetic acid, and various other substances, vermicompost promotes plant growth in multiple ways starting with faster germination, increased development of both lateral and vertical roots, increased area of root hairs, and even higher root activity (Quintern et al. 2016). Increased root area, root growth depth, and activity help in increasing nutrient uptake and access to more available soil water during drier seasons. Several researchers (Arancon et al. 2004; Roy et al. 2010; Lazcano et al. 2011; Joshi et al. 2013) recorded higher crop yield along with higher number of blossoms and flowers due to vermicompost application. Vermicompost also has the potential to suppress plant diseases and to control pests such as insects and nematodes (Edwards and Arancon 2004; Arancon et al. 2005; Edwards et al. 2007) Sustainable agriculture deals with production of food, fiber, or other plant or animal products using organic farming technique that protects the environment, public health, human communities, and animal welfare. It focuses on producing long-term crops and livestock while preserving the balanced ecological system within the environment. It helps to maintain soil quality, to reduce soil degradation and erosion, to increase biodiversity, to save water, and to maintain healthy environment. After independence India has achieved self-sufficiency in food production by increasing its net production from 50.8 million tons in 1950–1951 to 255.36 million tons in 2012–2013. India is also the second most populous country in the world with an estimated 1.2 billion people. The success of food security depends on the production and availability of food grains following sustainable agriculture. Management of paper mill wastes through vermicomposting, therefore, offers a great scope for sustainable agriculture in India.

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Social Benefits of Vermicomposting of Paper Mill Waste Vermicompost is a marketable product and offers a scope of extra income to housewives and other family members besides using it in potted plants or in kitchen garden. Pulp and paper waste solids can be combined with other high nutrient waste sources, such as municipal biosolids, to increase the nutrient value of the resulting vermicompost. The vermicomposting work is indeed a value addition to problematic paper mill wastes but also a great help for beneficial utilization of this by-product for better crop growth. Vermicomposting is an easy technique and easily understandable to all people irrespective to their age, gender, educational qualification, or cast. It creates an opportunity for future employment generation by selling the processed vermicomposts to organic farmers. Paper mill solid waste can be used as a resource material for producing good-quality vermicompost essential for better agricultural growth and for enhancement in soil fertility. This technique shows the path way for maintaining better health and hygiene of the society.

Conclusion Paper mill waste, though complex in nature, sometimes contaminated with toxic heavy metals, and problematic to healthy environment, can be managed following destructive method of incineration for energy recovery or by constructive method of vermicomposting for producing quality organic manure essential for sustainable agriculture. Human health depends on healthy food, non-contaminated water, and pollution-free environment. Low-cost easy vermicomposting of paper mill wastes plays an important role in maintaining health of the environment (soil, water, and air), crops, animals, and human being besides offering enormous scope of employment generation, provision for side income for family members, and economical upliftment to the society.

Cross-References ▶ Application of Novel Microbial Consortia for Environmental Site Remediation and Hazardous Waste Management Toward Low- and High-Density Polyethylene and Prioritizing the Cost-Effective, Eco-friendly, and Sustainable Biotechnological Intervention ▶ Biofilm-Based Systems for Industrial Wastewater Treatment ▶ Economic Assessment of Provisioning a Comprehensive Solid Waste Management System: A Case Study of Urban Agglomerations in Ganga River Basin ▶ Industrial Solid Waste Management in a Developing Country Governorate and the Opportunities for the Application of Cleaner Production Principles ▶ Investigation of the Chemical Content of Two Specific Streams in Municipal Waste: The Case of Hazardous Household Waste and Dental Waste ▶ Management of Hazardous Paper Mill Wastes for Sustainable Agriculture

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▶ Role of Earthworms in Managing Soil Contamination ▶ Soil Pollution and Remediation ▶ Techniques for Remediation of Paper and Pulp Mill Effluents: Processes and Constraints ▶ Wastewater Management to Environmental Materials Management

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Management of Municipal Solid Waste in Morocco: The Size Effect in the Distribution of Combustible Components and Evaluation of the Fuel Fractions

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A. Ouigmane, Otmane Boudouch, Aziz Hasib, and M. Berkani

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Waste Management in Morocco . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . General Data on Waste Management in Morocco . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Accompanying Measures in the Waste Sector in Morocco . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Waste Recovery: SRF Production . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Development of SRF Production in Developed Countries . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Recovery Modes of SRF . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

702 703 703 703 704 704 705

A. Ouigmane Laboratory of Spectro-Chemistry Applied and Environment, University Sultan MoulaySlimane, BeniMellal, Morocco Laboratory of Environment and Valorization of Agro-Resources, University Sultan MoulaySlimane, BeniMellal, Morocco Transdisciplinary Team of Analytical Sciences for Sustainable Development, University Sultan MoulaySlimane, Beni Mellal, Morocco e-mail: [email protected] O. Boudouch Transdisciplinary Team of Analytical Sciences for Sustainable Development, University Sultan MoulaySlimane, Beni Mellal, Morocco Environmental and Agro-Industries Processes Team, University Sultan Moulay Slimane, Beni Mellal, Morocco e-mail: [email protected] A. Hasib (*) Laboratory of Environment and Valorization of Agro-Resources, Sultan Moulay Slimane University, Beni Mellal, Morocco e-mail: [email protected] M. Berkani Laboratory of Spectro-Chemistry Applied and Environment, University Sultan MoulaySlimane, BeniMellal, Morocco e-mail: [email protected] © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_82

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Production of SRF in Developing Countries . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Initiation to a Study of the Combustible Fractions Contained in the MSW of the City of Beni Mellal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Presentation of the Study Area: The City of Beni Mellal, Morocco . . . . . . . . . . . . . . . . . . . . . . . Materials and Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Results . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Discussion of Results . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cross-References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

705 706 706 706 707 710 712 712 712

Abstract

The management of Municipal Solid Waste (MSW) is an obstacle in several local communities, seen the socioeconomic and environmental impacts that may be related to this sector. Waste generation is usually related to changing demographic figures and the improvement of living standards. The aim of this study is to search the problems of waste management in the region of Beni Mellal-Khenifra located in Morocco and to initiate a feasibility study in order to produce a Solid Recovery Fuel (SRF) from MSW. The generation of MSW in Morocco is in continuous evolution with an annual production of 5.3 million tons in urban areas (0.76 kg per capita per day) and 1.47 million tons in rural areas (0.28 Kg per capita per day). The waste management mode practiced in Morocco is disposal in controlled or noncontrolled landfill. In order to resolve problems associated with the landfilling impacts, efforts must be made to look for valorization solutions adopted for the Moroccan context. This chapter aims to initiate a study on combustible fractions in MSW to encourage investment in the SRF production which has shown benefits in developed countries. We have carried out a characterization of Beni Mellal city waste according to the MODECOM method in order to identify the effect of the particle size distribution on repartition of fuels fractions in urban waste in Morocco. The results we found show that the size less than 80 mm contains a lot of fermentable materials and that the combustible parts have almost the same distribution for size between 80 and 250 mm and more than 250 mm with abundance of PLDE, textiles, and baby diapers. Keywords

Municipal solid waste · Environmental impacts · Landfill · Solid recovery fuels · MODECOM · Fuel fractions

Introduction MSW is heterogeneous residues, which is constantly increasing as a result of changes in the population and the improvement of living standards. Thus, the impacts that may be linked to landfilling have promoted nations to look for recovery solutions. As a result, several recovery methods have replaced landfilling (composting, methanization, incineration, pyrolysis, gasification, production of

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Management of Municipal Solid Waste in Morocco: The Size Effect in. . .

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SRF, etc.). The production of SRF is a sustainable way of managing nonhazardous waste; it converts fuel fractions into energy. Several studies have been carried out in this topic in the developed countries (Wavrer et al. 2010; Federec and Compte-R 2015; Rada et al. 2008; Raggazi and Rada 2012). The SRFs are prepared from the combustible fractions contained in the rejects of mechanical-biological treatment (MBT) units, commercial waste, and nonhazardous industrial waste (Rada et al. 2008; Raggazi and Rada 2012; Nasrullahet al. 2015). Other by-products such as waste water treatment plant (WWTP) sludge and tree pruning olives can be added to SRF to improve calorific value (Tezanou et al. 2005; Lei et al. 2016; Casado et al. 2016). The use of SRF as a substitute for fossil fuels has shown several economic and environmental benefits (Porteous 2005; Psomopoulos et al. 2009; Abu-Qudais and AbuQudais 2000; Kara 2012; Chakraborty et al. 2013). In this regard, we decided to study the possibility of setting up a process for the energy recovery of the fuel fractions contained in the MSW produced in developing countries. The first step that we did is studying the effect of waste size on distribution of SRF.

Waste Management in Morocco General Data on Waste Management in Morocco Morocco has not escaped from problems of waste management. Several efforts have been made to improve the quality of waste management sector (collection, sorting, disposal, etc.). It has tried to reduce the degree of waste management risks by preparing an integrated strategy to achieve the objectives. • Techniques (typology and characteristics of waste, collection techniques adapted to the urban environment, development of landfill and recovery centers, recovery, etc.) • Socioeconomic aspect (taking into account populations living directly or indirectly from the recovery and recycling of waste, creation of new professions, etc.) In Morocco, collection of waste in urban areas is estimated to 5.3 million tons per year. The industry generates more than 1.5 million tons annually, of which 256,000 tones are hazardous waste. The health sector generates 6000 tons of medical waste per year (Data from the Ministry of the Environment in Morocco). Morocco has failed to achieve its goal of converting all its uncontrolled dumping to controlled landfills by 2015. Today, only about 48% of waste is landfilled in controlled sites. Tangier and Casablanca remain the two main black spots.

Accompanying Measures in the Waste Sector in Morocco • National Plan of Municipal Waste (PNDM) Defined in 2008, this plan was aimed in particular at setting up controlled landfills in 100% of the urban communes of Morocco by 2015. The first PNDM action plan

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implemented between 2008 and 2012 has fallen considerably. In 2012, Morocco was so far from its objectives that the Ministries of the Environment and the Interior decided in all discretion to extend the deadline to 2020. By 2016, at least three landfills have been completed and are operational in Meknes, Marrakesh, and Ifrane. They bring the number of operational landfills to 17 at a target of 75, set at 2020. Controlled landfill should therefore be around 48% today, according to our calculations based on data from the Ministry of Environment in June 2016. There are, however, two major black spots in the urban landscape: Casablanca and Tangier. Tangier produces more than 300,000 tons of household waste per year and Casablanca 1.2 million tons. The other objectives of PNDM: • Achieve a MSW collection rate of 85% in 2015, 90% in 2020, and 100% in 2030: in 2016, 85.2% of municipal household waste was collected. • Rehabilitate 100% of old landfills in 2015 (delayed to 2020): by June 2016, 26 dumps were rehabilitated or closed out of a total of 220. • Achieve a recycling rate of 20% of the waste produced in 2015 (pushed back to 2022): in 2016, it accounted for 10% mainly thanks to the work of the informal collectors who intervene in the garbage disposed cities and on the landfills themselves. – Sorting, Recycling, and Recovery The management of MSW in Morocco started by improving access and reducing environmental impacts, while ensuring landfilling in accordance with national and international standards, and by closing and rehabilitating wild dumps. Despite the orientation to landfilling mode, recycling remains the best solution for a sustainable waste management. Indeed, landfill remains an option whose environmental impacts are often difficult to control, despite the precautions taken. It consumes spaces, sometimes at the expense of productive land, and is not without environmental risks (leachate management, risk of contamination of water resources and soil, greenhouse gas emissions, etc.) Moreover, with the continuous increase in the number of landfills, the social acceptance of these controlled storage sites will be increasingly difficult. As a result, the development of waste sorting and recycling systems is undoubtedly one of the pillars of sustainable waste management and remains the solution to limit the extent of landfills. It also helps to reduce the environmental impacts of the sector by reducing the amount of waste to be disposed of or treated and saving raw materials.

Waste Recovery: SRF Production Development of SRF Production in Developed Countries SRF is a fuel derived from nonhazardous waste produced in accordance with the requirements of the European standards for SRF, specifically in accordance with EN15359.This standard provides for the classification of SRF according to an

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economic criterion (LCV or lower calorific value), a technical criterion (chlorine content), and an environmental criterion (mercury content). Five levels have been defined for each of these criteria. The production of SRF is complementary to material recovery, seeking to promote waste that cannot be recycled. Their preparation aims to meet requirements of both energy recovery and the reduction of contaminants during combustion. The search for energy sources that are more economical than fossil fuels and less greenhouse gas emitters is observed in energy-intensive industries. Renewable energies, but also waste, can constitute these new sources of supply. From nonhazardous waste, and after extraction of the recyclable fraction, the SRFs are prepared in such a way as to enable efficient energy recovery in heat and/or electricity, as a substitute for fossil energy. They are storable and have a relatively high energy potential (measured by the lower calorific value). The content of pollutants (halogenated, heavy metals,), moisture content, and particle size must be compatible with energy recovery, smoke treatment, and requirements for ash elimination. SRF can be divided into two categories: • SRF with high quality (or cement grade) (LCV > 18 MJ/kg and chlorine content 100 mm (Kodwo et al. All flow 2015) (Youb et al. 2014) >100 mm (Koledzi 2011)

>100 mm

Fuels fractions Paper and Plastic cardboard % % 37.60 16.28

Various Textiles fuels % % 34.43 11.71

46.92 63.16

25.77 22

25 9

2.07 6

20

25.07

53.83

1.32

46.5

17.13

21

15.38

Conclusion This study allows us to draw some conclusions in terms of MSW management in Morocco, which have an impact on human health and the environment given the management methods adopted by the municipalities. In this context, two scenarios can be developed; the first concerns the sorting of waste at the source in two fractions: fuel and nonfuel fraction. The nonfuel fraction will easily extract the biodegradable material and treat it biologically and the fuel part will be chained to an SRF production unit. The second scenario is to set up a grain size grid upstream of the mechanical-biological treatment units to sort the fraction with size more than 80 mm in part combustible and noncombustible. The second one will be chained to SRF unit production. The preparation of SRF does not require sorting of several fractions as other recoveries modes. It can be a sustainable solution for waste management in developing countries by reducing the stream of waste sent to landfills.

Cross-References ▶ Hazardous Waste Management with Special Reference to Biological Treatment ▶ Household Hazardous Waste Quantification, Characterization, and Management in Developing Countries’ Cities: A Case Study ▶ Investigation of the Chemical Content of Two Specific Streams in Municipal Waste: The Case of Hazardous Household Waste and Dental Waste

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Hazardous Waste Management with Special Reference to Biological Treatment

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Soumya Nair and Jayanthi Abraham

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Characteristic Features of Hazardous Waste . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Classes of Hazardous Waste . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Treatment of Hazardous Wastes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Physical Treatment Techniques . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Thermal Treatment of Hazardous Waste . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chemical Treatment Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biological Treatment Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Limitations of Treating Hazardous Wastes Using the Biological Treatment Method . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

Hazardous waste refers to any unwanted waste or refuse which may be solid or liquid in nature and is considered lethal, chemically unstable, reactive, combustible, or corrosive, posing a threat to the environment. The most commonly found hazardous waste in the household includes batteries, pesticides, dry cleaning agents, metal particles, organic solvents, lubricating oil, etc. Maximum hazardous wastes are produced by nuclear weapon and power plants. These wastes are toxic to human health when exposed via contact, breathing, or ingestion. Household hazardous wastes are less controlled than the industrial hazardous wastes as they are likely to be discarded inappropriately posing a negative impact to human health, a threat to the biodiversity and ecosystem at a large scale. These hazardous wastes may be treated chemically, physically, thermally, or biologically. The present chapter focusses on the cost-effective and eco-friendly microbial S. Nair · J. Abraham (*) Microbial Biotechnology Laboratory, School of Biosciences and Technology, VIT University, Vellore, Tamil Nadu, India e-mail: [email protected]; [email protected] © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_121

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treatment processes involved in hazardous waste management, both unaided and in combination with the existing treatment methods. The primary goal is to convert hazardous wastes to its nonhazardous counterpart so that it does not pose a danger when released in the environment, even better if it can be reused or recycled. With suitable microorganisms and under optimum environmental parameters, hazardous bio-refractory wastes can also be degraded. Furthermore, genetic engineered microorganisms could be a revolution in the microbial treatment of hazardous wastes. Keywords

Hazardous waste management · Bio-refractory · Microbial treatment · Environment · Industrial hazardous wastes · Genetic engineered microorganisms · Household hazardous wastes

Introduction Hazardous waste management (HWM) is a major challenge in all the metropolitan areas of the world. Without an adequate and economical waste management program, the waste generated from various industries, or due to anthropological activities, can lead to health jeopardy and have a negative impact on the ecosystem (ADB 2016). Uncontrolled and continuous discharge of industrial, agricultural, and civic wastes into the surrounding environmental sink has become a major global concern to be taken care of (Ash and Fetter 2004; Bullard et al. 2008; Chakraborty and Green 2014). The industrial and civic activities had also led to the contamination of the agricultural lands, which results in the loss of biodiversity. Although the use of pesticides and herbicides increases the productivity of crop, they also increase the contamination of the soil, water, and air (Gokhale-Welch 2009; Gilbert and Chakraborty 2011; Downey and Crowder 2011; Das 2015; Dixit and Srivastava 2015). Hazardous wastes can be categorized as solids, liquids, or gases. For example, paints, batteries, pesticides, polychlorinated biphenyls (PCBs), solvents, motor oil, radioactive gas, mercury, etc. As mentioned earlier, hazardous waste management is a global issue. Each year, technologically advanced nations with stringent environmental laws and regulations export more than three million tons of hazardous waste for disposal in poorer or developing nations with less stringent waste disposal laws, making these lands polluted (Haq and Chakrabarti 2000, 2007; Kathuria and Haripriya 2000; Kathuria and Khan 2007; Kathuria 2009; Kohli and Menon 2015). Hazardous wastes are known to cause serious issues for years after their improper disposal. Most of the current existing industrialized waste disposal sites were set up, dumped, and concealed long before establishing the present day standards for the waste management, which includes the collection, transport, processing, and finally the disposal of the hazardous wastes. At times, the dumping sites are themselves very difficult to locate (Ravi 2014; Taylor 2014; Sabapathy et al. 2015). These wastes that

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are toxic, highly flammable, or corrosive in nature are often persistent in the environment for a longer period of time. At times, the different risks of their improper disposal are unknown, but eventually poses a serious threat to its surroundings and indirectly to the human well-being (Ash and Fetter 2004; Anyamani 2009). The industries that are answerable for many of such hazardous waste sites prior to 1960 are no longer in business and are shut down (Planning Commission 2007; Planning Commission 2012). The current government legislation gives industries fairly a broad flexibility to synthesize chemicals, monitor and regulate their own waste disposal practices, and even to challenge cases of possible environmental or damage to human health. It is often tremendously difficult to prove a scientific link between an incident of drinking water poisoning, or a human disease cluster, and a facility that improperly handles industrial chemicals.

Characteristic Features of Hazardous Waste Hazardous wastes can be identified using a set of characteristic features. They can be classified on the basis of their nature of the waste as a whole: biological, chemical, and physical properties. These properties generate one or more of the factors below on reaction with self or other substances (Carter-Whitney 2007; Balkwaste 2010; Bernstad et al. 2011).

Classes of Hazardous Waste There are nine classes of hazardous waste as illustrated in Table 1. Some of the classes are further subdivided (EPA 2008; Lambeth 2012).

Treatment of Hazardous Wastes Hazardous waste treatment can be categorized in a number of ways. Categorization could relate to the treatment method: (I) Physical (II) Thermal (III) Chemical (IV) Biological Often waste is subjected to a combination of these methods for effective and safe disposal. Treating hazardous waste can be complex and expensive. For certain common hazardous substances, best practice treatment has already been well established through research and practical experience. For such substances, treatment options are generally easily determined and applied. However, for a number of more uncommon or problematic wastes and/or mixtures of contaminants, specialist,

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Table 1 The different classes of hazardous waste UN Class number (1989) 1

Characteristics Explosives

Types Projection, Mass explosion, Minor hazard

2

Gases

Poisonous, Flammable gas, Nonflammable gas

3

Flammable liquids

4

Flammable solids

5

Oxidizing substances

Low flash point range, Medium flash range, High flash range Flammable solids, Spontaneous combust, Emitting flammable gases when wet Oxidizing compounds of all types

6

8

Toxic and infectious substances Radioactive material Corrosives

Poisonous agents, Infectious agents Radioactive compounds of all kinds Corrosive in nature

9

Toxic

Toxic to the ecosystem or society

7

Examples Ammonium perchlorate, Cyclonite dinitrophenol, Hexanitrodiphenylamine, Explosive articles Compressed oxygen, Aerosols, Butane, Ammonia Acetone, Alcohol, Ethyl ether White phosphorus, Yellow phosphorus, Alkali metals Sodium peroxide, Potassium permanganate, Potassium super oxide Arsenic, Clinical waste Uranium Mineral acids, Organic acids, Strong bases Environmentally hazardous chemicals

experts may need to identify or develop an appropriate treatment option/process. Some contaminants may be so complex and/or difficult to treat that the best option remains to safely store the waste until shipment to an appropriate treatment facility. Safe storage and transportation in itself may be complex and specialized.

Physical Treatment Techniques Different physical treatments techniques have been deployed for treating the hazardous wastes. A waste is subjected to physical methods or processes so as to contain the hazard, to immobilize the hazardous component(s) or substance(s), and/or to prepare it for further treatment, recycling, or landfill. Physical treatment methods do not destroy the wastes altogether, instead there are conformational changes that are easier for further treatment or disposal. Conventional physical treatment methods include carbon adsorption, filtration, flocculation, distillation, reverse osmosis, and

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ion exchange (Lagrega et al. 2001; Sheth and Pandya 2001). Other physical treatment techniques include electro-dialysis, solvent extraction, separation by filtration, chemical precipitation, solidification and stabilization, chemical fixation, encapsulation, wetting, and ozonation to name a few.

Thermal Treatment of Hazardous Waste Thermal treatment methods are used at relatively high temperatures (415  C–1650  C) to break down the organic chemicals present in the hazardous waste into more simple, less toxic forms in systems with oxygen (incineration) or without oxygen (pyrolysis). Wastes are typically combusted in two stages during pyrolysis. The first stage occurs in the main chamber. The next stage occurs in the secondary chamber, where gases formed in the main chamber are burned at 976  C–1648  C. In theory, this second chamber burns off carbon monoxide and organic vapors generated in the first chamber and avoids vaporization of inorganic material. Inorganic compounds, which include heavy metals, form an insoluble residue, and those not destroyed by incineration are to be disposed. Some of the thermal treatment processes of hazardous wastes are as follows: thermal desorption, incineration, vitrification, wet and supercritical oxidation, plasma arc A, rotary kilns, cement kilns, and boilers.

Chemical Treatment Methods Chemical treatment of hazardous wastes helps us to convert highly toxic and lethal wastes to its lesser hazardous state or render it nontoxic altogether. This mode of treatment also helps in recovering beneficial and valuable by-products post the treatment process of the hazardous wastes. Hence, in this way the overall costs of waste disposal reduces tremendously. Usually, in most cases, the chemical treatment method is employed before disposing these hazardous wastes into landfills, so that the degree of contamination or pollution is reduced to an extent. During a chemical treatment process, a chemical transformation of the various waste components is brought about causing either the removal of the lethal component or in the oxidation/ reduction of the particular component of the waste. As mentioned earlier, oxidation and reduction is two of the examples of a chemical treatment process. The combination of both physical and chemical treatment helps in concentrating the effluent first followed by removal of the toxicity before it is dumped into the landfills. A number of chemical treatment methods are adopted in managing the hazardous wastes from a number of industries such as tannery, iron and steel, etc. Some of the many chemical processes are as follows: neutralization, sedimentation, pressure filtration, hydrolysis, ozonolysis, precipitation, coagulation, and ion exchange to name a few.

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The selection of a particular treatment process among the others depends solely upon the nature of the waste and the final outcome of the effluent from the waste stream.

Biological Treatment Methods Conventional biological treatment methods have been used for over a decade in industrial waste treatment and hazardous waste treatment plants. Current progress in the understanding of the existing biological processes has led the researchers to the development of new biological tools, enhancing the opportunities for bioscience applications in many areas, including the treatment of hazardous waste. The impact of these advancements on waste treatment techniques along with the process modifications and end-product substitutes are an important factor governing the entire treatment process using biological treatment methods. Biotechnology in some way or another has a direct role to play in the waste treatment plants. It is a means by which the researchers can degrade or detoxify the hazardous chemicals present in the waste effluents. Aims of using the biological treatment method over the other conventional methods are as follows: 1. 2. 3. 4. 5.

Study of degradation of recalcitrant compounds Measure of the tolerance of severe or frequently changing operating parameters Study of multicompound degradation Study of rates of degradation of heavy metals from the effluents Ability to concentrate the nondegradable compounds and using biological methods to detoxify the hazardous wastes

Recalcitrant such as benzene, halogenated compounds, and toluene have been reported to get degraded by certain isolated strains of bacteria and fungus. These organisms have been further taken forward for strain improvement and genetic manipulations. It was observed that post strain improvement, these organisms could degrade the compounds at a faster rate and more efficiently. These strains can also be used in different remedial situations, such as clean-up of oil spills, precipitating, and degrading heavy metals from the hazardous wastes. The improvement of the various conventional biological systems through the development and improvement of specific microbial strains capable of degrading multiple compounds at a given time has been proposed (Contreras-Ramosa et al. 2008). However, this approach has also many drawbacks: engineering difficulties, development of a consortium of organisms working together, etc. Development of biological pretreatment systems for waste streams has some potential for those wastes that contain one or two recalcitrant compounds. A pretreatment system designed to remove a specific toxic compound could reduce the shock effects on a conventional treatment process. In some cases, a pretreatment system may be used with other nonbiological treatment methods (i.e., incineration) to remove toxic

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compounds that may not be handled in the primary treatment system or to make them more readily treated by the primary system. In other cases, pretreatment might render a waste nonhazardous altogether. One area of research in advanced plant genetics is in the use of plants to accumulate metals and toxic compounds from contaminated soils. Current research is directed to four areas. The first area involves the use of plants to decrease the metal content of contaminated soils, through increased rates of metal uptake. Plants could be used to remove metals from the soil which gets accumulated in the plant fibre which can be disposed. The second area of development focuses on direct metal uptake in nonedible portions of the plant. For example, the development of a grain crop like wheat that could accumulate metal from soil in the nondurable parts of the plant would allow commercial use of contaminated land. A third area of research is directed toward development of crops that can tolerate the presence of metal without incorporating these toxic elements in plant tissue. Finally, research is being conducted concerning the use of plants in a manner similar to microorganisms to degrade high concentrations of hazardous constituents. Changes in process design incorporating advances in biological treatment systems may result in less hazardous waste. The development of organisms capable of degrading specific recalcitrant materials may encourage source separation, treatment, and recycling of process streams that are now mixed with other waste streams and disposed. The replacement of chemical synthesis processes with biological processes may result in the reduction of hazardous waste. Two methods of increasing the rate of chemical reactions are through higher temperatures and catalysts. One type of catalyst is biological products (enzymes) that inherently require milder, less toxic conditions than do other catalytic materials. Historically, many biological processes (fermentations) have been replaced by chemical synthesis. Genetic engineering offers opportunities to improve biological process through reduced side reactions, higher product concentrations, and more direct routes; thus, genetic engineering offers a means of partially reversing this trend. The development of new process approaches would require new reactor designs to take advantage of higher biological reaction rates and concentrations. Biotechnology also could lead to substitution of a less or nonhazardous material for a hazardous material, particularly in the agricultural field. One of the primary thrusts of plant genetics is the development of disease-resistant plants, thus reducing the need for commercial products such as fungicides. Genetic engineering to introduce nitrogen-fixation capabilities within plants could reduce the use of chemical fertilizers and potentially reduce hazardous waste generated in the manufacture of those chemicals (Wang and Chen 2006). Some methods employed in the treatment of hazardous wastes are as follows:

Activated Sludge An activated sludge treatment process refers to a multichamber bioreactor unit that makes use of extremely concentrated microorganisms in order to degrade the organic material present in the wastes and remove nutrients from it and produce a superior quality effluent. A continuous and well supply of oxygen to the treatment system is

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Preliminary treatment

Primary clarifier

Primary Sludge

Aeration tank

Secondary clarifier

Disinfection

Final effluent

Waste activated sludge

Sludge treatment and disposal

Fig. 1 Flowchart of activated sludge treatment process

required to maintain the aerobic conditions running well and to keep the activated sludge suspended.The activated sludge consists of different flocs of bacteria, protozoans, algae, and fungi which are suspended and mixed with the waste influent in an aerated tank. The bacteria utilize the organic pollutants to grow and develop. At the end, the biomass is converted to energy, water, carbon dioxide, and new cell material. Flowchart of activated sludge treatment process is shown in Fig. 1.

Microbes and Their Role in Activated Sludge There are five major groups of microorganisms generally found in the aeration basin of the activated sludge process. Some of the microorganisms with their active role in active sludge are listed in Table 2. 1. Bacteria • They are mainly responsible for the removal of organic nutrients from the waste water. • Aerobic bacteria are most commonly found in an activated sludge. 2. Protozoa • Their primary function is to remove and digest the bacteria and other organisms that are dispersed and suspended in the sludge. • Many kinds of protozoans are used in the treatment of hazardous wastes. The type of protozoans used in the activated sludge indicates the overall performance of the system. Some of them are listed below:

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Table 2 List of microorganisms playing active role in activated sludge

Microorganisms (Genus/Family) Acinetobacter Akkermansia Prevotella Acidaminococcus Cloacibacterim Megasphaera Moraxellaceae Comamonadaceae Flavobacteriaceae Rhodocyclacea Thiotrichaceacae Pseudomonadaceae

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References Fredriksson et al. 2012 Yu and Zhang 2012 Wells et al. 2014 Ivanov et al. 2010 Vandewalle et al. 2012 Mazzoli et al. 2007 Felföldi et al. 2010 Caravelli et al. 2007 Jin et al. 2011 Phuong et al. 2012 Yadav et al. 2012 Childress et al. 2014

1. Amoebae – It has a very little effect on hazardous waste treatment and dies off instantly as the amount of food decreases. 2. Flagellates – They feed primarily on soluble organic nutrients available in the sludge. 3. Ciliates – Their function is to clarify water by removing the suspended bacteria. The crawling ciliates are known to dominate the sludge. This denotes that the performance of the sludge is good. Presence of the crawling ciliates is only toward the end of the process. 3. Metazoans • These groups of microorganisms are set to dominate the older and longer age systems such as lagoons. • Metazoans are multicellular organisms, larger than the protozoans. They play very little role in the removal of the organic material from the treatment effluent. • They are also known to feed on the algal and the protozoan population. • Three common types of metazoan in the activated sludge are as follows: 1. Rotifers – Their basic function is to clarify the effluent. They are the first group to get affected in presence of toxic loads. 2. Nematodes – Feed on bacteria, protozoans, fungi, and other nematodes. 3. Tardigrades (water bear) – They are rarely found in the activated sludge. They are best known to survive in extreme conditions. This particular group of metazoans are very sensitive to the toxic load. 4. Filamentous bacteria • Their presence is observed when the operational conditions of the activated sludge system change drastically. • Growing in long filaments proves advantageous to them.

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• Change in environmental parameters such as temperature, pH, available nutrients such as nitrogen, carbon, phosphorous, etc., brings upon a change in the activity of these bacteria. • The dominance of this group in the activated sludge system can pose problems with sludge settling. As a result, the sludge may become bulky. The settling efficiency of the bulky sludge is very poor. This results in the turbid effluent. • Some filamentous bacteria may cause foaming in the aeration basin and clarifiers. 5. Algae and fungi • Their presence is observed with change in pH and age of the sludge. • They are photosynthetic organisms. • Their presence denotes poor working condition of the activated sludge. • It may also indicate the change in the pH or increase in the age of the sludge.

Aerated Lagoon An aerated lagoon is a type of biological method of treating hazardous wastes. It is a suspended-growth process in waste treatment unit. The treatment system consists of a large earthen pond, lagoon, or a basin that is equipped with automated aerators to sustain an aerobic environment and to prevent the settling of the suspended biomass. The unit is provided with an inlet and outlet at two different ends to enable the wastes to flow through it and allow it to be retained for a specific time. Initially, the population of microorganisms in the aerated environment is much lower than that in an unaerated setting because there is no scope of sludge recycles. Consequently, a considerably longer habitation time is required to attain the same waste quality. However, a longer habitation time may be an improvement when complex organic compounds are to be degraded simultaneously. Also, the microorganisms in aerated lagoons are more resilient to process setbacks caused by feed deviations than those in an unaerated setting because of the larger tank capacities and longer dwelling times used. Aerated lagoons are an effective and an economical alternate system for hazardous waste treatment in small populations. They can be incorporated very well into the surrounding backdrop. If suitable aerators are used, the power input is similar to an analogous activated sludge plant. In addition to the necessary oxygen transfer, the aerators have to be provided for proper mixing and circulation. A flow chart of a typical aerated lagoon system is depicted in Fig. 2.

Fig. 2 An aerated lagoon system

Raw waste

Lagoons with aerators

Treated effluents

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There are two types of aerated ponds: 1. Facultative ponds (common aerated lagoons) 2. Completely mixed aerated ponds The effluents from the aerated ponds may be recycled or reused. The settled sludge requires an additional treatment before its proper disposal. Microorganisms in Aerated Lagoon Microbes found in aerated and facultative ponds are more varied than those observed in other hazardous waste treatment processes. This is mainly due to the different growth parameters present in the surrounding environments. Both aerobic and anaerobic bacteria are involved along with algae and fungi. The aerobic bacteria are similar to those found in activated sludge. Four different groups are found in the aerated lagoon: (a) (b) (c) (d)

Freely distributed Single cell Floc-forming bacteria Filamentous bacteria

All the abovementioned bacteria function likewise in order to oxidize organic carbon (BOD) to produce carbon dioxide (CO2) and new bacteria. A number of filamentous bacteria are also present usually when optimum growth environment is available. Most heterotrophic bacteria function in an extensive range of environmental parameters. They play a role in removal of BOD over a wide range of pH and temperature. BOD removal takes place well in a pH range of 6.5–9.0 and at temperatures ranging from 3–4  C to 65–80  C. Efficiency of BOD removal declines rapidly below 2–4  C and terminates at 0–2  C. A very specific group of bacteria occurs to some degree in lagoons that oxidize ammonia to nitrate. This group of bacteria are strictly aerobes. Two specific bacteria are involved in nitrification. Further reports suggest that at least five classes of bacteria oxidize ammonia and at least two of the genera belong to bacteria oxidizing nitrite. Anaerobic, heterotrophic bacteria commonly occurring in lagoons are mostly involved in methane production and in sulfate reduction. Anaerobic methane production from the lagoons involve two different groups of anaerobic bacteria that function together to convert organic compounds to methane via a three-step process. Acid-forming bacteria are diverse group which convert products from the aerated lagoon under anaerobic conditions to simpler alcohols and organic acids. Examples of such organic acids are acetic and butyric acid. These bacteria are enduring and occur over an extensive range of environmental parameters, especially pH and temperature range. Anaerobic photosynthetic bacteria occur in almost all aerated lagoons and are the predominant photosynthetic microorganisms growing in anaerobic condition. The anaerobic sulfur bacteria are grouped into the red and green sulfur bacteria.

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They are represented by above 25 genera. These organisms oxidize the reduced sulfur compounds (H2S) using the light energy to produce sulfur and sulfate ions. All the anaerobic photosynthetic bacteria are either strict anaerobes or microaerophilic in nature. Most commonly found are Chromatium, Thiocystis, and Thiopedia, which grow in abundance and as a result the lagoon appears pink or red in color. Their presence in the lagoon indicates that there has been an increase in the organic material load and anaerobic environment in an anticipated aerobic system. Conversion of the odorous sulfides to its sulfur and sulfate forms by sulfur-reducing bacteria help significantly in odor control mechanism in many lagoons (facultative and anaerobic lagoons). Algae are aerobic organisms. They are photosynthetic in nature. They are capable of growing in simple inorganic compounds using light as an energy source. These organisms produce oxygen in the day hours and consume the same in the night hours. Usually, algal isolates are desirable in lagoons as they help in the generation of oxygen which is required for the stabilization of the wastes. There are three major groups of algae. They are characterized based on the color of the pigment. (a) Brown algae (diatoms) (b) Green algae (c) Red algae There is one more group that is very common in the lagoons. They are the bluegreen algae, also called as the cyanobacteria. Most of these organisms are capable of fixing nitrogen. Best studied blue-green bacteria in hazardous waste treatment systems include Aphanothece, Microcystis, and Anabaena.

Trickling Filter A trickling filter is a type of waste treatment process consisting of a fixed bed of rocks, gravel, peat, etc.; they are also referred to as biofilters or biological filter. They are operated mostly under aerobic condition. Already settled waste water is continuously sprayed over the filter bed. When the water passes through the pores of the filter, the organic compounds are degraded aerobically in a continuous process. Flow chart of trickling bed treatment process is given in Fig. 3. Rotating Biological Contactors Rotating biological contactors (RBC), also called rotating biological filters, are fixed-bed reactors consisting of stacks of rotating disks mounted on a horizontal shaft. They are partially submerged and rotated as waste flows through. They are used in conventional wastewater treatment plants as secondary treatment after primary sedimentation of domestic grey- or blackwater, or any other biodegradable effluent. The microbial community is alternately exposed to the atmosphere and the wastewater, allowing both aeration and assimilation of dissolved organic pollutants and nutrients for their degradation. Flow chart depicting the rotating biological contactor process is shown in Fig. 4.

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Primary sedimentation

Trickling filter

Secondary sedimentation

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Sludge To the digester

To primary tank

Fig. 3 Flow chart of trickling bed treatment process

Fig. 4 Flow chart depicting the rotating biological contactor process Primary effluent

Rotating biological contactors

Secondary clarifier

Effluent

Fig. 5 Flow chart of using stabilization ponds for the treatment of hazardous wastes Feed storage

Anaerobic reactor

Stabilization Stabilization Pond 1 Pond 2

Effluent

Stabilization Ponds Waste stabilization ponds (WSPs) are large, man-made water bodies in which blackwater, greywater, or fecal sludge are treated by natural occurring processes under the influence of solar light, wind, microorganisms, and algae. The ponds can be used individually or linked in a series for improved treatment. Flow chart of using stabilization ponds for the treatment of hazardous wastes in shown in Fig. 5. There are three types of ponds (1) Anaerobic ponds (2) Facultative ponds (3) Aerobic ponds

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BOD and pathogen removal is high. But, large surface areas and expert design are required. The effluent still contains nutrients and is therefore appropriate for the reuse in agriculture, but not for direct recharge in surface waters.

Packed Bed Reactors Packed bed reactors can be used in chemical reaction. These reactors are tubular and are filled with solid catalyst particles, most often used to catalyze gas reactions. The chemical reaction takes place on the surface of the catalyst. The advantage of using a packed bed reactor is the higher conversion per weight of catalyst than other catalytic reactors. The conversion is based on the amount of the solid catalyst rather than the volume of the reactor. Bioremediation Bioremediation is not only a process of removing the pollutant from the environment but also is an eco-friendly and more effective process. The pollutants can be removed or detoxified from the soil and water by the use of microorganisms, known as bioremediation. The purpose of bioremediation is to make environment free from pollution with the help of environmental friendly microbes. Bioremediation can be broadly divided into two categories, i.e., in situ bioremediation and ex situ bioremediation (Singal et al. 2000; Singh et al. 2009). In situ bioremediation provides the treatment at contaminated sites avoiding excavation and transport of contaminants. There is a biological treatment for cleaning the hazardous substances on the surface. It involves the use of oxygen and nutrient in the contaminated site in the form of aqueous solution in which bacteria grow and help to degrade the organic matter. It can be used for soil and groundwater (Gupta and Mohapatra 2003; Gosavi et al. 2004). Generally, this technique includes conditions such as the infiltration of water containing nutrients and oxygen or other electron acceptors for groundwater treatment. Most often, in situ bioremediation is applied to the degradation of contaminants in saturated soils and groundwater. It is a superior method to cleaning contaminated environments since it is cheaper and uses harmless microorganism to degrade the chemicals. Chemotaxis is important to the study of in situ bioremediation as microorganisms with chemotactic abilities can move into an area containing contaminants. So, by enhancing the cells’ chemotactic abilities, in situ bioremediation will become a safer method in degrading harmful compounds (Rahuman et al. 2000). In situ bioremediation is further subdivided into following categories (a) Bioventing (b) Biostimulation (c) Biosparging (d) Fixed biobarriers/biowalls (e) Bioaugmentation (f) Phytoremediation (g) Constructed/artificial wetland

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Ex situ bioremediation is further subdivided into following categories (a) Landfarming (b) Composting (c) Biopile/Biocells/Bioheap

(a) Bioventing Bioventing is an in situ treatment, which is a combination of oxygen (O2) and nutrients along with vapor extraction. It is a technique used to degrade any aerobically degradable compound. In bioventing, the oxygen and nutrients like nitrogen and phosphorus are injected into the contaminated site. The distribution of these nutrients and oxygen in soil is dependent on soil texture. Bioventing is pumping of air into contaminated soil above the water table through well. Enough oxygen is provided through low air flow rate for the microbes. Bioventing is more effective if the water table is deep from the surface and the area having high temperature. It is mainly used for the removal of gasoline, oil, petroleum, etc. The rate of removal of these substances varies from site to site due to the differences in the soil texture and the composition of hydrocarbons. (Singh and Tripathi 2007; Strong and Burgess 2008) (b) Biostimulation Biostimulation refers to the modification of contaminated areas to enhance the growth of indigenous microbes already present. This process may include utilizing fertilizers and other nutrients to stimulate the microbes. (c) Biosparging This approach aims to increase biological activity of the soil by increasing the O2 supply. Air is initially injected through wells, after pure O2 is injected. In biosparging, air is injected below the ground water under pressure to increase the concentration of oxygen. The oxygen is injected for microbial degradation of pollutant. Biosparging increases the aerobic degradation and volatilization. There must be control of pressure while injecting the oxygen at the contaminated site to prevent the transfer of volatile matter into the atmosphere. The cost can be reduced by decreasing the diameter of injection point. Before injecting the oxygen the soil texture and permeability is analyzed. This technology is applied to a known source of gasoline contamination in order to quantify the extent of remediation achieved in terms of both mass removed and reduction in mass discharge into groundwater. Heavier products (e.g., lubricating oils) generally take longer period to biodegrade than the lighter products, but biosparging can still be used at these sites (Vidali 2001; Lambert et al. 2009).

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(d) Fixed Biobarriers/Biowalls Fixed biobarriers use solid or viscous amendments placed across the flow path of contaminated groundwater to form a permeable reactive barrier. The fixed biobarrier approach can use engineered trenches or barriers containing solid phase, slowrelease substrates or viscous substrates placed cross gradient via direct-push injections. In situ enhanced bioremediation in the form of a fixed biobarrier is a suitable technology for large plumes having poorly defined, widely distributed, or inaccessible source areas (Funari et al. 2016). (e) Bioaugmentation Bioaugmentation refers to the addition of naturally occurring microbes to contaminated materials and sites in order to achieve bioremediation. The process ensures that the suitable microbes are added in sufficient quantities. Microorganisms having specific metabolic capability are introduced to the contaminated site for enhancing the degradation of waste. At sites where soil and groundwater are contaminated with chlorinated ethenes, such as tetrachloroethylene and trichloroethylene, bioaugmentation is used to ensure that the in situ microorganisms can completely degrade these contaminants to ethylene and chloride, which are nontoxic. Monitoring of this system is however difficult (Niu et al. 2009). (f) Phytoremediation Phytoremediation is the use of plants for removal of contaminants from soil or water. Contaminants are fixed in the ground, accumulated in the plant tissue, or released to the atmosphere. Phytoremediation consists of four different plant-based technologies each having a different mechanism of action for the remediation of soil polluted with heavy metals or water. These technologies comprise (Cobbett 2000; Cobbett and Goldsbrough 2002): (a) Rhizofiltration: It involves the use of plants to remediate aquatic environments. (b) Phytostabilization: It involves stabilization of polluted soil by using plants. (c) Phytovolatilization: Involves the use of plants to extract metals from the soil and then releasing them into the atmosphere by volatilization and (d) Phytoextraction: It involves the use of plants to absorb metals from the soil followed by translocation in the harvestable shoots. The main physiological steps in phytoremediation include: (a) Stimulation of microorganism-based transformation by plant exudates (b) Slowing of contaminant transport from the vegetated zone due to adsorption and increased evapo-transpiration (c) Plant uptake, followed by metabolism or accumulation

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Fig. 6 Uptake mechanism of pollutants and other wastes by phytoremediation process

Phytoremediation is an interdisciplinary technology that can benefit from many different approaches. Results already obtained have indicated that some plants can be effective in toxic metal remediation. The processes that affect metal uptake, availability, translocation, degradation, chelation, and volatilization need to be investigated in detail. Uptake mechanism of pollutants and other wastes by phytoremediation process is given in Fig. 6. (g) Constructed/Artificial Wetlands Wetlands constructed for remediation purposes are examples of phytoremediation. Constructed wetlands have been used for decades for the management and treatment of many wastewaters, including municipal, acid mine drainage, agriculture, petrochemical and textile industries, and storm water. However, constructed wetlands are being used for the remediation of groundwater for surface water impacted by industrial chemicals and wastes such as landfill leachate and explosives such as 2,4,6-trinitrotoluene (TNT) or 1,3,5-trinitro-1,3,5-triazinane (RDX). The trend towards increased use of constructed wetland technology relates to the low capital cost on operating and maintaining the passive technology. Ex Situ Bioremediation The treatments are not given at site. In ex situ, the contaminated soil is excavated and treated at another place. This can be further subdivided into following categories: (a) Land Farming Landfarming refers to a “low tech” biological treatment which involves the controlled application and spread-out of a more-or-less defined organic bioavailable

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waste on the soil surface, and the incorporation of the waste into the upper soil zone. It is typically used for biological removal of petroleum products from contaminated soil. In land farming make a sandwich layer of excavated soil between a clean soil and a clay and concrete. The clean soil at bottom and concrete layer should be the upper most layers. After this, allow it for natural degradation by providing oxygen, nutrition, and moisture. pH should be maintained at 7 by using lime. Land farming is also useful for pesticides contaminated sites. Land farming is a bioremediation technology. Contaminated soils are mixed with soil amendments such as soil bulking agents and nutrients, and then they are tilled into the earth. The material is periodically tilled for aeration. Contaminants are degraded, transformed, and immobilized by microbiological processes and by oxidation. Soil conditions are controlled to optimize the rate of contaminant degradation. Moisture content, frequency of aeration, and pH are all conditions that may be controlled. Land farming differs from composting because it actually incorporates contaminated soil into soil that is uncontaminated. Composting also generally takes place in aboveground piles (Maila and Cloete 2004). (b) Composting Due to its common use for household garden waste, this is the well-known controlled biological decomposition of organic material in the presence of air to form a humus-like material. Methods of composting include mechanical mixing and aerating, ventilating the materials by dropping them through a vertical series of aerated chambers, or placing the compost in piles out in the open air and mixing it or turning it periodically. Compositing is a process in which microorganism degrades the waste at elevated temperature that ranges from 55 to 65  C. During the process of degradation, microbes release heat and increase the temperature which leads to more solubility of waste and higher metabolic activity in composts. In composting rocks and other larger particles are removed from excavated contaminated soil. The soil is transported to a composting pad with a temporary structure to provide containment and protection from weather extremes. Amendments (straw, alfalfa, manure, agricultural wastes, and wood chips) are used for bulking agents and as a supplemental carbon source. Soil and amendments are layered into long piles known as windrows. Wastes are two type, i.e., inorganic waste and organic waste. The inorganic waste includes mainly heavy metals and organic waste includes agricultural waste, plastics, rubbers, etc. Researchers have found a variety of the ways by which we can degrade the solid waste. But bioremediation makes its leap to tackle the problem of heavy metals associated with different categories of waste with the help of microorganism. (c) Biopile/Biocells/Bioheaps It is a hybrid of composting and land farming. The basic biopile system includes a treatment bed, an aeration system, an irrigation/nutrient system, and a leachate

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collection system. For proper degradation, there should be control of moisture, heat, nutrients, oxygen, and pH. The irrigation system is buried under the soil. Air and nutrients are provided through vacuum. To prevent the run off, the soil is covered with plastic and due to which evaporation and volatilization is also prevented and promote solar heating. Biopile treatment takes 2 to 3 months to complete the procedure. Factors that may limit the applicability and effectiveness of the process include: • Digging of contaminated soils is required. • Treatability testing should be conducted to determine the biodegradability of the contaminants and appropriate oxygenation and nutrient loading rates. • Solid phase processes have questionable effectiveness for halogenated compounds and may not be very effective in degrading transformation products of explosives. • Similar batch sizes require more time to complete cleanup than slurry phase processes. • Static treatment processes may result in less uniform treatment than processes that involve periodic mixing. Organisms Used in Bioremediation As stated previously, bioremediation involves various microorganisms that are able to degrade and reduce toxicity of environmental pollutants. Therefore, the interactions of microbes with the environment and pollutants are significant in determining effectiveness of bioremediation. Those microbes can be either naturally present in the site of bioremediation or isolated from other sites and inoculated artificially. Biodegradation often occurs as part of microbial metabolism and in some cases, microbes are able to directly harvest carbon and energy by breaking down pollutants. Sections below go over bacteria and fungi, the commonly used organisms in bioremediation, and archaea, the more recently discovered group of organisms with unique potential in bioremediation. Bacteria

Bacteria are widely diverse organisms, and thus make excellent players in biodegradation and bioremediation. There are few universal toxins to bacteria, so they are likely to break down any given substrate, under optimum conditions (anaerobic vs. aerobic environment, sufficient electron donors or acceptors, etc.). Below are several specific bacteria species known to participate in bioremediation. (i) Pseudomonas putida: Pseudomonas putida is a Gram-negative soil bacterium that is involved in the bioremediation of toluene, a component of paint thinner. It is also capable of degrading naphthalene, a product of petroleum refining, in contaminated soils. (ii) Dechloromonas aromatica: Dechloromonas aromatica is a rod-shaped bacterium which can oxidize aromatics including benzoate, chlorobenzoate, and toluene, coupling the reaction with the reduction of oxygen, chlorate, or nitrate.

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It is the only organism able to oxidize benzene anaerobically. Due to the high propensity of benzene contamination, especially in ground and surface water, D. aromatica is especially useful for in situ bioremediation of this substance. Nitrifiers and Denitrifiers

Industrial bioremediation is used to clean wastewater. Most treatment systems rely on microbial activity to remove unwanted mineral nitrogen compounds (i.e., ammonia, nitrite, and nitrate). The removal of nitrogen is a two-stage process that involves nitrification and denitrification. During nitrification, ammonium is oxidized to nitrite by organisms like Nitrosomonas europaea. Then, nitrite is further oxidized to nitrate by microbes like Nitrobacter hamburgensis. In anaerobic conditions, nitrate produced during ammonium oxidation is used as a terminal electron acceptor by microbes like Paracoccus denitrificans. The result is N2 gas. Through this process, ammonium and nitrate, two pollutants responsible for eutrophication in natural waters, are remediated. (i) Deinococcus radiodurans: Deinococcus radiodurans is a radiation-resistant extremophile bacterium that is genetically engineered for the bioremediation of solvents and heavy metals. An engineered strain of Deinococcus radiodurans has been shown to degrade ionic mercury and toluene in radioactive mixed waste environments. In anaerobic conditions, nitrate produced during ammonium oxidation is used as a terminal electron acceptor by microbes like Paracoccus denitrificans. The result is dinitrogen gas. Through this process, ammonium and nitrate, two pollutants responsible for eutrophication in natural waters, are remediated. (ii) Methylibium petroleiphilum: Methylibium petroleiphilum (formally known as PM1 strain) is a bacterium capable of methyl tert-butyl ether (MTBE) bioremediation. PM1 degrades MTBE by using the contaminant as the sole carbon and energy source. (iii) Alcanivorax borkumensis: Alcanivorax borkumensis is a marine rod-shaped bacterium which consumes hydrocarbons, such as the ones found in fuel, and produces carbon dioxide. It grows rapidly in environments damaged by oil, and has been used to aid in cleaning the more than 830,000 gallons of oil from the Deepwater Horizon oil spill in the Gulf of Mexico Fungi (Mycoremediation)

Current bioremediation applications primarily utilize bacteria, with comparatively few attempts to use fungi. Fungi have fundamentally important roles because of their participation in the cycling of elements through decomposition and transformation of organic and inorganic materials. These characteristics can be translated into applications for bioremediation which could break down organic compounds and reduce the risks of metals. In some cases, fungi have an advantage over bacteria not just in metabolic versatility but also their environmental flexibility. They are able to oxidize a diverse amount of chemicals and survive in harsh environmental conditions such as

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low moisture and high concentrations of pollutants. Therefore, fungi are potentially an extremely powerful tool in soil bioremediation and some versatile species such as white rot fungi have been a hot topic of research. (i) Phanerochaete chrysosporium P. chrysosporium was the first fungi linked to degradation of organic pollutants. Extensive research has shown that it has strong potential for bioremediation in pesticides, PAHs, dioxins, carbon tetrachloride, and many other pollutants. Among fungal systems, P. chrysosporium has become the model for bioremediation. Other notable species of white rot fungi include Pleurotus ostreatus and Trametes versicolor. Archaea

The role of archaea in bioremediation has not been studied as commonly as that of bacteria. Nevertheless, numbers of researchers have shown their ability to degrade various pollutants and scientists began to discover more about their potential in participating in bioremediation. Below are lists of some important facts regarding archaea’s potential role in bioremediation. – Hydrocarbon-contamination is observed in some extreme environments, including hypersaline (high salt concentration), high or low temperature, or extreme pH. Archaea’s adaptation to extreme environment gives them the potential to participate in biodegradation and bioremediation in these environments; in fact, microorganisms naturally adapted to the cold environments are known to be important degraders of hydrocarbons in those environments. – Extreme halophilic archaea has potential to biodegrade pollutants in hypersaline environment, in which bacteria typically used in bioremediation cannot survive or function properly. – Some archaea are known to be resistant to variety of antibiotics, including penicillin, cycloheximide, streptomycin, etc., which gives them great advantage in participating in bioremediation in the presence of antibiotics. Examples of studies of Archaea involved in bioremediation: Four extreme halophilic strains of archaea (belonging to genus Halobacterium, Haloferax, and Halococcus) were studied to evaluate their potential to degrade crude oil and hydrocarbons (Singh and Tripathi 2007). All four strains could use various kinds of hydrocarbons as their carbon or energy sources. Two strains of Haloferax grew on n-alkanes with different lengths, ranging from C8 to C34, and also benzene, toluene, biphenyl, and naphthalene. The research demonstrated the important fact that archaea have potential to carry out biodegradation at high temperatures, in the range of 40–45  C, which is advantageous as hydrocarbons have higher solubility and bioavailability at these higher temperature. The four strains studied were resistant to six different antibiotics, including penicillin, streptomycin, and cycloheximide, and this gave them the potential to carry out biodegradation in conditions

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unfavorable for bacteria. Research suggests other genera of archaea are also capable of biodegrading in hypersaline environments. Archaeglobus fulgidus, a hyperthermophile that can use sulfate as an electron acceptor, can also break down various aromatic hydrocarbons. Microorganisms used for the bioremediation process has been listed in table below. Bioremediation of Hydrocarbon Pollutants Hydrocarbons are stored deep underground but are brought up to the surface to be transformed and utilized, primarily as an energy source known as fossil fuels. The majority of pollution currently comes from these by-products in the form of polycyclic aromatic hydrocarbons (PAHs), which are xenobiotic environmental pollutants that form when carbon materials are incompletely combusted. Some of examples of PAHs include burning wood, fossil fuels, and cigarette smoke. Currently, bioremediation is only effective for soils contaminated with low-molecular weight PAHs because of bacterial commercial use. However, fungi are effective at PAH degradation in comparison to bacteria for few reasons (Blanca et al. 2007). Firstly, they are capable of degrading PAH’s that are high in molecular weight; bacteria in comparison are better at degrading smaller molecules. Secondly, fungi can function well in nonaqueous environments and low oxygen conditions, both are conditions where PAH’s can accumulate. Many fungi have evolved mechanisms that allow them to target specific PAHs. Fungi produce extracellular enzymes that degrade lignin, a process called mineralization that produces carbon dioxide as the end product. The different methods of bioremediation used by the microorganisms for degrading pollutants have been listed in Table 3. Remediating Metals Toxic metals can enter the environment at all life cycle stages of metal compound. For example, metal leaching can occur from the mining process till the disposal of metal wastes. However in nature, the mobility of metals comes from the geological processes that can be released into the soil and aquatic environments. The environmental largest risk from metal contamination comes from the relationship between metals and compounds that are inherently of incapable of being degraded by any natural procedures. The best solution to treating contamination is transporting the metals to location where they cannot produce negative environmental effects. Fungi have various ways of interacting with metals; some of the techniques are increasing or decreasing the mobility of metals, sorption, or even cellular uptake. After the metals have been absorbed by the fungus, they can be chemically altered to be stored or translocated through the hyphae into various plants that participate in symbiosis. Pesticide Degradation Pesticide accumulation is an issue of great concern among the public, because they are directly associated with food products and water supplies. There are number of technologies used for pesticide cleanup; however, these technologies are generally expensive and inefficient because they require contaminated soil to be excavated and sent to a separate storage location for processing. Bioremediation offers a potential solution that treats contaminated soil and groundwater

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Table 3 The different methods of bioremediation used by the microorganisms for degrading pollutants Pollutants 2,2,6- Trinitrotoluene (TNT) Atrazine Chlorpyrifos Dibenzothiophene (DBT) Hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) PAHs

Phenanthrene, PAH

Polychlorinated biphenyl (PCB)

Polycyclic aromatic hydrocarbon (PAH)

Organisms Methanococcus sp. Pseudomonas sp. (ADP) Enterobacter strain B-14 Rhizobium meliloti Acetobacterium paludosum

Function Biotransformation Biodegradation Biodegradation Biodegradation Biodegradation

Pseudomonas sp. Pycnoporus sanguineus Coriolus versicolor Pleurotus ostreatus Fomitopsis palustris Daedalea elegans Rhizobium sp. Agrobacterium sp. Bacillus sp. Burkholderia sp. Rhodococcus erythropolis TA421 Rhizobium sp. Fungi

Biodegradation

Biodegradation

Biodegradation

Biodegradation

without needing excavation. Studies show that white rot fungi has high promise for soil bioremediation application; however, most tests have been conducted in the lab rather than in the actual environment. This fungus demonstrates the ability to transform and mineralize specific pesticides in soil. Environmental Applications Although fungi demonstrate significant biochemical and ecological useful qualities, they are hardly utilized for biotechnological purposes. Instead, bacteria are most commonly used because they usually produce superior results in their numerous advantages ranging from their highly specific biochemical reactions to their capabilities of breaking down pollutants efficiently. Fungi are underused primarily because of the costs that come from providing oxygen to fungi in polluted environments. However, filamentous fungi could be highly valuable in situations where bacteria cannot perform. For example, fungi are useful in situations where contaminants are physically blockaded and bacteria cannot reach or in circumstances of environmental extremes such as high acidity or dryness prevent bacteria from functioning.

Limitations of Treating Hazardous Wastes Using the Biological Treatment Method Some common environmental limitations to biodegradation are related to hazardous chemical wastes which possess high waste concentrations and toxicity. Because some time this toxicity either inhibits the growth of microorganism or even kill

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them. For proper growth of microorganism, it requires favorable pH condition, sufficient amount of mineral nutrients and temperature on which maximum microbes can survive, i.e., 20  C–30  C. Once the limitations by environmental conditions are corrected, the ubiquitous distribution of microorganisms, in most cases, allows for a spontaneous enrichment of the appropriate microorganisms. In the great majority of cases, an inoculation with specific microorganism is neither necessary nor useful. Besides other factors also affect bioremediation such as solubility of waste, nature and chemical composition of waste, oxidation-reduction potential of waste, and microbial interaction. Hence, researchers should search genetically different type of microbes which can also work on slightly adverse conditions. Therefore, bioremediation is still considered as a developing technology to regulate the day-to-day environmental problems faced by men residing in an area.

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Part IV Management of Radioactive Wastes

Management of Radioactive Wastes

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Bouchra El Hilal, Mohammed Hussein Rafeq Khudhair, and Ahmed El Harfi

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Notions of Radioactivity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . The Laws of Radioactivity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Radioactive Activity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Radioactive Decay . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Radioactive Waste . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Classification of Radioactive Waste . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Radioactive Waste Management: Case of Morocco . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . International Obligations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Purpose of National Radioactive Waste Management Policy and Scope . . . . . . . . . . . . . . . . . . . . . . Policy Objectives . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Long-Term Management of RW . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . General Principles of the Morocco Policy for Radioactive Waste Management . . . . . . . . . . . . . . Organizational Framework and Allocation of Responsibilities . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . National Council for Nuclear Energy . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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B. El Hilal (*) Laboratory of Agro resources Polymers and Process engineering (LAPPE), Team of Macromolecular and Organic Chemistry, Faculty of sciences, Ibn Tofail University, Kenitra, Morocco Operation Unit of the Radioactive Waste, Center of Nuclear Studies of Maamora (CENM) (CNESTEN), Kenitra, Morocco M. H. R. Khudhair Laboratory of Agro resources Polymers and Process engineering (LAPPE), Team of Macromolecular and Organic Chemistry, Faculty of sciences, Ibn Tofail University, Kenitra, Morocco Laboratory of Chemistry of Solid State, Faculty of Science, Ibn Tofail University, Kenitra, Morocco A. El Harfi Laboratory of Agro resources Polymers and Process engineering (LAPPE), Team of Macromolecular and Organic Chemistry, Faculty of sciences, Ibn Tofail University, Kenitra, Morocco © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_83

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Establishing a New Unified and Independent Regulatory Body . . . . . . . . . . . . . . . . . . . . . . . . . . . Defining Role of Operators and Organization Responsible for Centralized and Long-Term Radioactive Waste Management . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . National Strategy for Radioactive Waste Management . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Disused Sealed Radioactive Sources . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Strategic National Framework for Radioactive Waste Management . . . . . . . . . . . . . . . . . . . . . . . . . . . Monitoring Strategy Implementation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Updating the Strategy . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . National Model for Radioactive Waste Management . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Waste Generation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Predisposal Management of Radioactive Waste . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Waste Management End Points . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Long-Term Waste Management . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Management of Disused Sealed Radioactive Sources . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Management of Spent Radioactive Ion Exchange Resins . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . National Process for Implementing the Radioactive Waste Management Strategy . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

The radioactivity is a natural phenomenon and natural sources of radiation present in the environment. Radiation and radioactive substances have many useful applications, ranging from power generation to medical, industrial, and agricultural applications. At the end of their use, the radioactive elements containing in the used materials (radioactive waste) cannot be destroyed by any known chemical or mechanical process. Their final destruction is the fact either of a transformation into stable isotopes by radioactive decay or of a nuclear transmutation under the effect of the bombardment of atomic particles. The management of the radioactive waste, therefore, consists of controlling radioactive waste and maintaining it at a tolerable level in order to control the risks presented to man and his environment. The stakes of the management of radioactive waste are therefore of several kinds, technical and scientific, but also political and ethical; the fact that certain wastes generated pose safety problems in the very long term because their harmfulness can extend over very long periods of time. Keywords

Radioactve waste · Management of radioactive wastes · Radioactivity · Radioactive decay · Classification of radioactive waste · Disposal · Quality management system ion exchange resins · Radiological ris · Exempted waste · Very short lived waste · Very low level waste · Low level waste · High level waste · Low and intermediate level waste · Waste minimization · Safety and security Abbreviations

CENM CGEM CNEN

Centre d’Etudes Nucléaires de la Ma^ a mora General Confederation of Moroccan Companies Conseil National de l’Energie Nucléaire

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CNESTEN CNRP CNSN DSRS EW IAEA LLW MEMEE MoH NORM NRSSA QMS RAIS RW RWM SNF TENORM TSO VLLW VSLW WHO WMO

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Centre National de l’Energie, des Sciences et des Techniques Nucléaires Centre National de Radioprotection Commission Nationale de Sûreté Nucléaire Disused sealed radioactive sources Exempt waste International Atomic Energy Agency Low level waste Ministère de l’Energie, des Mines, de l’Eau et de l’Environnement (Ministry of Energy, Mines, Water and Environment) Ministry of Health (Ministère de la Santé) Naturally occurring radioactive materials Nuclear and Radiological Safety and Security Agency (Agence de Sûreté et de Sécurité Nucléaires et Radiologiques) Quality Management System Regulatory Authority Information System Radioactive waste Radioactive Waste Management Spent nuclear fuel Technologically enhanced naturally occurring radioactive materials Technical support organization Very low level waste Very short lived waste World Health Organization Waste Management Organization

Introduction Most human enterprises produce waste, some of which is radioactive (Ojovan et al. 2013; Venturini 2010; Giusti 2009). Radioactive waste contains materials that emit ionizing radiation, which has been recognized as a potential hazard to human health since the beginning of the twentieth century. The safe management of radioactive waste is therefore essential for the protection of human health and the environment, in the present and future (Jorda and Forest 1999). Radioactive waste is produced during the operational and decommissioning phases of facilities associated with the following activities: • Operation of the nuclear reactor and other facilities within the nuclear fuel cycle. • Production and use of radioactive materials in the fields of scientific research, medicine, industry, agriculture, trade, and education. • Extraction, isolation, and treatment of raw materials (such as oil, gas, and other strategic industries) containing naturally occurring radioactive materials (NORM). • Secondary radioactive waste resulting from the management of waste. • Disused or spent radioactive resources.

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Radioactive waste may occur in a gaseous, liquid, or a solid form that may range from low radioactivity (for example, medical waste, laboratory waste, and certain mining wastes) to highly radioactive waste (for example, spent nuclear fuel and certain spent radioactive sources). The physical and chemical characteristics of the various wastes (e.g., the activity concentration, half-life, mixture of radioactive nuclides, chemical toxicity, and radiotoxicity) vary widely. Radioactive waste may also occur together with other hazardous chemical or biological materials. The levels of radiation associated with radioactive waste should be seen in perspective to the natural background radiation to which everyone is exposed in everyday life. Radioactive wastes generated by facilities range from low volumes, such as spent radioactive sources, to large and diffuse volumes, such as legacy decommissioning waste or NORM materials from mining, and their radioactive decay products. In Morocco, the use of sealed or nonsealed radioactive sources in solid or liquid form in the healthcare sectors (radiodiagnostics, nuclear medicine, radiotherapy, etc.), industry (gamma radiography, level gauges, or tracers), agriculture (sterilization or conservation of products), mines, and other (teaching and research) is the source of the production of radioactive waste of various kinds and form (liquid or solid).

Notions of Radioactivity All the bodies of nature consist of atoms or assemblies of atoms. An atom is the smallest amount of an element that can combine to form a molecule or crystal structure (Zaitar 2012). The atom consists of: • A central nucleus made up of an assembly of protons and neutrons. • A peripheral cloud composed of a series of electrons. The nucleus (or nucleus of atom) is characterized by the total number of its nucleons A, called mass number, and by the number of its protons Z, called the atomic number. Most nuclei of atoms are naturally stable (Fig. 1) The arrangement of nucleons within the nucleus results from a balance between electrostatic repulsive forces that tend to spread protons and attractive forces with very short radii that tend to bring the nucleons closer together. The radioactivity results from an imbalance between these forces, due to an excess of either neutrons or protons or both (Horowitz 1952; Radvanyi and Bordry 1984; Boudia 2001; Curie and Curie 1899; Douce and Guillot-Salomon 1970; Joliot and Curie 1934). Two nuclides are isotopes if they have the same atomic number Z but different numbers of masses A. An isotope, when it is radioactive, is called a radionuclide or radioisotope. The chemical properties of radioactive isotopes are the same as those of stable isotopes. One element may thus have radioactive isotopes and nonradioactive isotopes (Fig. 2). The hydrogen atom has two isotopes: deuterium and tritium. The electrical charge of the three cores as well as the chemical properties are identical. Deuterium is stable, as for tritium, it is unstable and radioactive.

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Fig. 1 Illustrative diagram of an atom (Zaitar 2012) Neutron

Proton

Electron

Fig. 2 The different isotopes of the hydrogen atom

L’Hydrogène 1H

Noyau : 1 électron 1 proton

Le deutérium 2H ou D

Noyau : 1 électron 1 proton 1 Neutron

Le tritium 3H ou T

Noyau : 1 électron 1 proton 2 Neutrons

The Laws of Radioactivity Radioactive Activity Radioactive activity is the number of disintegrations that occur per unit time in a given amount of the radionuclide that constitutes it (Zaitar 2012), expressed as: AðtÞ ¼ λ NðtÞ Or λ is the radioactivity constant, characteristic of the element under consideration. The unit is Becquerel (Bq): equal to one decay per second.

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Fig. 3 Radioactive decay

N N0

N( t ) = N0 e –λt

N0 2

t 1/

2

t

Radioactive Decay The radioactive decay of a given nucleus is a random phenomenon. On the other hand, the calculation of the probability makes it possible to know with precision the number of nuclei that will transform per unit of time (Devillers et al. 1975; Zaitar 2012; Braham et al. 1979; Vieu et al. 1973). dN ¼ λ N dðtÞ If one knows the activity of the radionuclide and the number of atoms present N0, one can then deduce the number of remaining atoms Nt as a function of the time elapsed: Nt ¼ N0 eλt This is the radioactive decay, which is therefore an exponential function of time. However, for each radionuclide a radioactive period T is defined at the end of which half of the radioactive atoms disappear by spontaneous transformation (Fig. 3). Nt=N0 ¼ eλt ¼ ½etlog2 ¼ λT T ¼ log2=λ

Radioactive Waste The International Atomic Energy Agency (IAEA) defines radioactive waste as “any material for which no use is foreseen and which contains radionuclides in concentrations above the values that the competent authorities consider to be permissible materials in materials suitable for noncontrolled use” (International Atomic Energy Agency 2002, 2008, 2011).

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Classification of Radioactive Waste Safe management of radioactive waste depends on knowledge of its characteristics and characterization may occur in several stages. Knowledge of the origin of the waste provides the first indication of its characteristics; direct measurements and observations then improve or confirm this information. The aim of characterization is to define the properties sufficiently to allow the waste to be accepted for successive waste management steps and ultimately to meet waste acceptance criteria for the disposal facility. Waste generators have responsibility for characterizing the wastes they generate. The regulatory body must be satisfied that this has been performed adequately. Demonstration of an adequate characterization regime is aided by the application of a quality management system (QMS) that produces, for each waste package, records that are reliable, traceable, and retrievable. The QMS will also include training, method development, qualification, and documentation. Remembering that radioactive waste is an industrial product, there is a possibility that it may contain hazardous substances that are nonradioactive. Disposal of such substances will need to satisfy environmental legislation, which means that they too should be considered during waste characterization. The table below identifies the radioactive waste classification scheme (Table 1).

Radioactive Waste Management: Case of Morocco As a Member State of the IAEA and a Party to the Joint Convention on the Safety of Spent Fuel Management and the Safety of Radioactive Waste Management, the Kingdom of Morocco Country is committed to managing radioactive waste in a safe, secure and sustainable manner in accordance with internationally recognized principles related to nuclear and radiation safety. The application of those principles aims to ensure adequate protection of workers, the general public and the environment, now and in the future, from the harmful effects of ionizing radiation. Requirements for nuclear and radiation safety apply to all types of radioactive waste present in the country, regardless of their physical and chemical characteristics or origin, at all stages of their management. Compliance with International Standards implies in particular the implementation of ICRP recommendations on radiation protection and the application of Fundamental Safety Principles as specified in IAEA Safety Publications. In addition to matters of safety, the Kingdom of Morocco is also committed to establishing a beneficial, responsible, and sustainable RWM system in accordance with the IAEA Nuclear Energy Basic Principles and Fundamental Safety Principles. When applied to the management of radioactive waste and spent nuclear fuel, these require: • Minimization of generation and optimization of the management of radioactive waste

Solid

Waste type Liquid

Various sources Research and medical institutions

EW

VSLW

Organic low and intermediate level waste (OLILW) High level waste (HLW) Contamination from damaged spent nuclear fuel

Reactors, research institutions, medical institutions, universities, and industry Decommissioning waste, medical institutions, research institutions, operation of nuclear facilities, uranium mining and milling Decommissioning waste, isotope production, fuel fabrication, enrichment, reprocessing of spent nuclear fuel Research processes, reactors, spent nuclear fuel

Very short lived waste (VSLW) Very low level waste (VLLW)

Low level waste (LLW)

Waste origin Various sources

Waste class Exempted waste (EW)

Table 1 The radioactive waste classification scheme

Annual dose for population>10 μSv T1/2 < 100 day

Contact dose rate > 2 mSv/ h, no long-lived radionuclides

Contains organics

3.7GBq/m3>Activity >37MBq/m3 T1/2 M3+ mainly due to decrease in ionic radius and increase in Lewis acidity of the metal ion (Altmaier et al. 2013). The initial step in the hydrolysis of a cation is usually the formation of the mononuclear species MOH(z-1)+ and in general it is represented as Mzþ þ H2 O , MOHðz1Þþ þ Hþ

(1)

Complete reaction is represented by including water of solvation: ðz1Þ

þ MðOH2 Þzþ n , MðOHÞðOH2 Þn1 þ H

(2)

In the above reaction a proton is lost by a solvating water molecule and on further hydrolysis, ðz1Þ

þ MðOHÞðOH2 Þn1 ! MðOHÞ2 ðH2 OÞz2 n2 þ H ðz2Þ

þ MðOHÞðOH2 Þn2 ! MðOHÞ2 ðH2 OÞz3 n3 þ H

(3) (4)

and so on. But in reality complete hydrolysis does not happen; instead other chemical reactions or aggregation of two or more hydrolyzed species takes place and in general it is represented as follows.

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N. Priyadarshini et al. (y-1)+

(y-1)+ H2 O

OH (H2O)d-2M

M(H 2O)d-2

+ HO

OH2

ð5Þ (y-1)2+ H O M(H 2O)d-2

(H2O)d-2M

+

2H2O

O H

When this reaction further continues, then there is a possibility of formation of high-molecular-weight polynuclear chains. The determination of the identity and stability of dissolved hydrolysis products has proven to be a difficult and challenging task primarily for two reasons: I. The hydroxide complexes formed are often polynuclear, that is, they contain more than one metal ion. It will be readily perceived that this can result in the formation of a greater variety of species during the hydrolysis of cations. More hydrolysis products present simultaneously in appreciable amounts. II. The range of pH over which the formation of soluble hydrolysis products can be studied is often limited by the precipitation of the hydroxide or the oxide of the metal cation (Baes and Mesmer 1981).

Hydrolysis of Actinides Hydrolysis reactions are the primary complexation reactions of actinide elements in aqueous solution. It is mainly correlated with the electrostatic interaction energy between the actinide ion and the OH ligand (Choppin 1983; Grenthe and Puigdomenech 1997; Neck et al. 2001). The general order is Pu4+ > Np4+ > U4+ > Pa4+ > Th4+. The effective charge of the actinide ions decreases in the order of An4+ > AnO22+ > An3+ > AnO2+. Due to high electric charge, tetravalent actinides can hydrolyze even under very acidic conditions (pH < 3 depending on actinide) (Walther et al. 2007). In the near-neutral pH, An(OH)(aq) species dominates in the absence of other complexing ligands. Actinide ions in the trivalent and tetravalent states in acidic solutions are in the form of the simple hydrated ions An3+ and An4+, whereas An5+ and An6+ exist as trans-dioxo cations. This is due to large positive charges which make them readily strip oxygen atoms from water

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Actinide Speciation in Environment and Their Separation Using. . .

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molecules. They are extremely stable and have a symmetric and nearly linear structure [O = An = O]n+ where n = 1,2. This structure decreases the effective charge on the central actinide ion making it less prone for hydrolysis compared to tetravalent actinides (Choppin 1999).

Plutonium Among the actinides, plutonium is the most chemically diverse and fascinating element in the periodic table. Plutonium is generated by transmutation of uranium in nuclear reactors in large scale. Plutonium has 18 isotopes. The most important among these are 238Pu (t1/2 = 86 years) and 239Pu (t1/2 = 24,110 years) (Clark et al. 2006). The ground-state electronic configuration of atomic plutonium is [Ru]5f 67s2. Plutonium ions in solution exist in +3, +4, +5, and +6 oxidation states as Pu3+, Pu4+, PuO2+, and PuO22+ (Cleveland 1970). The chemistry of plutonium is in great contrast to the light elements of periodic table and it is very difficult to control. Some of the complexities of plutonium chemistry in aqueous medium are listed as follows.

Complexities in Plutonium Aqueous Chemistry Three reaction pathways have to be considered while investigating plutonium in aqueous systems: hydrolysis, polymerization and colloid formation, and redox reactions of different oxidation states of plutonium. It is difficult to study one particular reaction without the interference of the other. Because all the reaction takes place simultaneously, that is, mainly due to the same redox potentials of the couples, Pu (III)/Pu (IV) (E0 = 1.031 V) and Pu (IV)/Pu (V) (E0 = 0.936 V), an equilibrium of two or more oxidation state may occur in solution. Differences in redox potentials for different oxidation states of plutonium are given in Fig. 1. In other words, plutonium acts both as oxidizing and reducing agent at the same time. This is called disproportionation reaction. The disproportionation reaction mainly happens at elevated temperature. The equations governing the redox reactions for plutonium ions under acidic conditions are given as Eqs. 6, 7 and 8: þ 2Pu4þ þ 2H2 O , Pu3þ þ PuOþ 2 þ 4H

(6)

3þ Pu4þ þ PuOþ þ PuO2þ 2 , Pu 2

(7)

þ 4þ 2PuOþ þ PuO2þ 2 þ 4H , Pu 2 þ 2H2 O

(8)

778

Pu(VI) PuO22+

N. Priyadarshini et al. 0.94 V

Pu(V) PuO2+ 0.99 V

1.01 V

1.04 V Pu(IV) Pu4+

Pu(III) Pu3+

-2.00 V Pu(0)

-1.25 V

Fig. 1 The redox potential differences for the plutonium aquo ions in 1 M perchloric acid, as well as the potential difference between the plutonium aquo ions and pure Pu (Cleveland 1970)

• Conversely, under some conditions, two plutonium ions of different oxidation states can react by means of a reproportionation reaction. The two ions are simultaneously oxidized and reduced to form two ions of the same oxidation state. • Another important complexity is that all plutonium isotopes are radioactive. One milligram of plutonium emits about 106 alpha particles per second, and the radioactive decay is constantly adding energy to the plutonium solution. This leads to radiolytic decomposition of water-producing redox reagents such as short-lived radicals H, OH, and O. These radicals recombine to form H2, O2, and H2O2. Thus radiolysis tends to reduce Pu(VI) and Pu(V) to Pu(IV) and Pu(III) states (Clark 2000).

Hydrolysis and Polymerization of Tetravalent Plutonium The tendency towards hydrolysis, polymerization, and further to colloid formation is strongest for tetravalent plutonium as compared to other oxidation states due to its high effective charge. The effective charges for different oxidation state of plutonium are Pu4+(~ + 4) > PuO22+ (~ + 3.3) > Pu3+ (~ + 3) > PuO2+ (~ + 2.2). Pu(IV) is stable only at very high acidities and considerable amounts of Pu(III), Pu(V), and Pu (VI) are formed at pH > 0 as shown in Fig. 2 (Walther 2008). Only at pH < 0.5, Pu4+(aq) is the dominant species whereas with increasing pH the hydrolysis reaction leads to the quantitative formation of mononuclear complexes Pu(OH)n 4n where n = 1–4. Rapid polymer formation takes place even in 1  103 and 1  104 M Pu(IV) solutions and close to solubility (Kumar and Koganti 1997). The polymeric plutonium refers to a hydrolytic form which is characterized by its bright emerald green color. The only means by which polymer formation can be avoided is to detect colloids that are formed initially in the solution. This is because the freshly formed polymer can be depolymerized where as aged ones are difficult due to conversion of amorphous to crystalline form which involves the modification of hydroxyl bridges to oxo bridges as shown in Eq. 9 (Seaborg and Loveland 1990).

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Actinide Speciation in Environment and Their Separation Using. . .

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Fig. 2 Plutonium oxidationstate distribution depending on –log[H+] [21] (Walther 2008)

m+ H O xPu(OH)n 4-n

- yH 2 O

H O Pu

Pu

Pu O H

O H

Hydroxy bridged

ð9Þ

- yH 2 O m+ O

O Pu O

- yH 2 O Pu

Pu

PuO 2

O

oxo bridged

Parameters Influencing the Formation of Plutonium Polymers/Colloids The formation and properties of polymers and colloidal hydroxides are directly dependent on pH of the solution. At alkaline pH, some hydroxides can form anionic

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complexes and possibly polymers. With increase in pH the degree of polymerization and formation of colloidal aggregates also increases (Schuelein 1975). Ionic strength of the electrolyte affects the solubility. Increase in ionic strength increases the coagulation. Concentration of metal ion has direct effect on polymerization and colloid formation. With decrease in Pu(IV) concentration, the pH of colloid formation increases (Bitea et al. 2003b). The rate of polymerization is a function of temperature. Polymerization can occur even at room temperature in solution less acidic than 0.3 M H+ and in 1.26 M H+ at boiling temperatures. The amount of polymer formed at 75  C is ~3.5 times higher than the amount at 25  C when the initial acidity is 0.075 M (Costanzo et al. 1973). The free energy of plutonium polymer is estimated as about 341.9 kcalmol1 whose solubility product of Pu(OH)4 corresponding to this number is 2.5  1056 (Silver 1983). This shows that the polymerization phenomenon is a very rapid and irreversible process (Toth et al. 1981). Under conditions of low acidity and elevated temperature, the polymeric species require drastic treatment for conversion to the ionic state. The presence of low concentrations of Pu in natural waters and in disposal streams is of much interest to those concerned with reactor and radiation safety. These solutions are usually nearly neutral and any plutonium is probably in the polymeric state. Formation of plutonium colloids can be detected by spectroscopic studies as each oxidation state of plutonium and polynuclear plutonium has its own characteristic color (Fig. 3a). The absorption spectrum of polymeric Pu(IV) state is different from those of other plutonium species (Fig. 3b). At pH 0–2, there is a steady decrease in the characteristic absorption band at 470 nm characteristic of ionic Pu(IV). They also

a

b

000 100

absorptivity (M–1cm–1)

60

colloids

50 40 Pu(IV)

Pu(VI)

Pu(III)

30 20 Pu(V)

10 0 400

500

600 700 wavelength (nm)

800

Fig. 3 (a) Characteristic colors of each oxidation state of Pu and Pu colloids (Clark 2000). (b) Absorption spectra of different oxidation state of Pu and Pu colloids (Rabideau and Kline 1960)

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Actinide Speciation in Environment and Their Separation Using. . .

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show relatively broad, low-intensity bands and rather intense with increasing absorption below 460 nm (Rabideau and Kline 1960). Polymerization of Pu(IV) leads to alterations in redox equilibria and additional easily identifiable spectral feature around 620 nm appears for the presence of small Pu(IV) colloids. Absorption spectra of aged polymers do not show any difference with that of fresh polymers except the pronounced 620 nm peak (Cleveland 1967).

Probing the Formation and Structure of Pu(IV) Polymer Though first step in the formation of polynuclear species is the condensation of mononuclear species, there are no reports on the solution structure of Pu (IV) mononuclear hydroxides. As an evidence for the presence of dimer, an X-ray diffraction study has reported a solid-state structure of Pu2(OH)2(SO4)3.4H2O with dimeric [Pu2(OH)2]6+ units linked by bridging sulfate ligands (Wester 1982). Electrospray ionization mass spectrometry (ESI-MS) studies on dimers, trimers, and tetramers revealed the presence of mixed oxidation states of plutonium, i.e., Pu(III) and Pu (V) (Walther et al. 2009). This gives an insight into the coexistence of monomeric and pentavalent plutonium species with colloids and also their interference in redox reactions (Newton et al. 1985; Rai and Swanson 1981; Neck et al. 2007). A hexanuclear [Pu6(OH)4O4]12+ core with glycine ligands containing both oxo and hydroxo bridges was also reported (Knope and Soderholm 2013a). The same linkage was also found in case of lanthanides (Wang et al. 2000). The formation of the polynuclear species is assumed to involve hydroxide or oxygen bridging of plutonium ions but some reviews consider such species to be thermodynamically unstable (Lemire et al. 2001; Guillaumont et al. 2003). As an alternative they favor colloid formation from mononuclear hydroxide complexes. Thus Pu(IV) colloids are more stable in contrast to other tetravalent species such as Th(IV) and U(IV). The most widely accepted mechanism of formation of plutonium colloid involves the condensation of [Pu (OH)n](4n)+ through an olation reaction to yield hydroxo-bridged species. On further condensation, the hydroxo-bridged oligomers produce mixed plutonium oxide hydroxides (Knope and Soderholm 2013b). X-ray absorption fine structure (XAFS) and laser-induced breakdown detection (LIBD) measurements proposed that the formation of Pu(IV) colloids follows by stacking eightfold coordinated Pu units. It involves the formation of a trinuclear species through hydrolysis and condensation of a monomeric Pu(OH)2(H2O)62+ unit with a binuclear species to form single-edge sharing, leading to the formation of neutrally charged species as product or a doubleedge sharing species with common corner. The dimers and trimers agglomerate and eventually form nm-sized colloids (Fig. 4). The predominant species contain highly asymmetric oxygen coordination which indicates the presence of different Pu-O bond lengths from different coordinated oxygen atoms (–O, –OH, –OH2). PuxOy(OH)4x2y(H2O)z (0  y  2x) has been proposed as colloid structure formed by stacking of mononuclear or polynuclear building blocks on top of each other. During this process, the cubic subunits are preserved and form a distorted fluorite like plutonium lattice (Rothe et al. 2004). Single crystals of 38 plutonium

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Fig. 4 Schematics showing mechanism of hydrolysis and polymerization of Pu4+(aq) leading to formation of nm-size colloids Rothe et al. (2004)

nanoclusters [Li14(H2O)20[Pu38O56Cl54(H2O)8] and Li2[Pu38O56Cl42(H2O)] 15H2O] were isolated for structural characterization using single-crystal X-ray diffraction. Both structures consist of the [Pu38O56]40+ core, wherein 38 8-coordinate Pu(IV) cations are bridged exclusively via oxo linkages. The data clearly shows that intracluster packing and structural topology of clusters of composition [Pu38O56Cl54(H2O)8]40+ are identical with PuO2 and the surface is occupied by chloride ion and water molecules (Soderholm et al. 2008; Wilson et al. 2011). However the solid-state structure of the crystal does not contain Pu-OH moieties which contradict the structure proposed by Rothe et al. (2004). Small-angle neutron

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scattering (SANS) experiments were carried out on colloidal Pu(IV) suspensions in aqueous phase and polymer extracted into C6D6 solutions by alkyl esters of phosphoric acid. The results revealed the presence of long, thin rodlike polymer in both the phases. However the diameter of ellipsoids of extractant-Pu(IV) polymer complexes in C6D6 is larger than that of aqueous polymer. This increase in diameter is due to the attachment of extractant molecules around the rod-shaped Pu polymer. The length of aqueous Pu(IV) polymer is larger than the polymer extracted in C6D6 (Thiyagarajan et al. 1990). A comparative study was also made between freshly prepared and 5-year-old Pu(IV) colloids using Pu L3-edge extended X-ray absorption fine structure (EXAFS). The aged colloidal particles contain only 3–4 Pu atoms with a structure very similar to solid Pu(IV) oxide but with shorter Pu-O and Pu-Pu distances which was caused due to partial oxidation of Pu(IV) to Pu(V)/Pu (VI) (Ekberg et al. 2013).

Size and Morphology of Pu(IV) Colloid Polymeric plutonium was observed in two forms: amorphous and crystalline primary particles and secondary particles which appeared to be aggregates of the primary particles. X-ray diffraction patterns of precipitated plutonium show a faint PuO2 structure, suggesting that the hydroxide is in reality hydrated PuO2 rather than Pu(IV) hydroxide. The precipitate becomes more crystalline when aged at elevated temperatures (Haire et al. 1971). Similarly the precipitates aged in a basic or neutral medium display a crystalline pattern that corresponds to the cubic PuO2 structure and was confirmed by electron diffraction pattern. Colloids formed in highly acidic media differ from those formed at higher pH (Bell et al. 1973). Thus colloid morphology varies strongly depending on the method of preparation, conditions of formation, and age of the solutions. For instance, fast neutralization of Pu(IV) solutions with a base or water usually forms less ordered structures (Reed et al. 2006). Scanning electron microscopy (SEM) and transmission electron microscopic (TEM) images show amorphous morphology for these polymers (Neu et al. 1997). The IR spectrum of Pu(IV) polymer showed major bands at 360 and 3400 cm1. They are assigned to Pu-O vibration similar to crystalline PuO2 and to OH stretching vibration of a water group which is either hydrated or occluded in the polymer structure. The comparison of the IR spectrum of PuO2 and the Pu (IV) polymer precipitated and dried in air showed close resemblance of the polymer bands in the 250–600 cm1 region suggesting a similar lattice structure in the two compounds. Unlike PuO2, water and possibly hydroxyl groups are present in the polymer (Toth and Friedman 1978). The size of Pu(IV) colloids ranges from a few nanometers to almost micrometers, depending on the conditions of generation (Lloyd and Haire 1978; Triay et al. 1991). Electron micrograph shows that the primary particles are extremely small of the order of 5–20 Å whereas the amorphous primary particles are of the order of 10 Å (Lloyd and Haire 1973). TEM analysis of dried solutions shows evidence of small PuO2-like clusters with a diameter of about 2 nm. LIBD experiments gave mean particle size

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Fig. 5 TEM image of (a) intrinsic Pu nanocolloid cluster. (b) TEM image of intrinsic Pu nanocolloid and tubular crystal of goethite (Powell et al. 2011)

of 5–12 nm. Particle size may increase to the point of precipitation. In acidic solutions (pH 0.4–1) the weighted mean colloid size increases from 12 to 25 nm with increasing degree of oversaturation with respect to amorphous Pu (IV) hydroxide. They are said to be composed of small crystalline particle covered by an amorphous layer (Walther and Denecke 2013). The small-angle X-ray diffraction (SAXS) and XRD affirmed the presence of U(IV) > Np(IV) > Pu(IV), due to decrease in An-O distance in the lattice (Neck and Kim 2001; Guillaumont et al. 2003). The same trend is observed for crystalline An(IV) dioxides. The solubility products Ksp0 and Kspo (at infinite dilution) of amorphous Pu(IV) oxide or hydroxide, Pu(OH)4(am), are given by Eq. 13: PuðOHÞ4 ðamÞ , Pu4þ þ 4OH-

(13)

  0 o 0 ¼ Pu4þ ½OH 4 and K sp ¼ K sp ðγPu Þ ðγOH Þ4 K sp log Kspo for crystalline PuO2 is 64.0 and for amorphous hydrated hydroxide 58.3 [30]. Thus crystalline PuO2 is far less soluble than its amorphous hydroxide. The slope of the solubility curve obtained from LIBD studies (Fig. 6) provides further evidence that the dihydroxo-complex (Pu(OH)22+) is the prominent species in the range of 1.0 < pHc < 1.6. The slope = 2 for the solubility curve infers that colloid formation proceeds from Pu(OH)22+ via consumption of 2OH per Pu(IV) ion suggesting Eq. 14 (Walther et al. 2007): þ

PuðOHÞ2 2 þ 2OH $ PuðOHÞ4ðam=collÞ

(14)

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Fig. 6 Solubility of amorphous Pu(OH)4(am) comparing available literature data Walther et al. (2007)

Methods Adopted to Remove Polymers Two ways can be adopted to avoid the polymers formed in the system. They can be either extracted or depolymerized.

Extraction of Pu(IV) Polymer/Colloid Attempts were made to extract plutonium polymers or colloids using neutral extractant TOPO (trioctylphosphine oxide), CMPO (octylphenyl-N,Ndiisobutylcarbamoylmethylphosphine oxide), and the bifunctional extractants DHDECMP (dihexyl-N,N-diethylcarbamoylmethyl phosphonate). But in most of the cases there is crud formation. They also co-extract other oxidation-state metals (Muscatello et al. 1983). Back extraction of the polymer either was done in a silica sol or required large concentrations of salts to reverse the extraction (Chaiko 1992). A novel approach was made by dissolving the [Pu38O56Cl54(H2O)8]14 nanocluster in aqueous solution of LiCl producing green solution of colloidal Pu which was

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confirmed by the optical spectra. Addition of 6 M HCl changes green-colored solution to red with change in the spectrum. The rapid change involves ion- or ligand-exchange reaction between water and Cl. But the core moiety Pu38O56 remains unperturbed. Trichloroacetic acid (TCA) in n-octanol was chosen as extractant due to its immiscibility with water. The well-defined surface of Pu colloids can be altered and manipulated which can involve in ion- and ligandexchange reactions resulting in selective extraction of only plutonium colloid leaving other plutonium species in aqueous phase.

Depolymerization Depolymerization of plutonium polymer is a slow process. Freshly formed polymer can be depolymerized whereas aged or polymers heated to higher temperatures are difficult to depolymerize and it needs drastic conditions. In other words it leads to very low depolymerization rates due to conversion of amorphous to crystalline primary particles which involves more extensive cross-linking (Hunter and Ross 1991; Brunstad 1959). Some of the reported methods applied for depolymerizing the polymer if formed accidentally are listed below.

Proposed Methods for Depolymerizing Pu(IV) Polymer • Pu(IV) polymers can be depolymerized in the presence of reductants and oxidants such as hydroquinone, H2O2, hydroxylamine, U(IV) and KMnO4, H2O2 + Fe (NO3)3, Na2S2O8 + catalysts, and Co(III) in 0.533 M HNO3. This is because Pu (III) and Pu(VI) formed due to reduction and oxidation are not prone to polymerization (Ermolaev et al. 2001; Oak et al. 1983). • Passing direct current through an aqueous nitric acid solution which contains Pu(IV) polymer converts the polymer to Pu3+ and PuO22+ ions in turn are converted to Pu4+ ions in the solution, followed by the step in which Pu4+ ions are extracted from the solution. This can be done by contacting the acid solution with a solution containing about 30% tri-n-butyl phosphate and 70% hydrocarbon diluent (such as dodecane). It can also be done with ion-exchange resin (Tallent et al. 1982). • To avoid the complications of colloid formation, plutonium has often been injected as the citrate complex, prepared by adding Pu(NO3)4 in concentrated nitric acid to a 0.3–2.0% solution of trisodium citrate followed by adjustment with base to pH 4–7. The rate of depolymerization is directly proportional to citrate concentration, time, and acidity (Lindenbaum and Westfall 1965). • Depolymerization is accelerated by introducing fluoride, sulfate, and other ions forming strong complexes with Pu4+ (Weigel et al. 1997). • The process of depolymerization is accelerated on exposure to UV radiation (Bell and Friedman 1976; Friedman et al. 1977).

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Uranyl Nitrate Effect • Solutions of plutonium are commonly used in the presence of large amounts of U. Therefore the effect of uranyl ion on the polymer chemistry is of considerable interest. Much uncertainty exists about the effect of UO2(NO3)2 on the formation rate of Pu(IV) polymer because this solute provides nitrate ions which could either complex with Pu(IV) and stabilize it with respect to hydrolysis and polymerization or, through complexation with Pu(IV), shift the disproportionation equilibrium to the left, thereby decreasing the real acid concentration of the solution and causing a corresponding increase in the rate of hydrolysis and polymerization. UO22+ functions as a chain terminator through the formation of hydroxyl bridges to terminal uranyl groups (Quiles and Burneau 1998). Also, Pu(IV) polymer formed in the presence of uranium is shorter than those which are formed in the absence of uranium. This shows that presence of uranium in a polymerizing Pu(IV) solution reduces the rate of polymer formation about 30% without appreciably entering the polymer. • Costanzo and Biggers (1963) prepared polymer in 5 N HNO3 and aged for several months. It is observed that for aged polymers the depolymerization halftime is 320 h and it is only 20 h for freshly prepared polymer. • Depolymerization occurs easily at 90  C in 6 M HNO3. Also, Pu(IV) polymer can be dissolved using Na2S2O8. Decomposition of S2O8 yields H2SO4 favoring Pu(IV) depolymerization (Savage and Kyffin 1986).

Hydrolysis of Th(IV) Th(IV) exhibits a strong tendency towards hydrolysis and subsequent polymerization over a wide range of pH and concentrations. Thorium has an ionic radius of 1.09 Å in nine coordinations making it the largest stable tetravalent metal ion and in spite of its high valence it is least susceptible to hydrolyze when compared to U (IV) or Pu(IV) (Torapava et al. 2009). Many studies have reported Th (IV) hydrolysis and the species formed in solution depending on pH and concentration. Th4+(aq) ion can be found as a predominant species only at pH < 3. It has been observed that mononuclear ionic complexes dominate the species distribution at very low concentrations as well as at high acidity (Khazaei et al. 2011). At nearneutral pH, and as the thorium concentration approaches the solubility limit of the amorphous hydroxide Th(OH)4, polymers become the dominant species in solution, starting with dimers which grow and form pentamers. Pentamers have been found to be the most abundant and stable complex rather than tetramers and hexamers (Walther et al. 2012). The polymeric species are found to contain one or more thorium atoms linked by hydroxy bridges (Bacon and Brown 1969). With increasing pH and thorium concentration it forms thorium nanoparticles or the so-called colloids.

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Probing the Formation and Structure of Th(IV) Hydrolyzed Species Th(IV) hydrolysis and polymerization reactions follow the same pattern as Pu(IV). Th(IV) colloid formation is preceded by polynuclear hydroxide complexes such as Th2(OH)26+, Th2(OH)35+, Th4(OH)88+, Th4(OH)128+, Th6(OH)1410+, and Th6(OH)159+. But in contrast to Pu(IV) polynuclear hydroxide complexes, Th (IV) polynuclear hydroxide complexes are thermodynamically unstable (Walther et al. 2008). Structural studies on aqueous solution of the hydrolysis products of Th (IV) have identified three different types of hydrolysis species: a μ2O-hydroxo dimer, [Th2(OH)2(H2O)12]6+; a μ2O-hydroxo tetramer, [Th4(OH)8(H2O)16]8+; and a μ3O-oxo hexamer, [Th6O8(H2O)n]8+ (Rand et al. 2007). Single crystals of oxo- and hydroxo-bridged haxameric [Th6(μ3-O)4(μ3-OH)4]12+ (Takao et al. 2009; Knope et al. 2011) and octameric [Th8O4(OH)8]16+ (Knope et al. 2012) have also been isolated by addition of complexing ligands. SAXS investigation revealed the presence of microcrystalline particles with the thorium oxide structure in highly hydrolyzed Th(IV) solutions (Magini et al. 1976). In another study, when a thorium chloride solution was left to evaporate at room temperature, it produced hydroxobridged Th dimers. Thorium structure of Th(IV) pentamers was investigated by XAFS and by quantum chemical calculations. It was considered that the most favorable structure contains two Th(IV) dimers linked by a central Th(IV) cation through hydroxide bridges. These polymers are then subjected to continuous hydrolysis; that is, the number of hydroxide ligands increases with increase in pH. The ultimate end product of Th hydrolysis is an oxo-bridged ThO2 with fluorite structure (Sellers et al. 1954).

Size and Morphology of Hydrolyzed Th(IV) Polymer/Colloid Freshly formed small particles have a mean diameter in the range of 16–23 nm (Rothe et al. 2002; Bundschuh et al. 2000). Three different oxides/oxo-hydroxide solid phases have been reported for thorium, that is, crystalline ThO2 whose particle size is >50 nm, microcrystalline ThO2.xH2O of 15–30 nm, and amorphous ThO2. xH2O of 2–5 nm. Upon heating a highly hydrolyzed thorium solution the particle size was estimated to reach as large as 20–50 Å in diameter. But if that same solution was left to age without evaporation it produced small ThO2 particles with sizes of about 20 Å. The small 20 Å ThO2 particles are X-ray amorphous (Neck et al. 2002). DLS measurement shows that well-defined colloids were formed at two different pH domains. The colloids formed at pH~0.8 were smaller, when compared to those formed at pH~2 (Priyadarshini et al. 2016).

Hydrolysis and Solubility of Th(IV) M-O bond length of Th(IV) is largest compared to Pu(IV) due to actinide contraction and hence anionic ligand forms much stronger for Pu(IV) than for Th(IV).

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Accordingly the first hydrolysis constant of Th(IV) is smaller than Pu(IV). Some of the reported values of formation constants for the first mononuclear hydroxide complex M(OH)3+ are logβ1,1 = 11.8  0.2 and logβ1,1 = 11.5  0.5. In order to measure the solubilities, two regions with respect to pH and Th(IV) concentration need to be considered, which correspond to different solid phases, microcrystalline and X-ray amorphous ThO2 (Ekberg et al. 2000). Rai et al. (2000) determined the solubility product of crystalline ThO2 in acidic solution (pH < 2.5) as log Kspo = 54.2  1.3. A combined coulometric and pH titration in the pH range 1.5–2.5 with LIBD gave a solubility product of log Kspo = 52.8  0.3. The ambiguity in the solubility data is mainly due to the presence of polynuclear species or intrinsic colloids prevailing in the system. With increase in pH and close to the solubility limit, where Th(OH)4 is the predominant aqueous species, the solubility product of crystalline ThO2 is similar to that of amorphous ThO2 (log Ksp (Th(OH)4 (am;hyd)) = 47.8). Hence it can be concluded that in near-neutral solutions the bulk crystalline ThO2 is covered with an amorphous and more soluble surface layer. Solubility studies conducted on Th(IV) intrinsic colloids in concentrated NaCl (0.5 and 5.0 M) and MgCl2 (0.25, 2.5, and 4.5 M) solutions within the pH range 8.8–10.8 revealed that the hydrophilic oxo-hydroxide intrinsic colloids formed through chemical polynucleation reactions are stable equilibrium species and contribute to the total solubility of Th(IV).

Stability of Th(IV) Hydrolyzed Colloids Laser-induced breakdown (LIBD) spectroscopy combined with ultrafiltration was used to investigate the generation of Th(IV) colloids in the concentration range from 105 to 102 M at pH 3–5 in 0.5 M NaCl. Surprisingly the colloids generated by coulometric titration are found to be small in size and remain stable up to 400 days of investigation, without a tendency towards agglomeration or precipitation. There was no change in size and concentration of colloids during the course of investigation which shows their extreme stability. These suspensions on dilution lead to an equilibrium between colloids and ionic species (Bitea et al. 2003a).

Uranyl Effect on Polymerization of Th(IV) Previously it was shown that the presence of uranyl ion retards the rate of Pu (IV) hydrous polymer formation. Raman spectroscopic studies were done on Th (IV) in order to characterize the uranyl-thorium(IV) interaction. By monitoring the changes in the frequency of the uranyl symmetric stretching vibration, that is, the shift in 869 cm1, Raman band corresponding to unassociated UO22+ to 851 cm1 for UO22+ attached to Th(IV) polymer. In the same solutions when aged, there was change in bridging structure (i.e., hydroxyl to oxygen) which was confirmed by the appearance of band at 665 cm1 and by the change in color of dispersed polymer from a uranyl pale yellow to a urinate golden orange (Toth et al. 1984).

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Molecular Weight of Th(IV) Hydrolyzed Polymer Static light scattering measurements showed that the Th(IV) colloids formed at pH ~0.8 have a molecular weight of ~3449 Da and it shows that ~15 Th atoms are present in the polymeric network at the initial stage. The second virial coefficient is negative. It implies that there was aggregation in the system leading to precipitation. The colloids formed at acidic pH tend to agglomerate causing precipitation in the system. But, the colloids formed at pH > 2.0 have a molecular weight of ~6223 Da with positive second virial coefficient. This polymeric colloid formed due to polynucleation contains around 20 atoms of Th when initially formed. The solute-solvent interaction in the system is higher leading to greater stability of the system. The colloids upon ageing for 7 days did not show any considerable change in their molecular weight but when it was aged for 55 days the molecular weight increased to 43,573 Da (Priyadarshini et al. 2016).

Uranium In nuclear reactors, uranium is handled in large quantities when compared to other actinide elements. The behavior of uranium in radioactive waste repositories and the migration of uranium in the environment depend on its oxidation state and its speciation is determined by pH, presence of complexing anions, and total salt background. Uranium exists in different oxidation states (from +3 to +6). The known redox states of uranium are U(II), U(III), U(IV), U(V), and U(VI) and their corresponding reduction potentials are given in Fig. 7. Uranium(III) can exist only under anaerobic and strongly reducing conditions. U(IV) is also considered as toxic waste under strongly reducing conditions that is often accepted to be present in deep geological repositories. U(V) is unstable and disproportionates. U(VI) is the most common species in the environment. Under natural aquatic systems, uranium exists as U(IV) and U(VI).

Hydrolysis and Polymerization of Uranium(IV) The known redox states of uranium are U(II), U(III), U(IV), U(V), and U(VI). U (IV) aqueous chemistry is of particular interest under reducing conditions which prevails in both inside of nuclear waste repositories and under deep uranium mines. Due to hydrolysis, U(IV) occurs in the form of [U(OH)x](4x)+. The hydrolysis -1.60 V 0.19 V

UO2

2+

0.28 V

-4.7 V

-0.61 V

0.37 V

UO2+

4+

U

-0.1 V

U2+

U3+ -1.35 V

Fig. 7 The redox potential differences for the uranium ions and pure U

U

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reactions of U(IV) resulting in the formation of polynuclear hydrolysis complexes are given in Eq. 15: h ið4xyÞþ  4þ x UðH2 OÞn þ yH2 O , Ux ðOHÞy ðH2 OÞny þ yH3 Oþ (15) U(IV) is readily oxidized to U(VI) when reducing conditions are not maintained carefully. Due to hydrolysis U(IV) occurs in the form of [U(OH)x](4x)+. There are limited literature data on hydrolysis studies on U(IV).

Probing the Formation and Structure of U(IV) Hydroxide/Oxide In contrast to UO22+ for which several oxides and hydroxides are known, only few U(IV) oxide/hydroxide structures are known. Uranium(IV) dioxide adopts a fluorite structure. The compound prepared by hydrothermal hydrolysis of a U(IV) sulfate solution at 100–150  C showed a structure isostructural with that described for thorium and consists of infinite chains of hydroxo-bridged (U(OH)2)n2n+ units with UU distances of 3.90 Å.

Size and Morphology of U(IV) Colloid Attempts were made to produce colloids of U(IV) by electrochemical reduction of U (VI). EXAFS investigations were performed and compared with solid UO2 and amorphous UO2.nH2O. They showed correspondence with the latter (Opel et al. 2007). Around 20 nm colloidal particle was detected in 2.5 mM U(IV) solutions at pH ~3.6 by DLS experiment. It decreased to ~5 nm when the concentration of U (IV) was increased to 19 mM (Priyadarshini et al. 2014).

Hydrolysis and Solubility of U(IV) The solubility product of crystalline UO2(cr) formed at pH ~1 and amorphous hydrous oxide UO2.xH2O(am) formed at pH ~3 was determined by combined coulometric titration and LIBID experiments as log Kspo = 59.6  1 and log Kspo = 54.1  1 (Manfredi et al. 2006). Compounds containing U(IV) are insoluble in mildly acidic and alkaline medium. U(OH)2SO4 was prepared at 100–150  C from the hydrothermal hydrolysis of a U(IV) sulfate solution. The structure is isostructural with that described for thorium and consists of infinite chains of hydroxo-bridged (U(OH)2)n2n+ units (Qiu and Burns 2013).

Molecular Weight of Hydrolyzed U(IV) Polymer Light scattering measurements performed on freshly formed polymers of U(IV) showed a weight average molecular weight (Mw) of 1820 Da and it increased to

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13,000 Da when aged for 3 days. Around 40–50 atoms of U are considered to be present in the aged polymer. Positive value of second virial coefficient shows that the solute-solvent interaction is high leading to stable suspension of aged polymers (Priyadarshini et al. 2014).

Hydrolysis of Uranium(VI) Under aerobic conditions, hexavalent uranium is more relevant in aqueous medium. U(VI) is thermodynamically most stable form with relatively high solubility (Zanker et al. 2007; Zhao and Steward 1997). Being a hard Lewis acid, it shows strong interactions with hard donor ligands resulting in a tendency to undergo hydrolysis, formation of soluble complexes with carbonates and phosphates, and the most important is its ability to form colloids which will increase the migration behavior in natural systems (Eliet and Bidoglio 1998). Hence it becomes mandatory to study the speciation and related reaction products of uranium in aqueous systems, under well-defined conditions such as pH, ionic strength, concentration, and presence of other ligands. The hydrolysis equilibrium reaction of uranyl ion in general can be written as given in Eq. 16:  ð2mnÞþ mUO2þ þ nHþ 2 þ nH2 O , ðUO2 Þm ðOHÞn

(16)

The solution chemistry of U(VI) is dominated by the linear dioxo cation UO22+ (Kirishima et al. 2004). In aerated aqueous solutions at pH  2.5, the uranyl ion is very stable. The hydrolysis of uranyl ion starts at pH values greater than 3 (Clark et al. 1999) and also if the concentration of uranyl exceeds 103 M (Steppert et al. 2012). UO2OH+ is the only mononuclear hydrolysis product present at appreciable concentration up to pH ~7 in 105 M or more concentrated U(VI) solutions (Brooker et al. 1979). The wellestablished dinuclear species (UO2)2(OH)22+ is present over the pH range of 3–6 and accounts for at most 40% of the total U(VI) in 102 M solutions. (UO2)3(OH)5+, (UO2)4(OH)7+, and (UO2)3(OH)7 are the dominant polynuclear species from pH 5 to 9 in solutions of 102 to 105 M U(VI), accounting for up to 100% of the total uranium in solution. Potentiometric measurement (Palmer and Nguyen-Trung 1995), timeresolved fluorescence spectroscopy (Eliet et al. 2000), Raman spectroscopy (Quiles and Burneau 2000), attenuated total reflection infrared (Muller et al. 2008), and electrospray mass spectrometry (Clark et al. 1999) studies reveal that in acidic solutions of pH levels (2  pH  5) UO22+, (UO2)2(OH)22+, and (UO2)3(OH)5+ are the predominant species. Complexes (UO2)3(OH)7, (UO2)3(OH)82, (UO2)3(OH)104, (UO2)3(OH)115, and (UO2)3(OH)42 have also been suggested by various methods (Kirishima et al. 2004). There is consensus that several polynuclear species coexist in alkaline solution. A number of structures that contain chains or extended structures of hydroxo- and oxo-bridged UO22+ have been reported. The hydrolyzed U(VI) species aggregate and reach colloidal dimensions and exist as suspended sols in the aqueous medium. Particle size is an important property to

794

UO2(CO3)34–

100 Relative distribution (%)

Fig. 8 Speciation diagram of U(VI) ions as a function of pH in aqueous solution at atmospheric condition (Scierz and Zanker 2009)

N. Priyadarshini et al.

UO2(CO3)(OH)3–

UO22+

80 β-UO2(OH)2(s)

60 40

(UO2)3(OH)5

+

UO2(CO3)22–

UO2(OH)

20 0

3

4

+

5

6

7

8

9

10

pH

define a colloidal system. It influences the behavior of colloids during diffusion, aggregation, etc. DLS studies were performed on different concentrations of uranyl solution diluted to neutral pH. A peak around 32–36 nm was observed on all the concentration range of U(VI) in DLS (Priyadarshini et al. 2014). Hence it has been confirmed that the colloids have size of ~32–36 nm. With increasing concentration of uranyl ions the pH for colloid formation decreases considerably. The colloids formed were not stable for a long period of time. There is no change in average particle size of the colloids as the concentration of U(VI) increases. Well-defined colloids are formed when the concentration reaches the precipitation point. The molecular weight of U(VI) colloid was determined as ~763 Da which was much less when compared to U(IV) and Pu(IV) (Priyadarshini et al. 2014). The calculated molecular weight shows that there may be 2–3 units of uranyl groups bridged by hydroxyl groups. Hence it is only oligomer that has been formed. Figure 8 represents the relative abundance of U(VI) species as a function of pH in the aqueous solution at atmospheric normal environmental.

Nanomaterials for Separation of Actinides Safe management of radioactive waste with minimum impact to environment is emerging as one of the major challenges in separation science. Accumulation of large volume of the radioactive wastes is highly risky. The risk is directly proportional to volume of the waste. Therefore, these waste streams need to be treated to reduce their activity to a level at which they are permitted as per national regulations. This leads to more demanding efforts in removal of radionuclides from aqueous solution for environmental concern. It is one of the most urgent technological problems that mankind face worldwide. Decay is the only natural way of reducing radioactivity. Since radionuclides have decay rates ranging from seconds to thousands of years, proper segregation of wastes depending on their half-lives and

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conditioning is an important factor. The major aim is to separate long-lived alphaemitting actinide radionuclides. Hence, safe and efficient management of long-lived actinide is of utmost importance. Under this situation, versatile materials and technologies that can remove the radionuclides and remediate the environment are of severe demand. Although a wide range of separation techniques are employed for removal of radionuclide, amalgamation of nanotechnology into conventional techniques has revolutionized the separation processes. Nanomaterials offer high surface area, high sorption capacities, selectivity, and reusability (Dash and Chakravarty 2017). In recent years, nanomaterials such as carbon nanostructures, meso porous materials, and nanopolymers/dendrimers have been extensively investigated for actinide separation. The potential application of aforementioned nanomaterials in removal of long-lived actinides from aqueous media is briefly discussed in this chapter.

Carbon Nanostructures Carbon nanostructures, including carbon nanotubes (CNTs), nanodiamonds, fullerenes, graphene, and other carbon materials, have been the most effective materials for environmental remediation, and have attracted great attention recently. The large specific surface area, the outstanding thermal and chemical stabilities, and the recent development in large-scale synthesis make carbon nanostructures attractive as interesting possible solid-phase extraction (SPE) materials (sorbents) for treatment of radioactive waste. Various types of bulk sorbents such as organic (Sureshkumar et al. 2010), inorganic (Li et al. 2012), biosorbent (Ghasemi et al. 2011), composites (Gao et al. 2010), and other carbon-based material (Starvin and Rao 2004) have been developed for the recovery of radionuclides from aqueous systems. However, the low sorption capacities or efficiencies of these materials have obviously restricted their applications. Search for alternatives to conventional bulk adsorbents has intrigued intense interest on use of nanomaterial-based adsorbents due to their unique physicochemical properties as compared to their bulk equivalents. Nanomaterials circumvent many limitations of bulk material-based adsorbents. Carbon nanostructure characteristics such as large surface area, high specificity, fast adsorption kinetics, and ability to interact with different chemical species make them excellent candidates for radionuclide separation.

Carbon Nanotubes (CNTs) Carbon nanotube (CNT) is a well-known carbon nanostructure. CNTs can be described as a graphite sheet rolled up into a nanoscale tube; they are typically several nanometers (nm) in diameter and both ends are normally capped by fullerene-like structures. There are two main types of CNTs: (i) single-walled CNTs and (ii) multi-walled CNTs named based on the number of graphene sheet. CNTs exhibit unique properties such as strong tensile strength, large elastic module,

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high heat conductivity and electrical conductibility, and large surface area (Belloni et al. 2009). Due to its extraordinary physicochemical properties, CNTs have been widely applied in various scientific areas such as electronic, power engineering, medical, and catalysis. With a continuously decreasing cost of production and required properties (large surface area, hollow structure, strong affinity for ionic and organic species, chemical and radiation stability), its application has been stretched to sorbent material for extraction of radionuclides. For example, pristine MWCNTs were investigated as a sorbent for europium (Tan et al. 2008), americium (III) (Wang et al. 2005), thorium (Chen et al. 2007a), uranium, and plutonium radionuclides due to the strong complexation of sorbates on the MWCNT surface in the aqueous and NaClO4 medium. Functional groups like –COOH, –C=O, and –OH are incorporated on the surface of MWCNTs by treating with strong oxidizing agents like conc.HNO3 or mixture of conc. HNO3 and conc.H2SO4 at elevated temperature (Chen et al. 2007b; Darmstadt et al. 2003). The oxidized MWCNTs offer a more hydrophilicity and surface area. The oxidized MWCNTs show fast kinetics and a high sorption activity towards uranium, thorium, and plutonium radionuclides in weakly acidic to weakly basic solutions (pH 3–8) (Myasoedova et al. 2009). The pH of aqueous solution plays an important role; it influences both the metal ion speciation and total surface charge on sorbents. The isoelectric point of pristine and oxidized MWCNTs is ~7.1 and ~5, respectively (Wang et al. 2005). The isoelectric point of MWCNTs (ξ-potential) depends on their preparation, pretreatment methods, and different functionalities on MWCNTs. At pH less than isoelectric point (ξ-potential), the surface of MWCNTs is positively charged due to protonation reaction and negatively charged due to deprotonation at pH greater than isoelectric point. The electrostatic repulsion between the positively charged surface of the sorbent and positively charged species of uranium/thorium/plutonium resulted in a very low sorption capacity at lower pH. In the pH range of 5–8, the transition in the surface polarity of MWCNTs favored the sorption of hydrolyzed species of uranium (UO2(OH)+, (UO2)2(OH)22+, (UO2)3(OH)5+, UO2(OH)2)/thorium (Th(OH)3+, Th (OH)22+, and Th(OH)4) and maximized the uptake capacity. But at higher pH values (pH > 8), the electrostatic repulsion between the negatively charged MWCNTs surface and the negatively charged uranium species (UO2(CO3)22 and UO2(CO3)34) resulted in a rapid decrease in sorption. In case of thorium at pH > 4, deposition of precipitated thorium complexes onto the MWCNTs resulted in decrease of sorption. Plutonium is known to take part in disproportionation, hydrolysis, and polymerization reactions and form concurrently polymeric forms in different oxidation states (Walther 2008). The oxidized CNTs are capable of extracting both the polymeric Pu4+ and ionic forms (Pu3+, Pu4+, PuO2+, PuO22+) from weakly acidic to weakly basic solutions (Perevalov and Molochnikova 2009). Pu sorption on oxidized MWCNTs follows the Pu4+ > PuO22+ > PuO2+ trend due to their charges. Both the pristine- and acid-pretreated CNTs were found to be useful for actinide sorption; however their removal efficiency and selectivity towards particular radionuclides remain quite limited. In recent years, much attention has been diverted towards making the MWCNTs highly selective and less hydrophobic. Due to higher hydrophobicity, the dispersion of MWCNTs in an aqueous medium is less, resulting in aggregation of MWCNTs due to van der Waals interaction and thus hindering in the sorption process to a greater extent. To enhance the

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dispersibility in aqueous solution, sorption performance, and selectivity towards particular radionuclides, CNTs have been functionalized without much alteration of their physical and chemical properties. Modification of CNTs is accomplished through covalent or noncovalent functionalization. MWCNTs were easily dispersed in solution by plasma-induced grafting of carboxymethyl cellulose on MWCNT (MWCNT-gCMC). MWCNT-g-CMC has much higher adsorption capacity for U(VI) compared to the raw MWCNTor oxidized MWCNT (Wang et al. 2009). Similarly, MWCNTs grafted with chitosan (MWCNT-g-CS) showed high sorption performance for uranium (Shao et al. 2010). MWCNTs functionalized with organic ligands are shown to be more selective than pristine and oxidized MWCNTs. Specifically, the CNTs are functionalized with amide- and phosphorus-based organic ligands used in the liquid extraction. The sorption behavior of functionalized MWCNTs depends on the functionality and the mechanism adopted for the sorption. Sorption performances of diglycolic acid (DGA)-functionalized MWCNTs (DGA-MWCNTs) (Deb et al. 2012, 2013) have been evaluated for U(VI) and Th(IV) in aqueous solution. The sorption behavior of U (VI) and Th(IV) on DGA-MWCNTs is highly influenced by pH of the solution. Effective sorption of U(VI) and Th(IV) on DGA-MWCNTs was found to be at pH 5–7 and 4, respectively. DGA-MWCNTs have been investigated for U(VI), Am (III), and Pu(IV) from nitric acidic medium (Sengupta et al. 2017a; Deb et al. 2012). The sorption of U(VI), Am(III), and Pu(IV) increases gradually with nitric acid concentration and reaches maximum at 3–5 M HNO3. Further increase in acid concentration does not alter much the sorption. At acidic condition (pH < 2), DGA group in MWCNTs remains unionized and acts as neutral chelating group which can bind effectively with neutral metal nitrate species. The increase in formation of M(NO3)n (Pu(NO3)4, Am(NO3)3, UO2(NO3)2) species with nitric acid concentration effectively coordinates with DGA group; thereby increase in sorption was observed:   Mnþ þ nNO3  $ MðNO3 Þn M ¼ Am3þ =Pu4þ =UO2 2þ and n ¼ charge MðNO3 Þn þ DGA-MWCNTs $ MðNO3 Þn :DGA-MWCNTs

(17) (18)

Sorption of U(VI), Pu(VI), Am(III), and Pu(IV) on dihexylamide (DHA)functionalized MWCNTs (DHA-MWCNTs) also found to be increasing with nitric acid concentration (Gupta et al. 2017b). However, DHA-MWCNTs have shown low sorption performance for UO22+ and Th4+ compared to DGA-MWCNTs. Sorption of Pu4+ or Am3+ on poly(amidoamine) (PAMAM) dendrimer of generation 1 and 2 functionalized MWCNTs (MWCNT-PAMAMG1 and MWCNT-PAMAMG2) was studied (Kumar et al. 2017). PAMAM-functionalized MWCNTs (MWCNTPAMAMG1 and MWCNT-PAMAMG2) have shown much higher sorption capacity for Pu4+ and Am3+ compared to DGA-MWCNT or DHA-MWCNT sorbents. PAMAM dendrimer has more number of amide and amine functional groups and they increase with dendrimer generation MWCNT-PAMAMG(n = 1,2. . .). These amide and amine functional group of PAMAM forms strong complex with f-block element leads to have higher sorption capacity than other functionalized MWCNTs.

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The amidoamine-functionalized MWCNTs (MWCNT-AA) were also employed for the sorption of Pu4+,NpO2+, PuO22+, and NpO22+ in an acidic aqueous medium (Sengupta et al. 2017b). In aqueous medium neptunium exists as NpO2+ without disproportion with lower chemical potential making it very difficult to separate. However, the sorption performance of MWCNT-AA for Np was found ten times higher than ligand-impregnated MWCNTs making it an efficient solid-phase sorbent for the separation of Np from nuclear waste solution. The sorption of neptunium species (NpO2+ and NpO22+) increases with increase in aqueous nitric acid concentration and remains almost constant above 3M HNO3 concentration. However, an exponential increase in sorption of Pu4+ and PuO22+ on MWCNT-AA was observed with aqueous nitric acid concentration (Sengupta et al. 2017b).

Ligand-Impregnated MWCNTs Ligands which showed high distribution coefficient along with selectivity for a particular oxidation state of actinides in liquid–liquid extraction have been impregnated on MWCNTs. Various phosphorus-based ligands like diphenyl (dibutylcarbamoylmethyl) phosphine oxide (CMPO), tri-n-octylphosphine oxide (TOPO), tri-n-butylphosphate (TBP) [Murali and Mathur 2001; Yaftian et al. 2003], nitrogen-based ligands like N,N‘-dimethyl-N,N‘-dioctylhexylethoxymalonamide (DMDOHEMA) (Serrano-Purroy et al. 2005), and task-specific ionic liquids like tetra alkyl ammonium hydrogen phthalate (Gupta et al. 2017a) and phosphonium ionic liquid Cyphos IL-101 (Quinn et al. 2013) were shown distribution ratio and selectivity towards actinides in the presence of various other fission products. Hence, these ligands were impregnated on the MWCNTs by stirring a suspension of MWCNTs and a ligand directly in nitric acid solution. Impregnation is mostly a combination of pore filling and surface adsorption. The extractant gradually fills the pore space starting with the smallest pores and moving up to pores of about 10 nm and then surface adsorption becomes the dominant force. Interaction between the extractant and support is usually quite weak, consisting of only the attractive forces between alkyl chains and/or aromatic rings of the ligand and those of the support. The sorption of Np(V), U(VI), Pu(IV), Am(III), and Eu(III) onto CMPO-, TOPO-, and Cyphos IL-101-impregnated MWCNTs (Taunit) had shown some promising results in nitric acid solution (Zakharchenko et al. 2012). Taunit-TBP and TaunitDMDOHEMA sorbents are also useful for the separation of plutonium from acidic nuclear waste solution (Zakharchenko et al. 2012).

Graphene Oxide (GO) Graphene oxide (GO) is one of the most important graphene derivatives. It has unique structure and exceptional physicochemical properties. The remarkable properties of graphene include high Young’s modulus, fracture strength, thermal conductivity, mobility of charge carriers, specific surface area (theoretical calculated

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value of 2630 m2 g1), high chemical stability, large pore volume structure, and fascinating transport phenomena (Chen et al. 2012). GO has many oxygencontaining functional groups in the form of epoxy, hydroxyl, and carboxyl groups on its basal plane and edges. These oxygen-containing groups can bind metal ions and organic pollutants through coordination, electrostatic interaction, hydrogen bonding, etc., which ensures its potential application in environmental remediation (Nupearachchia et al. 2017). GO is hydrophilic in nature due to existence of many oxygen-containing functional groups. In addition, the delocalized π electron systems of graphene layer as Lewis base can form electron donor–acceptor complexes with radionuclides as Lewis acid through Lewis acid–base interaction. The U(VI), Am(III), Sr(II), and Pu(IV) radionuclides were removed using GO sorbent (Romanchuk et al. 2013). GO was shown as an efficient sorbent for Th(IV) due to large surface area and sufficient exposure of active sites of abundant oxygen-containing functional groups with Th (IV) (Bai et al. 2014). Removal of U(VI) ions from an aqueous system using few-layered GO nanosheets was reported. However, the aggregation and accumulation of few-layered GO nanosheets resulted in limited sorption of U(VI) which was observed in the aqueous systems (Zhao et al. 2012). GO-based materials are easier to process and easily functionalized. Therefore, polymers including polyaniline, polyacrylamide, cyclodextrins, chitosan, and amidoxime were grafted onto GOs to introduce various functional groups and to enhance their dispersibility in solutions and thereby improve the removal ability of some radionuclides (Wang et al. 2016). Such approaches, which added functionality to groups already present on the GO surfaces, make GO a more versatile sorbent (Dreyer et al. 2010). Cyclodextrinmodified GO nanosheets (CD/GO) were synthesized using an in situ polymerization method (Song et al. 2014). The mutual effects on the simultaneous removal of U (VI) and humic acid from an aqueous system by CD/GO were investigated. The sorption of U(VI) and humic acid (HA) sorption on CD/GO highly depend on pH and ionic strength. The presence of HA enhanced U(VI) sorption at low pH and reduced U(VI) sorption at high pH, while the presence of U(VI) enhanced HA sorption.

Magnetic Nanoparticles for Sorption Magnetite nanoparticles (MNPs) have received much attention recently to remove radioactive metal ions due to the easy separation of sorbent from solution with the aid of magnetic field. The bisphosphonate-modified magnetite Fe3O4 nanoparticles were used to remove U(VI) from blood at pH 7.0 (Wang et al. 2006). However, the Fe3O4 nanoparticles are unstable at acidic condition, which limits their application for nuclide sorption from acid nuclear wastewater. Aggregation and decomposition of Fe3O4 nanoparticles under extreme pH conditions hinder the separation process of actinides. This situation can be circumvented by designing core-shell type of nanomaterials (Kaur et al. 2013a). The magnetic core material is protected by a shell made up of nonmagnetic materials such as polymer, silica, carbon, and gold. This

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sort of core-shell dimension reduces aggregation as well as improves extraction efficiency. For instance, silica-coated magnetic nanoparticles were developed for the separation of actinides from acidic nuclear waste, in which the Fe2O3 nanoparticles were coated with the silica, followed by covalent attachment of the actinide-specific chelators. The silica-coated MNPs are stable even in 1 M HCl solution, and show enhanced actinide separation efficiency compared to the uncoated counterparts (Han et al. 2010). The affinity of the magnetic nanoparticle towards a particular metal ion is enhanced by functionalizing it with -thiol, -amine, and -carboxyl functional groups. These functional groups provide active sites for the capture of metal ions which get complexed by chelate formation, ion-exchange process, or electrostatic interactions. Therefore, surface-functionalized Fe3O4 magnetic nanoparticles are preferred over pristine Fe3O4 nanoparticles for the separation of metal ions (Singh et al. 2013). The selectivity of the chelators towards particular metal ion can be tailored by selecting ligands with oxygen, nitrogen, and sulfur donor atoms. This is done based on hardsoft nature of donor atoms and metal ions. Generally hard and soft metal ions form strong complexes with hard and soft ligands, respectively (Kaur et al. 2013a). This concept is exploited for the selective extraction of metal ion using ligands. Diethylene triamine penta acetic acid (DTPA) chelator covalently tethered to silica-coated magnetic nanoparticles was found to be efficient for the recovery of Am(III), Np(V), Pu(IV), and U(VI) under acidic conditions (Kaur et al. 2013b). N,N, N0 ,N0 -tetraoctyl diglycolamide (TODGA)- and bis(2-ethylhexy)phosphoric acid (HDEHP)-coated superparamagnetic Fe3O4 nanoparticles (NP) were investigated for the extraction of Am(III) and Pu(IV) from the solution of 3–4 M HNO3 (Ojha et al. 2017). TODGA-coated magnetic nanoparticles extracted both Am(III) and Pu (IV) whereas HDEHP-coated magnetic nanoparticles selectively extracted Pu(IV). This sort of magnetic nanoparticle-extractant framework can be effectively utilized for the pre-concentration of metal ions followed by their easy separation.

Nanocomposites Nanocomposite is a multiphase solid-state material with at least one of the phases in nano-dimension which might be zero-dimensional (quantum dots) or one-dimensional (nano rods) or two-dimensional (nanosheets) (Okpala 2013). Also, the multiphase materials with their phases separated by a distance of nanoscale are termed as nanocomposites. The idea of incorporating nanoscale materials is to promote the synergism among the constituents that emerges into novel properties capable of meeting desired specifications. Multifunctional capability of nanocomposites arises due to the benefit of nanoscale properties, huge interphase zone, and chemical functionalization. Different types of metal-, carbon-, ceramic-, and polymer-based nanocomposites possess a variety of properties (Camargo et al. 2009). The properties of nanocomposites can be modified by choosing the right parameters. These parameters include nanoscale dimension, loading, degree of dispersion, shape, orientation, matrix material, and interaction among the

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constituents. These parameters are meticulously selected to prepare materials with desired flexibility, ductility, strength, conductivity, compatibility, magnetism, and surface activity (Camargo et al. 2009). Nanocomposites are used in catalysts, sensors, opto-electronic devices, magnetic devices, food packaging, fuel cells, dentistry, construction, coating, medicine, and automobiles (Okpala 2014). Nanocomposites also found potential applications in the separation of heavy metal ions. The prospective of nanocomposites over the separation of actinides has emerged recent times. CNT nanocomposite was prepared by using hydroxylated fullerene (C60(OH)n), carboxylated fullerene (C60(C(COOH)2)n), and oxidized-MWCNTs (Wang et al. 2013). The presence of hydroxylated fullerene (C60(OH)n) and carboxylated fullerene (C60(C(COOH)2)n) in MWCNTs increased the sorption of Th4+ onto oxidized MWNCTs at pH < 4. However, the sorption of Th4+ lowered at pH > 4 with the increasing concentration of C60(OH)n or C60(C(COOH)2)n (Wang et al. 2013). This striking behavior was due to the good sorption affinity of C60(OH)n or C60(C (COOH)2)n onto oxidized MWCNTs by strong π–π electron donor–acceptor interactions between the flat surfaces of both aromatic C60(OH)n/C60(C(COOH)2)n and oxidized MWCNTs. Thus, the increasing concentration of fullerene created a competitive pathway for the sorption of Th4+. Polyvinyl alcohol MWCNT (PVA-MWCNT) composite was evaluated for the removal of U(VI) (Abdeen and Akl 2015). The maximum sorption of U(VI) on PVA-MWCNTs was observed at pH 3 and the adsorbed uranium could be desorbed by using 0.1 M EDTA. Various GO-based composites were synthesized and used for environmental remediation. GO-supported sepiolite (GO@sepiolite) composites were synthesized applied for the removal of U(VI) (Cheng et al. 2013a). GO-activated carbon felt (GO–ACF) composites were prepared by an electrophoretic deposition and subsequent thermal annealing(Chen et al. 2013b). GO–ACF showed high sorption capacity for U(VI) due to presence of carboxyl functional groups in GO–ACF. The sorption of U(VI) on GO-supported polyaniline (PANI@GO) nanocomposites was investigated. The PANI has a strong affinity for radionuclides and heavy metal ions due to the large number of amine and imine functional groups; these functional groups can form very strong complexes with radionuclides on the nanocomposite surfaces (Sun et al. 2013). The chemical binding of radionuclides with the nitrogen-containing functional groups is much stronger than that of radionuclides with the oxygen-containing functional groups. The sorption of U(VI) ions on GO-polypyrrole (GO/PPy) composites was much higher than that of U(VI) on either GO or PPy in an aqueous system (Hu et al. 2014). The highly selective sorption capacity towards U(VI) was observed due to the strong coordination between the U(VI) ions and nitrogen donor atoms on the GO/PPy composites. GO/PPy composites could be regenerated and reused. Hierarchical threedimensional composite (layered double hydroxide/graphene) was obtained via in situ growth of layered double-hydroxide (LDH) nanosheet arrays onto graphene sheets and sorption of U(VI) was performed from aqueous media (Tan et al. 2015c). Poly(amidoxime)-reduced GO (PAO/RGO) composites synthesized by in situ polymerization showed excellent sorption capability for U(VI) (Chen et al. 2014).

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The sorption capacity of PAO/RGO composites for U(VI) showed highest sorption capacity (872 mg g  1) compared to other GO-based composites.

Magnetic Nanocomposites Even though nanomaterials possess extraordinary features, it is practically complicated to remove the suspended nanoparticles from the solution. This problem can be tackled by introducing magnetic nanocomposites that can be easily removed by the application of magnetic field. Design of chelator-tethered magnetic nanomaterials revolutionized the process of separation of metal ions. They work on the simple principle of magnetism but they are versatile with high efficiency in the separation. The main advantage of using magnetic materials is the easy separation of the sorbent from the solution after the extraction of metal ions (Teja and Koh 2009). The MWCNT nanocomposites were prepared by incorporation of Fe2O3 or CoFe2O4 magnetic particle. These magnetic Fe2O3-MWCNT (Liu et al. 2014) and CoFe2O4-MWCNT (Tan et al. 2015a) composite were evaluated for the sorption of Th(IV) and U(VI) radionuclides, respectively. The removal of U(VI) using nanoscale zero-valent iron (nZVI) and reduced graphene oxide-supported nanoscale zerovalent iron (nZVI/RGO) was studied (Sun et al. 2014). The U(VI) species such as UO2 2+, UO2OH+, and (UO2)3(OH)5+) were efficiently removed in the weak aqueous acid medium. GOs have high dispersion properties in aqueous solution due to their hydrophilic nature; thereby it is difficult to separate GOs from aqueous solution by using the traditional separation techniques after the GOs are applied as adsorbents in the sorption process. The magnetic Fe3O4/GO composite was prepared from natural flake graphite and employed as adsorbent to remove U(VI) ions (Zong et al. 2013). After U(VI) sorption, magnetic Fe3O4/GO composite was easily removed from aqueous solution due to magnetic separation. Amidoximated magnetite/ GO (AOMGO) composites were also synthesized and applied to U(VI) sorption, and had a more sorption capacity compared to Fe3O4/GO composite (Zhao et al. 2013). MnO2–Fe3O4–RGO was successful for the removal of U(VI) (Tan et al. 2015d). In addition, U(VI)-loaded MnO2–Fe3O4–RGO was efficiently regenerated and reused. The CoFe2O4-reduced graphene oxide nanocomposite was used for the extraction of uranium (Tan et al. 2015b). The superparamagnetic GO-Fe3O4 nanocomposite was synthesized and evaluated for the uptake of Am3+, UO22+, Th4+, and Pu4+ in mildly acidic solutions (Gadly et al. 2017). The uptake of tetravalent metal ion reduced with increase in hydrolysis due to the reduction in the effective nuclear charge. These composites are recyclable and therefore can be reused. Magnetic ferberite-GO nanocomposite was utilized for the extraction of U(VI) from aqueous solution (El-Maghrabi et al. 2017). This composite possesses high recycling capacity and the morphology remained unaltered after several stages of reuse. Fe3O4-encapsulated ZIF-8 nanocomposite was used for the separation of U(VI) from Ln(III) under various experimental conditions (Min et al. 2017). The fast kinetics and high U(VI) uptake capacity were observed for this kind of metal organic framework composite. Attapulgite-Fe2O3 nanocomposites prepared by chemical

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route were used for the separation of U(VI) from its aqueous solution (Chen et al. 2013). At higher pH, sorption of U(VI) occurs through inner sphere surface complexation and at low pH sorption of U(VI) follows outer-sphere surface complexation.

Mesoporous Materials for Sorption Besides nanomaterials, mesoporous materials such as activated carbon (Carboni et al. 2013), silica (Vivero-Escoto et al. 2013), and polymeric membrane with pore size in the range of nano dimensions (size between 2 and 100 nm) also gained prominence in the separation of actinides. Ordered mesoporous carbon and silica materials (MCM-Mobil Composition of Matter-41,48 and SBA-15-Santa Barbara Amorphous-15) are novel families of the fascinating porous solids, which have the advantages of large surface area, well-defined pore size, excellent mechanical resistance, non-swelling, exceptional chemical stability and radiation tolerance, as well as extraordinarily wide possibilities of functionalization (Darmstadt et al. 2003; Lee et al. 2011). These advantages make the ordered mesoporous carbon and silica compounds attractive for nuclear waste disposal. The U(VI) was entrapped using MCM-41 and MCM-48 molecular sieves based on direct template ion exchange (Vidya et al. 2001, 2004). It was found that the entrapment of U(VI) was facilitated by the large pore size and the high surfactant content in the as-synthesized host materials. However, these sorbents normally show poor selectivity and slow sorption kinetics for U(VI). To promote the sorption selectivity, and achieve higher sorption capacity and faster sorption kinetics, functionalized mesoporous materials were synthesized. 4-Acetophenone oxime-functionalized ordered mesoporous carbon CMK-5 (Oxime-CMK-5) was developed for U(VI) sorption (Li et al. 2011). The U(VI) sorption by Oxime-CMK-5 was found to be rapid and pH dependent (Fryxell et al. 2005; Lin et al. 2005). It was found that the composite can be reused without considerable loss in sorption capacity. The 5-nitro-2-furaldehyde-modified mesoporous silica (MCM-41) was used for extraction of U(VI) (Yousefi et al. 2009). The sorbent exhibits good stability, reusability, high sorption capacity, and fast rate of equilibrium for sorption/desorption of U(VI). Recently, the phosphonatefunctionalized MCM-41(NP10) (Yuan et al. 2011) and amino-functionalized SBA-15 (APSS) (Liu et al. 2012) were synthesized by co-condensation and grafting method, respectively. These synthesized materials were used as sorbents for the removal of U (VI) from aqueous solution. These new sorbents offered large sorption capacity and possess ultrafast sorption kinetics. Furthermore, these silica materials show a desirable selectivity for U(VI) ions over a wide range of competing metal ions. Ordered mesoporous carbon (OMC) was synthesized and investigated for the sorption and desorption of Pu(VI) and Eu(III) (Parsons-Moss et al. 2014). Extraction of uranium in aqueous media by nanocomposite of two-dimensional layered double hydroxide with silica nanoparticles revealed that sorption of U(VI) is through innersphere surface complexation and electrostatic interaction (Yang et al. 2017a). Layered double-hydroxide carbon dot composite was used for the extraction of

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U(VI) and Am(III) from aqueous solutions (Yao et al. 2018). The dispersion of carbon dots into layered double hydroxides prevents agglomeration of carbon dots and thereby increased surface area and extraction efficiency of actinides. Self-assembled monolayer on mesoporous support (SAMMS) was prepared by assembling monolayers of molecules in the mesoporous support and functionalizing the free end of the monolayers (Fryxell et al. 1999). The large surface area of these functionalized mesoporous materials provides more active sites for the adsorption of metal ions. Moreover, the high porosity enhances the diffusion of metal ions and sorption kinetics. This kind of ceramic support is stable under extreme acidic, corrosive, and oxidizing environment. Glycinyl-urea-, carboxylate-, and phosphonate-functionalized SAMMS were investigated for the extraction of U (VI), Pu(IV), and Am(III) from aqueous solutions containing several foreign metal ions (Fryxell et al. 2005). Isomers of hydroxypyridinone-functionalized SAMMS were investigated to understand the extraction behavior of Pu(IV), Np(V), and U (VI) (Lin et al. 2005). The multiple ligand–metal interactions and intramolecular hydrogen bonding enhanced the extraction of actinides. The surface ion-imprinted mesoporous silica was investigated for the separation of U(VI) from highly acidic solutions (Yang et al. 2017b). The sorbent was synthesized using U(VI) (uranyl) as the template and diethylphosphatoethyltriethoxysilane as functional group. The surface ion-imprinted mesoporous silica have nano-sized three-dimensional cavity that is specific for U(VI). Using this sorbent the fast kinetics and higher selectivity for U(VI) were observed when compared to nonion-imprinted similar mesoporous materials. The material is also resistant to radiation such that the performance did not deteriorate after five consecutive usages of the ion-imprinted polymer.

Mesoporous Membrane for Sorption Removal of radionuclides from aqueous waste by membrane separation using nanofiltration (NF)/reverse osmosis (RO)/ultrafiltration (UF) membrane has become one of the emerging technologies with a rapid growth. It has drawn the attention of researchers in the field of separation technology with its better performance compared to conventional separation processes. In these filtration methods, the feed mixture is pressurized to pass through nano/micro-sized pores of the membrane and then separated into a retentate (part of the feed that does not pass through the membrane) and a permeate (part of the feed that passes through the membrane). During this process, the actinides are concentrated in the retentate by sieving and adsorption.

Reverse Osmosis (RO) Reverse osmosis (RO) can effectively remove nearly all inorganic contaminants from water. RO process was applied for the treatment of effluents containing uranium. In this process, uranium content in permeate is reduced to less than

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1 mg L1, where the removal of uranium is 99.5% (Hsiue et al. 1989; Prabhakar et al. 1992). RO process was conducted to remove uranium, technetium, tritium, strontium, and cesium from liquid effluents and groundwater (Garrett 1990; Gamal Khedr 2013). It was demonstrated that all of the above contaminants could be removed to less than the regulated limits, except tritium. Recently, RO was used in the framework of the Fukushima-Daiichi accident to treat contaminated seawater (Fournel et al. 2012).

Nanofiltration (NF) Nanofiltration membranes have a nominal pore size of approximately 0.001 microns and a molecular weight cutoff (MWCO) of 200–1000 daltons (Da) (Yong Du et al. 2016). NF membranes are very commendable for retention of multivalent cations and organic solutes. NF membranes are generally manufactured from cellulose acetate or polyamide materials. The polyamide-based NF membrane was used for treatment of radioactive effluents containing specific activity levels in 10 kBq/L. It was observed that the level of radioactive contaminants could be reduced from 10 kBq /L to few Bq/L. Studies on selective removal of uranium from aqueous solution using nanofiltration membranes showed relatively high selectivity, despite the high concentration of other divalent and monovalent cations (Reguillon et al. 2008; Raff and Wilken 1999).

Ultrafiltration (UF) Combined with Sorption/Precipitation/ Complexation Ultrafiltration has a pore size of approximately 0.002–0.1 microns, an MWCO of approximately 10,000–100,000 daltons. UF membranes are constructed from a wide variety of materials, including cellulose acetate, polyvinylidene fluoride, polyacrylonitrile, polypropylene, polysulfone, polyethersulfone, or other polymers. Each of these materials has different properties with respect to the surface charge, degree of hydrophobicity, strength, flexibility, pH, and oxidant tolerance. UF membranes have large pore size than NF membrane and they are not effective in retaining radioactive metal ions. Hence, in order to retain the radioactive ions, size enhancement of radioactive species by means of sorption/precipitation/complexation is done prior to UF. UF experiment was conducted with 239,240Pu and pseudo-colloids of Fe in water having dissolved organic carbon (DOC) in the range of 10–60 mg/L. The presence of DOC increases the colloidal size of Fe and acts as a matrix for high sorption of plutonium (Singhal et al. 2009). Studies on ferric hydroxide precipitation employed for the treatment of low-level waste (LLW) containing 241Am showed >99.9% removal of 241Am (Gao et al. 2004). The recovery of 241Am was also accomplished using sodium dodecyl sulfate (SDS) micellar solution at pH > 3 (Kedari et al. 2009).

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Nanopolymer/Dendrimer-Assisted Ultrafiltration Polymer-assisted ultrafiltration (PAUF) is a technology under development for selective concentration and recovery of valuable metal ions from wastewater. PAUF is a combination of ion exchange or chelating property of a functionalized water-soluble polymer and sieving power of an ultrafiltration membrane. As the molecular weight of metal polymer complex is larger than the molecular weight cutoff of the membrane, they are concentrated in the retentate and permeate is free from metal ions. The polymers could be regenerated by changing pH of the solution. With the recent significant advantages in nanoscience and nanotechnology, various nanostructured polymers/dendrimers were devised for a wide range of advanced applications. Examples include the use of polymer nanoparticles as drug delivery devices, polymer nanofibers as conducting wires, and polymer thin film as optoelectronic devices. Like most other nanomaterials, nanostructured polymers also possess interesting mechanical, electronic, optical, and even magnetic properties that are different from those of the bulk materials, depending on their size, shape, and composition. Dendrimers are a new class of polymeric nanomaterials (after linear, cross-linked, and branched polymers) having unique properties such as high degree of branching unit, high density of surface functional group, and narrow molecular weight distribution. They are spheroid or globular nanostructures that are precisely engineered to carry molecule encapsulated in their interior void space or attached to the surface. Size, shape, and reactivity are determined by generation (shell) and chemical composition of the core, interior branching, and surface functionalities. Dendrimers are constructed through a set of repeating chemical synthesis procedures that build up from the molecular level to the nanoscale region under conditions that are easily performed in a standard organic chemistry laboratory. The dendrimer diameter increases linearly whereas the number of surface groups increases geometrically. Dendrimers are very uniform with extremely low polydispersities, and are commonly created with dimensions incrementally grown in approximately nanometer steps from 1 to over 10 nm. The control over size, shape, and surface functionality makes dendrimers one of the “smartest” or customizable nanotechnologies commercially available (Tomalia et al. 1985). Figure 9 provides a schematic of a generalized dendrimer structure. Among the uses of dendrimer the important one is that they can be used to selectively bind to, or react with, a particular element, ion, or molecule of choice. They are very large, yet soluble, macroligands, and well-defined sizes and shapes can be made, with hundreds or even thousands of complexing sites and reactive chain ends. They can also be covalently linked to each other or to other macromolecules to form supramolecular assemblies of various size, shape, and topologies. Dendritic macromolecules can also be functionalized with surface group that make them soluble in selected solvents or bind to selected surface (Diallo et al. 2009). Currently, dendrimers are under investigation as metal-sequestering agents for waste remediation technologies (Diallo 2006). The highly branched molecules with numerous functional groups can be formulated to provide the required properties of water

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Fig. 9 Schematic of a generalized 2D and 3D view of dendrimer structure (Kukowska-Latallo et al. 1996)

solubility, selectivity, high loading capacity, and so on. The use of poly(amido)amine (PAMAM) and poly(propylene)imine (PPI) dendrimers and their derivatives in combination with ultrafiltration technique have shown potential applications in removal of metal ions such as Cu(II), Ni(II), Co(II), Pd(II), Pt(II), Zn(II), Fe(III), Ag(I), Au(I), and U(VI) from dilute aqueous solution (Rether and Schuster 2003; Diallo et al. 2008; Arkas et al. 2003). Removal of U(VI) and Th(IV) ions from aqueous solution by ultrafiltration (UF) and dendrimer-assisted ultrafiltration (DAUF) was investigated (Ilaiyaraja et al. 2014). Regenerated cellulose acetate ultrafiltration membrane was used during ultrafiltration and PAMAM dendrimer chelating agent used as a complexing agent in DAUF. In UF process, removal of U(VI) and Th(IV) metal ions is observed to be based on adsorption/mass deposition on membrane. PAMAM dendrimer contains a large number of nitrogen and oxygen atoms as amino and amide groups on branches and terminal surface showing strong chelating action towards actinides. In DAUF, the water-soluble PAMAM dendrimer effectively forms complex with hydrolyzed U (VI) and Th(IV) species in aqueous solution and gets concentrated in the retentate. In both UF and PAMAM-DAUF, the U(VI) and Th(IV) were more than 95% in the weakly acid medium (pH 5–6). However, in UF, the membrane fouling (plugging of membrane pores) was observed on repeated use. PAMAM-DAUF was shown to be effective in the removal of U(VI) and Th(IV) at pH > 4.

Conclusion Separation of radionuclides from aqueous streams is currently one of the most significant and challenging problems. The main challenge involved is the separation of various long-lived actinide species that exist simultaneously in aqueous streams and their effective separation in the presence of a large excess of competing ionic

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species. Therefore, materials to be used for separation of actinides species are required to be more efficient and specific enough. This chapter revealed the fact that there has been a large increase in utilization of nanomaterials over the last few decades. All above-mentioned studies have dealt with separation of radionuclides using nanomaterials. These studies emphasize that nanomaterials show several interesting aspects for application in nuclear waste disposal and environmental remediation. Due to nanostructural and surface features, there is a tendency for various interactions (hydrophobic, dipole, π–π interactions, formation of hydrogen and other bonds), good sorption capacity, kinetic properties, and thermal and chemical stabilities. Moreover, their modified or functionalized forms are promising candidates for specific actinide separation. The magnetic nanoparticles are incorporated into sorbent to facilitate the easy removal of suspended sorbent from aqueous solution after the extraction of metal ions. This approach promises industrial level scaling up of separation process. However, some nanomaterials are not quite stable under the conditions of ionizing radiation and cannot perform for relatively longer period of time in complex chemical environments. Furthermore, the issues on hazardous nature of nanomaterials itself to environment also need to be evaluated and clarified. The dendrimer-assisted ultrafiltration possesses potential application in decontamination of radioactive wastes. However, the synthesis of dendrimer involves a number of steps that are time consuming and expensive. Hence, attention shall be provided for reducing the number of steps involved in dendrimer synthesis. In these regards, full realization of a true potential of nanomaterials for radionuclide separation requires further studies concentrating on developing highly efficient, selective, radiation-resistant, renewable, economic, and environment-friendly nanomaterials. Therefore this chapter provides a deeper insight into speciation and separation of actinide species in aqueous stream.

Cross-References ▶ Advanced Treatment Technologies ▶ Environmental Treatment Technologies: Adsorption ▶ Management of Radioactive Wastes ▶ Nanomembranes for Environment

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Part V Management of Special Wastes: CO2, CH4, NOX, SO2, Carbon Particles, and Oil Spills

Prospects and Issues of Phosphorus Recovery as Struvite from Waste Streams

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Contents Phosphorus: Role, Issues, and Possibilities . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . P Recovery as Struvite: An Alternative P Fertilizer and Factors Affecting . . . . . . . . . . . . . . . . . . . Struvite: As Nuisance and Issues Concerning Prevention . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Benefits Incurred from Struvite Recovery . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Potential Sources for Recovery of Struvite . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Commercial Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Effects of Nonparticipating Ions on Crystallization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Technology Enhancement Through Process Modification . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Methods of Struvite Recovery . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chemical Precipitation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Electrochemical Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Ion Exchange Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Microbial Biomineralization Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Use of Different Pretreatments for Maximizing Struvite Recovery . . . . . . . . . . . . . . . . . . . . . . . . . . . . pH Modification Using Acid/Base Leaching . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chelating Agent Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Microwave Heating . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Enhanced Biological Phosphorus Removal (EBPR) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Use of Seed in Struvite Crystallization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Alternative Magnesium Sources for Struvite Precipitation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

822 824 825 826 827 834 841 842 842 844 844 847 847 848 849 849 849 850 850 852

S. Kataki (*) Energy Conservation Laboratory, Department of Energy, Tezpur University, Tezpur, Assam, India e-mail: [email protected] D. C. Baruah Energy Conservation Laboratory, Department of Energy, Tezpur University, Tezpur, Assam, India Department of Mechanical and Industrial Engineering, University of South Africa, Pretoria, South Africa © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_19

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Application of Struvite as Soil Fertilizer . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Fertilizer Properties of Struvite . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Effect of Struvite Application on Crop . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

854 854 855 858 859

Abstract

Finite resources of non-substitutable plant nutrient like phosphorus (P) make recovery of it an attractive option of renewed interest from alternative waste sources. In this context, feasibility of struvite recovery (MgNH4PO4. 6H2O), an alternative P fertilizer, is already established from different waste streams with reasonably high ortho-P recovery efficiency (~90%). Feasibility of struvite recovery has been established at laboratory scale for a range of sources of farm, municipal, and industrial origin. Municipal wastewater is the most common struvite recovery source, whereas farm wastes represent easily accessible stock with reliable availability. Depending on the source characteristics, the recovery of struvite may require some process modifications such as addition of P and NH4+ salt and incorporation of pretreatments. However, except for municipal sludge and urine, development of cost-effective, targeted, and environmentally friendly full-scale recovery of struvite is limited due to inherently heterogeneous nature of the sources and unfavorable economics. In recent years, an increasing research concern can be seen toward the techno-economical aspects of the process for development of competent and energy-efficient process from alternative potential struvite sources with incorporation of more efficient method, Mg source, and seed material. Studies on struvite’s application aspect identify its favorable impact on crop though with variation attributed to soil type, plant type, and climate; however, its field-scale long-term impact, its applicability within regulatory limits of fertilizer, and users’ perception remain other concerns. Nevertheless, considering the related benefits of recovery process, struvite recovery appears to be an attractive and feasible pathway provided uncertain aspects are addressed through appropriate research and development. Keywords

Struvite · Phosphorus · Phosphate · Recovery · Seed · Magnesium · Fertilizer · Nutrients · Pretreatments · Anaerobic digestion · Agriculture · Farm waste · Industrial waste · Municipal waste · Ammonium · Calcium

Phosphorus: Role, Issues, and Possibilities Currently, agricultural lands occupy 40% of global land and contribute 4% to the global GDP, making agriculture as one of the most important enterprises in the world (Warnars and Oppenoorth 2014). In recent years, agricultural growth has gained new significance and dimension, as appropriate intensification of agricultural sector is of

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vital importance to ensure food security for growing population. Fertilizer sector constitutes the backbone of agricultural intensification. In long-term conventional cultivation practices, substantial amount of plant essential nutrients is removed through crop. It is not possible for soil to meet crop nutrient requirement without external supplementation of proper nutrient inputs. Thus fertilizer has been and will continue to be one prime input in global agriculture to achieve self-sufficiency in food grain production. Currently, an average of 40–60% of the world’s food production is attributable to fertilizer inputs (Roberts 2009). Phosphorus (P) is one such non-substitutable fertilizer, and agroecosystems account for 80–90% of the world’s total P consumption (Childers et al. 2011). Current global P demand is forecast to grow at an annual rate of 1.9% over the period 2013–2018 with a more stabilized consumption in the developed countries, but demand is seen increasing in the developing world (Heffer and Prud’homme 2010, 2014) (Fig. 1). Presently the only source of commercial P fertilizer is natural phosphate rock (both sedimentary and igneous) with no known chemical or technological substitute. Rock P reserve is mostly spread in limited number of countries, viz., China, the USA, Morocco, and Russia, contributing 75% of world total production (Childers et al. 2011; Heckenmüller et al. 2014). However, there are some uncertainties associated with the actual extent of commercially viable global phosphate stock (Vaccari 2009). Further, increasing population, uncertain backup of global stock, and progressive global P consumption are expected to put more pressure on dwindling accessible P supplies. P appears as a key limiting nutrient in 40% of the world’s arable soils due to its slow diffusion and high fixation with other chemical constituents (Ca, Fe, Al, Mg, K) that becomes available through chemical weathering 25

Phosphorus(Mt)

20

15

10

5

0 1960 1964 1966 1972 1976 1980 1984 1988 1992 1996 2000 2004 2008 2012 2016

Year Developing countries

Developed countries

World

forecast

Fig. 1 Global phosphorus fertilizer consumption (Adapted from Heffer and Prud’homme 2010)

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Fig. 2 Issues favoring phosphorus recovery

(Shen et al. 2001; Vance 2001). All these issues (Fig. 2) argue in favor of recovery of P with the help of cost-effective, energy-efficient, and environmentally compatible means from non-conventional P-rich sources to produce products with enhanced nutrient values competitive with mineral fertilizers. P recovery technologies are based on the principle of fixation of P through biological method or chemical precipitation by metal salts (Le Corre et al. 2009).

P Recovery as Struvite: An Alternative P Fertilizer and Factors Affecting The most common method of P recovery is through production of mineral or salt precipitates from P-rich sources. Struvite or magnesium ammonium phosphate (MAP/MgNH4PO4˙6H2O) is such P-rich mineral which is by mass 44% crystal water, 39% phosphate, 10% magnesium and 7% ammonium (Gell et al. 2011). Precipitation of struvite needs the presence of three ionic species, magnesium (Mg2+), ammonium (NH4+), and orthophosphate (PO43) in an alkaline solution in equimolar (1:1:1) concentrations (Rahaman et al. 2008). pH, ionic strength of participating molecules, the presence of impurities or nonparticipating ions, mixing energy, residence time for crystallization, and the type of crystallization reactor are the factors that govern the precipitation of struvite in P-rich sources (Doyle and Parson 2002; Nelson et al. 2003; Le Corre et al. 2007; Koralewska et al. 2009; Hutnik et al. 2011) (Fig. 3). The basic chemical reaction to form struvite can be expressed using Eq. 1. It was estimated that globally 0.63 million tons of P per year could be generated as struvite from the wastewater treatment plants, which would be sufficient to reduce phosphate rock mining by 1.6% (Shu et al. 2006).

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Fig. 3 Defining factors affecting struvite precipitation

Mg2þ þ PO4 3 þ NH4 þ þ 6H2 O , MgNH4 PO4 6H2 O # pKs    ¼ 12:6 25 C

(1)

(Li et al. 1999) pH is considered to be one of the most significant factors of struvite precipitation as solubility of struvite is highly pH dependent. Minimum solubility of struvite is found within the pH range of 8–11, and struvite solubility decreases with increase in pH (Ohlinger et al. 1998). At constant pH, higher ionic strength or supersaturation of participating ions is reported to induce nucleation rate positively, reducing induction time (Bouropoulos and Koutsoukos 2000). Mixing energy is another controlling factor for crystal growth, increase of which reduces induction period, when supersaturation is constant (Ohlinger et al. 1999; Bhuiyan et al. 2007). On the other hand, the presence of foreign ions can inhibit formation of struvite by competing with struvite-forming participating molecules and can decrease purity of recovered struvite by coprecipitation (Jones 2002; Le Corre et al. 2005).

Struvite: As Nuisance and Issues Concerning Prevention Struvite formation appears as a physical nuisance mostly in wastewater treatment plants (in pipes, heat exchangers) where it occurs spontaneously when appropriate conditions of molar ratio of Mg2+, NH4+, and PO43, pH, and mixing energy are achieved (Bhuiyan et al. 2007). In wastewater plants, areas most affected by struvite scale formation are anaerobic digester units, digester liquor discharge line, heat exchangers, and centrifuge dewatering units downstream of the digester system. During anaerobic digestion of wastewater sludges, NH4+ is released from degradation of nitrogenous material in organic wastes creating suitable conditions for struvite formation (Al-Seadi and Lukehurst 2012; Bhuiyan et al. 2007). Areas of a wastewater treatment where there is an increase in turbulence (pumps, aerators, and pipe bends) are more prone to struvite scale formation as aeration helps to increase pH by stripping off CO2 (Borgerding 1972). Thus, overall system

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Fig. 4 Struvite control methods and limitations associated

efficiency and operational costs are affected due to struvite scaling (Doyle and Parsons 2002; Jaffer et al. 2002; Petzet and Cornel 2011). There are some control measures employed to control formation of struvite as shown in Fig. 4. However, none of the methods are completely affective to alleviate the problem and are associated with drawbacks, viz., cost affectivity in terms of labor and materials and time, increase in total sold content, non-suitability of sludge disposal, difficulty in P recovery from Fe/Al salt (Ohlinger et al. 1998; Wu et al. 2005), concern about agricultural application of remaining sludge (Wu et al. 2005), and multiple factor-dependent affectivity (Wu et al. 2005) (Fig. 4). Therefore, failing to achieve at an working mitigation strategy to deal with the problem, a controlled, designed, and intentional struvite precipitation has been attempted leading to consideration of “struvite production” a resource recovery method.

Benefits Incurred from Struvite Recovery Struvite recovery, as a method of nutrient recycling, has expansive multifaceted benefits and is receiving support from R&D. The recovery process of struvite can create multiple benefits directly or indirectly as shown in Fig. 5. Apart from development of a product with enriched value, P recovery process also adds to some associated benefits in terms of environmental quality improvement of P-rich source, increase in dry matter content, volume reduction of waste handled and supplementation to chemical fertilizer, closing of open P cycle, and enhancing performance efficiency of waste treatment facilities (Woods et al. 1999; Shu et al. 2006; Gell et al. 2011; Estevez et al. 2014). Integration of nutrient recovery with waste management is expected to result in cost-effective relocation of excess nutrients from otherwise polluting effluents (Burns and Moody 2002).

32

Prospects and Issues of Phosphorus Recovery as Struvite from Waste Streams

827

Fig. 5 Benefits derived from struvite recovery technology

Potential Sources for Recovery of Struvite The biogeochemical cycle of P is an open cycle. It has been reported that only 15% of the total P extracted from mines and used for food production is eventually consumed by humans, whereas remaining P is lost to the environment (Roy et al. 2006; Suh and Yee 2011). During crop cultivation and livestock, meat, and dairy production, more than 66% of the total P extracted is lost, and remaining 19% is lost through household food waste, mining waste, and fertilizer manufacturing waste (Karunanithi et al. 2015). Thus, during the cycling of P in the terrestrial and aquatic environment, the lost P ends up in some easily accessible and abundant natural sources (Atkinson et al. 2010). Therefore, a sustainable approach for effective P management is recycling back of P from various identified P-rich sources. To identify such sources, composition of source, abundance, and need for their management or treatment remain the main driving forces. In literature a range of waste sources has been investigated to understand the feasibility struvite recovery. Depending upon their occurrence or origin, these sources can be categorized into three groups, viz., farm waste, industrial waste, and municipal waste. Various waste sources of farm, industrial, and municipal origin reported in literature are shown in Tables 1, 2, and 3 along with waste characteristics and process conditions. Farm wastes: Manures are rich in P and NH4+, which is favorable for struvite recovery. It has been reported that more than 70% of consumed animal feed is excreted as urine and feces which contain organic matter, N, P, K, and other micronutrients (Barnett 1994). Thus manures represent inexpensive and most abundantly available sources for recovery of struvite, and struvite recovery could serve as

NR 1405 7732

532 234 732–931 985

NR 19 NR

72 42 30–56 161

55–139 1013–1426 3000 mg kg1 1900 mg kg1 DM DM

255–519

NR

NH4-N (mg L1) NR 1318

NR not reported, DM dry matter

Swine compost

Cattle urine Swine manure

Dairy manure

Sources Poultry manure

PO4-P (mg L1) 572 NR

Bittern MgCl2 MgCl2 Struvite pyrolysate MgCl2 MgCl2

Mg source MgCl2 MgO, MgSO4, MgCl2 MgCl2, Mg (OH)2 MgCl2 MgCl2 Brine 7.5–8.5 9 9 8–8.5

– KH2PO4 KH2PO4 H3PO4 10 7.3

8.5 7.2 9

– – –

Na3PO4 H3PO4

8.5–9.2

pH 9 9

Additional chemical – NaHPO4, KH2PO4, H3PO4 Na2HPO4

99 NR

73 89 97 96

 82% 69 NR

NR

P recovery (%) 91 NR

87 NR

NR 70 90 80

NR NR NR

95

NH4+ recovery (%) NR 85

Table 1 Potential sources of struvite recovery of farm origin (Adapted from Kataki et al. 2016a with modification)

Zhang et al. (2012) Fukumoto et al. (2011)

Suzuki et al. (2007) Perera et al. (2007) Ryu and Lee (2010) Huang et al. (2011a, b)

Zhao et al. (2010) Shen et al. (2011) Prabhu and Mutnuri (2014)

Reference Burns et al. (2001) Yetilmezsoy and SapciZengin (2009) Demirer et al. (2005)

828 S. Kataki and D. C. Baruah

520 3500

NR

56 4535

208–426

43–127

NR

100

286

Rare-earth wastewater

1400

24

AD effluent of molassesbased wastewater Semiconductor wastewater Anaerobic effluent from potato processing Coking/coke oven wastewater

210–220

6 2320

83–208

5.5–10

3490

20–368

NR

TN/NH4-N (mg L1) 119–1076

Carmine dye wastewater

Textile industry wastewater Abattoir wastewater/ meat packing effluent

Sources Leather tanning wastewater

Total P/PO4P (mg L1) 2.5–8

Brucite

MgCl2

MgCl2

MgCl2

MgCl2

MgCl2

MgO

MgCl2

MgCl2

MgCl2

Mg source MgCl2

8.5–9.5

97

NR

9–9.5

19–89

Na2HPO4, Ca (H2PO) H3PO4

8.5–8.7

–

70

NR

9

–

NR

9.5

8–9

–

100

NR

NR

NR

P recovery (%) NR

Na2HPO4

8.5–9

9.5

NaH2PO4 Na2HPO4

9

8–9.5

KH2PO4 Na2HPO4

pH 9

Additional chemical Na2HPO4

Table 2 Potential sources of struvite recovery of industrial origin (Adapted from Kataki et al. 2016a with modification)

95

95

84

NR

98

78–95

89

89.5

 78

84

NH4+ recovery (%) 89

Prospects and Issues of Phosphorus Recovery as Struvite from Waste Streams (continued)

Reference Tunay and Kabdasli (2001) Kabdasli et al. (2000) Tunay and Kabdasli (2001) Kabdasli et al. (2009) Chimenos et al. (2003) Türker and Celen (2007) Kim et al. (2009) Moerman et al. (2009) Zhang et al. (2009) Kumar and Pal (2013) Huang et al. (2011a, b)

32 829

415

17.4

10.8

Yeast industry

Wastewater

528

161

NR

NR

4.45%

Cola beverage

1197

NR

Fertilizers industry wastewater

1128

TN/NH4-N (mg L1) 550

36

Total P/PO4P (mg L1) NR

7Aminocephalosporanic acid wastewater

Sources Nylon wastewater

Table 2 (continued)

MgSO4

MgCl2

MgCl2

Mg source Brucite, MgSO4 MgCl2, MgO, MgSO4 Struvite pyrolysate MgCl2

H3PO4

Na2HPO4

NH4Cl

9–11

NH4Cl

9

9.5

9.5

NR

83

97

99.5

NR

NR

9

9.5

P recovery (%)  94

pH 8.5

Additional chemical H3PO4, Na2HPO4 H3PO4, Na3PO4, NaH2PO4 –

87.55

81

NR

NR

97

 74%

NH4+ recovery (%)  88

Yu et al. (2012) Matynia et al. (2013) Foletto et al. (2013) Khai and Tang (2012) Uysal and Demir (2013)

Reference Huang et al. (2012) Li et al. (2012)

830 S. Kataki and D. C. Baruah

Sewage sludge ash

Municipal wastewater

Landfill leachate

Sources Human urine

TN/NH4-N (mg L1) 6963

40

NR/7220 NR 2540

3200–4990 245 2750–2900

1795 2600 1150

1400

193/168

100–700 949 795 NR

Total P/PO4-P (mg L1) 240

460

206 NR 197

156–194 416 NR

10.5 NR 200

24

273

50–170 21 54 15–27%

MgCl2 MgCl2 MgCl2 MgCl2

MgCl2

MgCl2

MgO MgO, MgCl2 MgCl2,MgO, MgSO4 MgCl2 MgO MgO

MgCl2 MgO Mg anode

MgO, MgCl2

Mg source MgCl2

95 92 92 99 100 87

NR

9.4 8–11 9.2–9.5 8.9 9.1–9.3 9 9 9 9 9.5

8.5 9.14 8.5 9 8.5–9 10

– Na2HPO4 – – – – Na2HPO4, Ca (H2PO), H3PO4 KH2PO4 H3PO4 Triple superphosphate, H3PO4 H3PO4 – – H3PO4 KH2PO4 NH4Cl

 95% 95 NR 97

87

85 95–100 84

95

pH 9

Additional chemical –

P Recovery (%) 96

Table 3 Potential sources of struvite recovery of municipal origin (Adapted from Kataki et al. 2016a with modification)

Türker and Celen (2007) Demirer and Othman (2009) Pastor et al. (2010) Uysal et al. (2010) Latifian et al. (2012) Xu et al. (2012)

 97

NR 89 NR NR

46

Kim et al. (2006) Iaconi et al. (2010) Suschka and Poplawski (2003)

Reference Ronteltap et al. (2007) Wilsenach et al. (2007) Liu et al. (2008b) Ganrot et al. (2007) Hug and Udert (2013) Morales et al. (2013) Latifian et al. (2013) Li and Zhao (2003) 87  95 98

NR 90 NR

95 50 NR

NR

NH4+ recovery (%) NR

32 Prospects and Issues of Phosphorus Recovery as Struvite from Waste Streams 831

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S. Kataki and D. C. Baruah

an alternative to deal with issues related to their excess nutrient load. The average total P content in dairy, poultry, and pig manures are 9.3, 18, and 39 g kg1, respectively (Barnett 1994; Shen and Shen 2011). Among farm-based wastes, successful struvite recovery has been reported in dairy manures, swine manure, poultry manure, and cattle urine. Though P load of manures makes these suitable sources, P fractionation between available and non-available is an issue. Manure contains P mostly as inorganic P (60–90%). The rest of the P contained in manure is in the organic P fraction, and it remains the most poorly studied potential source for P recovery (Barnett 1994). Further, particulate P that remains mostly bound to mineral component is unavailable for recovery (Sharpley and Moyer 2000; Chapuis-Lardy et al. 2003). Hence, the use of some pretreatment methods (acid/base leaching, chelating agent treatment, microwave treatment, anaerobic digestion) in order to release P into an available form is helpful for making struvite recovery more effective in manures (Shen et al. 2001; Szogi et al. 2008; Moody et al. 2009; Qureshi et al. 2008). However, in farm-based waste, though availability and abundance of P are ensured, economics of incorporation of pretreatment methods has to be considered for effective struvite recovery. Industrial wastes: Most of the input nutrient load in industrial processes makes their way into the effluent coming out. However, the environmental regulations in most countries are becoming stringent in the use of appropriate technology to reduce the P content of waste streams before safe disposal. In this context, recovery of P from highstrength industrial effluents is a viable strategy for their safe disposal by stripping both P and NH4+ concentration (Altinbas et al. 2002). Successful struvite recovery has been reported from wastewater from tanning, food processing, slaughterhouse and meat packing, tannery, textile, dye, cola beverage, fertilizers and agrochemical, semiconductor, rare earth, yeast, and coking industries (Table 2). However, orthophosphate concentrations of these sources remain in the lower side compared to manures, which makes external supplementation of P sources a must in most of the cases to make the crystallization effective. Similarly for sources with less NH4+ (e.g., such as cola beverage, fertilizer industry wastewater), NH4+ is needed to add externally (Xu et al. 2012; Hutnik et al. 2012; Foletto et al. 2013). Literature reported direct use of industrial wastewater for struvite recovery without pretreatment. Municipal waste: It is the spontaneous precipitation of struvite in municipal sewer system that led to controlled struvite recovery attempt using other wastes also. Globally huge quantity of municipal water is produced annually which contains high amounts of nutrients and organic materials (Guo et al. 2010). Phosphorus from both domestic and industrial sources enters the municipal wastewater treatment plants. Phosphate can be recovered from the liquid phase (supernatant liquor of anaerobic digestion), sludge phase, and mono-incinerated sludge ash as shown in Fig. 6 (Cornel and Schaum 2009). Ninety percent of the incoming phosphorus load, from the wastewater, is incorporated into the sewage sludge. The phosphorus recovery rate from the liquid phase can reach up to 40–50%, while recovery rates from sewage sludge and sewage sludge ash can reach up to 90% (Cornel and Schaum 2009). The economically feasible recovery requires a liquid phase containing 50–60 mg l1of ortho-P (Cornel and Schaum 2009).

32

Prospects and Issues of Phosphorus Recovery as Struvite from Waste Streams

Primary sedimentation

Activated sludge Anaerobic pretreatment

1a

Anaerobic digestion

833

Secondary sedimentation

aerobic

anoxic

Anaerobic posttreatment

2a

2b

Dewatering

Incineration

3

1b

1a: side stream after anaerobic treatment; 1b: dewatering unit after anaerobic digestion; 2a: sludge from the digestor before dewatering; 2b: sludge from the digester after dewatering; 3: sewage sludge ash after incineration

Fig. 6 Possible locations for phosphorus recovery in wastewater treatment plant (Adapted from Desmidt et al. 2015)

Prior to struvite recover, fixation of P in wastewater sludge particles requires use of some pretreatment methods (acidic, basic, microwave, enhanced biological phosphorus removal) (Stark 2005; Pan et al. 2006; Pastor et al. 2008). Landfill leachate (Li and Zhao 2003; Kim et al. 2006; Iaconi et al. 2010) and human urine (Ganrot et al. 2007; Morales et al. 2013, Hug and Udert 2013) are also prospective sources for struvite recovery, which can be directly used without any pretreatment. Urinederived struvite is found to be free of heavy metal, and alkali addition is not required due to its inherent alkaline nature (Hug and Udert 2013; Morales et al. 2013). Combustion ash of municipal sewage sludge could be another prospective struvite source, but recovery process needs incorporation of mechanical, thermal (incineration), or chemical pretreatments (Hong et al. 2005; Xu et al. 2012). From Tables 1, 2, and 3, it is seen that, when the constituting elements are found to be limiting, the requirements are fulfilled by external addition of P and NH4+ salts. Most of these studies used stirred batch reactors which are frequently used to provide necessary mixing energy at small scale due to operational simplicity and reasonable recovery efficiency, whereas, at pilot scale, fluidized bed reactors are generally used. The pH range found to be favorable for struvite precipitation varies from of 8 to 11, which is generally adjusted using NaOH, MgO, KOH, or CO2 stripping. Recovery efficiency of P and ammonium as struvite is usually >90%. To understand the measure of suitability of a waste for struvite recovery based on its composition, a Feedstock Suitability Index (FSI) has been developed, which has been reported in previous publication by the authors (Kataki et al. 2016a). The index takes into account concentration of P, ammonia as favorable component, whereas concentrations of inhibiting ions such as Ca and Fe are considered to affect FSI

834 Table 4 Ranking and Feedstock Suitability Index (FSI) of potential struvite recovery sources (Adapted from Kataki et al. 2016a with modification)

S. Kataki and D. C. Baruah

Rank 1 2 3 4 5 5 6 7 8 9 10 11 12 13 14 14 15 16 17 18

Feedstock Rare-earth wastewater Fertilizers industry wastewater Cochineal insects processing wastewater (carmine dye industry) Nylon wastewater Human urine Leather tanning wastewater Cola beverage Landfill leachate Coking/coke oven wastewater Poultry manure wastewater AD effluent of molasses-based industrial wastewater Dairy manure Semiconductor wastewater Pharmaceutical wastewater Municipal wastewater Swine wastewater Yeast industry wastewater Anaerobic effluent from potato processing industries Textile printing industry wastewater Abattoir wastewater/slaughterhouse wastewaters

negatively. If compositional status of a particular waste source is known, the FSI index is expected to be useful in identifying comparative suitability of different waste sources for struvite production. As per the FSI index, rare earth wastewater is found to be the most suitable source; however, its abundance could be an issue. Among the top potent sources of struvite showing high FSI, urine appears to be a promising source considering its composition and reliable availability (Table 4).

Commercial Processes Several studies have evaluated the potential technologies of P recovery at bench and pilot scales, and only a few has been incorporated in full-scale system. A limited number of commercial struvite recovery units handling urine and municipal wastewater sludge are in operation in countries like Nepal, Japan, Canada, England, the USA, Australia, Germany, the Netherlands, and Italy. In these commercial processes, to make the recovered struvite competent with chemical fertilizer, the recovered struvite is washed, dried, purified, concentrated, sieved/sorted for granulometry, milled or pelletized, and packaged. Descriptions of some commercially available struvite recovery technologies are given below, and Table 5 summarizes the overview of these processes.

Digested sludge from EBPR WWTP

Rejection water of wastewater treatment

Centrates from digested sludge from EBPR WWTP

Centrates from digested sludge from EBPR WWTP

Airprex

Phospaq

Pearl

Crystalactor

Food processing, farm (dairy, swine) waste

Multiform

Seaborne

Phosnix

Digested sludge Dairy industry, potato processing industry, pharmaceutical, municipal wastewater Wastewater after digestion or sludge treatment Digested sludge

Stuttgart NuReSys

Potato processing

Input

Process

2000/Seaborne Environmental Research Laboratory 2012/Multiform Harvest Inc.

1987/Unitika Ltd.

2003–2004 2006/Akwadok bvba

2007/University of British Columbia/ Ostara DHV Water BV

2007/Paques BV

Berliner Wasserbetriebe

Year of initiation/ developed by

USA

Germany

Japan

Germany Germany, Belgium, the Netherlands

The Netherlands

Canada, UK

The Netherlands, Germany The Netherlands, Germany

Available/ place of operation

MgCl2/NaOH

MgOH/ NaOH H2SO4, Na2S, NaOH, MgO, flocculent

– MgCl2/NaOH

Sand, NaOH, H2SO4, Ca (OH)2

MgCl2/NaOH

80

90

80–90

– 72–90

70–80

85%

80%

80–90

MgCl2, CO2stripping MgO/ CO2stripping

P recovery efficiency

Mg source/pH adjustment

0.5–1 mm –

–

–

–

– –

0.8 mm

–

–

Separation of heavy metal problematic

–

No residue Need of high turbulence Compact high-purity crystal – Modular assembly

Struvite complies with EU standard Combined phosphate and COD removal –

–

–

0.7 mm

Remarks

Pellet size

8.2–8.8

– 8–8.5

8–8.5

7.25

8.2–8.3

8

pH

Prospects and Issues of Phosphorus Recovery as Struvite from Waste Streams

Fluidized bed

Fluidized bed CSTR

– CSTR

Fluidized bed

Fluidized bed

CSTR

CSTR

Reactor type

Table 5 Overview of commercial processes available for struvite recovery

32 835

836

S. Kataki and D. C. Baruah

Anaerobic effluent MgCI2 Air

NaOH

Effluent STRIPPER CRYSTALIZER

Crystal Purge

Fig. 7 Schematic overview of the NuReSys process (Adapted from Moerman et al. 2009)

NuReSys (Belgium): NuReSys (Nutrients Recovery Systems) is a Belgianbased company founded in 2011. The technology has been applied in case of digested sludge, dairy industry, and food processing industry. The main common features of the reactors are air stripper, crystallization reactor equipped with a top entry mixer and a transient quiescent settling zone, and pH control (using NaOH) (Fig. 7) (Moerman et al. 2009). The struvite pellets are harvested by intermittent purging, and up to 76% of P removal could be achieved (Moerman et al. 2009). Pearl ® Technology (North America, UK): The US patented Pearl ® Technology for struvite (commercially known as Crystal Green) was developed by the University of British Columbia, Canada. In Oregon (USA) the first commercial full-scale plant was installed in 2009. Ostara process is primarily applied to recover phosphorus from dewatering liquor and centrate after anaerobic digestion. Ostara technology utilizes upflow fluidized bed reactor comprising a liquid/solid separation device, a settling tank and a reaction tank, and a piping system and an injector arranged to inject chemicals (Fig. 8). Phosphorus and ammonia recovery using this technology was reported to achieve 80–90% and 14–42%, respectively (Britton et al. 2009; Oleszkiewicz 2015). Phospaq™ (The Netherlands): The technology was developed by Paques, the Netherlands. The technology is commercially being used in Lomm (in potato processing industry) and Olburgen (in potato processing industry and municipal wastewater plant) of the Netherlands (Remy 2013). A PO43 recovery efficiency of 75% and NH4+ recovery of 19% could be achieved using this technology (Remy et al. 2013). The struvite complies with EU standards for fertilizer. Recovery of struvite by

32

Prospects and Issues of Phosphorus Recovery as Struvite from Waste Streams

Magnesium Chloride (MgCI2)

837

Treated centrate effluent

Caustic

Centrate Influent

Crystal GreenTM Product

Fig. 8 Schematic overview of the Pearl technology (Adapted from Britton et al. 2009)

PHOSPAQ™ is feasible from a daily P load >100 kg P/d, ortho-P load >50 mg/l, and NH4+ load >200 mg/l (Paques 2016). The advantage of the process is combined phosphate, ammonia, and COD removal in one reactor. Phosnix (Japan): The Phosnix process was developed by Unitika Ltd. (Katsuura 1998) and has been in operation in Japan since 1987 (Münch and Barr 2001). The reactor consists of an aerated column containing a fluidized bed of granulated struvite, into which returned water from sewage sludge treatment is fed (Fig. 9). Recirculation of the effluent back to the initial wastewater treatment reduces the requirement for chemical supplementation (Ueno and Fuji 2001; Nawa 2009). Seaborne (Germany): The Seaborne process was developed by the Seaborne Environmental Research Laboratory, Germany. This combined treatment technology with multiple unit operations uses anaerobic digestion and sludge acidification by sulfuric acid to extract nutrients and heavy metals from the solid phase, followed by heavy metal precipitation as metal sulfide. After that, struvite is precipitated by addition of magnesium hydroxide and NaOH (to increase the pH to pH 9) in a continuous stirred tank reactor (Fig. 10). However, sometimes, ineffective removal of heavy metals may result in struvite with impermissible heavy metal concentration through this technology (Müller et al. 2007; Bergmans 2011). AirPrex™ (Germany, Netherlands): The AirPrex™ technology was developed by Berliner Wasserbetriebe and was implemented at the Wassmannsdorf wastewater treatment plant (Heinzmann and Engel 2006; Forstner 2015). A continuous stirred tank reactor is utilized in this technology where sludge is lifted upward by air bubbles in the aerated zone. Struvite is continuously removed from the bottom of

838

S. Kataki and D. C. Baruah

pH Treated Water Digestion Tank

Separation Zone

Struvite Hopper

Mg(OH)2 Storage Tank

P

Struvite Separator Recovered Struvite

NaOH Storage Tank

P

Digested Sludge

Granule Formation Zone

Dehydrator

Raw Water

Fines are returned to Granule Formation Column

Struvite B Blower

P Feed Pump Storage Tank

Sold as Raw Material for Fertilizer

Fig. 9 Schematic overview of the Phosnix process (Ueno and Fuji 2001) Flue gas

Fuel Ash

acid

incineration

Digested sludge

Drying Polymer

Ammonium- NaOH sulfate

acid

Centrifuge Wasterwater to the inflow of the wwtp

org. Residues

stripping

Filter

Digester Gas

Gas desulphurised to the block and heat power plant NaOH Mg(OH)

Filter

2

Filter

Centrifuge

Residues (Heavy MetalSulfides)

MgNH4PO4

Fig. 10 Process flow sheet of the Seaborne process at the (Müller et al. 2007)

tank. Sand washing equipment is utilized to ensure cleaning and purification of the recovered struvite (Desmidt et al. 2015; Oleszkiewicz 2015). These plants were reported to remove 80–90% of phosphate from the liquid phase of the digested sludge (Desmidt et al. 2015) (Fig. 11).

32

Prospects and Issues of Phosphorus Recovery as Struvite from Waste Streams

839

Digested Sewage Sludge

MgCI2

AirLife Reactor

Polymer

Decanter

Air

Dewatered Sludge

Sludge Liquor

Struvite sludge Water

Sand Washer

Process Water

Struvite

Fig. 11 Schematic overview of the AirPrex technology (Adapted from www.p-rex.eu)

Multiform (America): The patented Multiform technology was developed by Multiform Harvest Inc. The technology is commercially operational at wastewater treatment plants in Boise, Idaho, and the city of Yakima, Washington. The technology is also suitable for food processing and swine farm waste. The technology is reported to achieve an 80% and a 20% reduction in phosphate and nitrogen from wastewater, respectively (www.multiformharvest.com). This reactor consists of a conical shaped fluidized bed with no recycle flow (Fig. 12). Struvite pellets are collected from bottom of the reactor (Oleszkiewicz 2015). Bowers and Westerman (2005) demonstrated from laboratory-scale experiments that phosphorus removal process was optimal at pH 7.56 with 60 mg l1 Mg using this technology. Stuttgart: The technology was developed at the Institute of Sanitary Engineering, Water Quality and Solid Waste Management of the University of Stuttgart, Germany. The process is distinguished by the fact that municipal sewage sludge from wastewater treatment plants with simultaneous phosphate elimination with iron salts could be used without any changes in the process of wastewater purification. Though not much information about the field-level application is available about this technology, more information is available at www.iswa.uni-stuttgart.de. The pilot plant consists of two batch tanks, a sedimentation tank, and a chamber filter press as well as holding tanks and dosing equipment for the operational resources (Fig. 13).

840

S. Kataki and D. C. Baruah

Fig. 12 Schematic overview of the multiform technology

Crystalactor: This patented Crystalactor technology was developed by DHV, the Netherlands. The Crystalactor is a cylindrical fluidized bed reactor where wastewater is pumped in an upward direction, achieving up to 95% of P reduction (Fig. 14). The reactor is partially filled with a suitable seed material like sand or minerals. High crystallization rate is also reported to be achieved with concentrated solutions (>100 mg P l1) (Oleszkiewicz 2015); however, the influent rate must be constant during the treatment, as an increase can lead to an increase of secondary nucleation or a decrease not sufficient to maintain particles in a fluidized state (Regy et al. 2002). Furthermore, some amorphous phosphate is lost in suspended form; therefore, a dual-media filtration is required, which incurs additional capital and operational costs.

32

Prospects and Issues of Phosphorus Recovery as Struvite from Waste Streams

841

Fig. 13 Schematic overview of the Stuttgart technology

Fig. 14 The Crystalactor process flow diagram

Effects of Nonparticipating Ions on Crystallization Most waste sources comprise a heterogeneous mix of nutrients and ions, and certain nonparticipating ions can limit the struvite precipitation process. In typical wastes P can exist in particulate or suspended form and in soluble and insoluble form, often in association with other components (Le Corre et al. 2005; Marti et al. 2008). There are various impurities including aluminum ions, alkali metal ions (potassium, sodium), alkali earths (calcium), transition metals (iron, copper, zinc), anions (sulfates, chlorides, nitrates, fluorides, carbonates), and organic impurities (lactic acid) which impact on crystal growth kinetics (Table 2). These soluble cations provide background competition to precipitate Mg or Ca phosphates and therefore have

842

S. Kataki and D. C. Baruah

potential to contaminate the product or decrease the product yield. Metal ions mostly compete for phosphate ions and coprecipitation of their salts (mostly phosphate or hydroxide) salts along with struvite can reduce product purity (de Bashan 2004). Further, crystals growth is inhibited because of blockage of active growth sites through adsorption of impurity ions onto the surface of struvite crystals (Jones 2002; Kabdasli et al. 2006). Crystal size can be decreased by up to 46% in the presence of impurities like calcium, iron, and nitrates (Hutnik et al. 2011). The ions can affect negatively the growth rate and can lengthen the induction time preceding the first occurrence of crystals (Koutsoukos et al. 2003; Kabdasli et al. 2006). Impurities in solution are also known to affect the growth rates of crystalline compounds due to blocking of sites where crystals could formed, thus inhibiting the increase of crystal size (Jones 2002). In waste sources like animal manures, where Ca levels are relatively high, Ca ions can interact with phosphate or carbonate ions to form calcium phosphates (usually as apatite, hydroxylapatite) or calcium carbonates (usually calcite) (Le Corre et al. 2005). Interference from calcium can be minimized by either thermodynamically driven redissolution of calcium phosphate or by removing it via chemical precipitation at elevated pH (Huichzermeier and Tao 2012) (Tables 6, 7, and 8).

Technology Enhancement Through Process Modification Though laboratory feasibility for struvite precipitation is shown from a range of sources, full-scale installations are still limited. At present, municipal wastewater sludge and human urine are the two sources where commercial struvite recovery is carried out. Successful execution of recovery possesses some difficulty related to heterogeneous characteristics of source, chemical input, and recovery efficiency, making the overall struvite economics unfavorable. Therefore, in the recent literature, there is an increasing concern toward the techno-economical aspects of the process to increase the process efficiency and cost affectivity such as improved and competent method of production, effect of different seed material, and alternative and recyclable magnesium sources.

Methods of Struvite Recovery With time some advanced techniques have come up using established principles of electrochemistry, ion exchange separation, and biomineralization by microbes. The new approaches are developed to make the process more efficient in terms of recovery, which are discussed below.

32

Prospects and Issues of Phosphorus Recovery as Struvite from Waste Streams

843

Table 6 Effect of nonparticipating ions on struvite precipitation (Adapted and expanded from Kataki et al. 2016b) Ion Phosphocitrate

CO32

Effect on struvite recovery Inhibition was dose dependent leading to complete cessation at higher concentration Phosphocitrate induces very strong, crystal facespecific inhibition of struvite Crystal growth is slowed Disrupt struvite growth and formation directly through interference with the molecular growth processes on crystal surfaces Formation of amorphous Ca phosphate when Mg/Ca >1:1 Decrease in struvite purity when Ca/Mg >1 Precipitation of Ca phosphate at pH >10 Significant reduction in struvite precipitation (by 37%) in municipal sludge containing 59 mg L1 Ca compared to sludge with 10 mg L1 Ca Decrease in crystal size with increase in Ca concentration and formation of tubular crystal Coexistence of hydroxylapatite in product Insignificant change in P recovery efficiency Loss of product purity at Ca/P >0.5:1 Increased fine (20%) Decrease in homogeneity in product Formation of tubular crystal With increase in NO32 concentration and decrease in mean crystal size (by 29%) Formation of tubular crystal Moderate increase in crystal size (by 6%) The presence of Cu hydroxide in product Formation of tubular crystal

Reference Wierzbicki et al. (1997)

Downey et al. (1992)

Le Corre et al. (2005) Wang et al. (2013) Pastor et al. (2008) Hutnik et al. (2011) Huichzermeier and Tao (2012) Lee et al. (2013) Kabdasli et al. (2006) Kabdasli et al. (2006) Huichzermeier and Tao (2012) Kabdasli et al. (2006) Kozik et al. (2011) Hutnik et al. (2012)

Hutnik et al. (2013) Hutnik et al. (2013) (continued)

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Table 6 (continued) Ion Zn K Al Humic substance (fulvic acid)

Herbal extracts

Effect on struvite recovery Appearance of Zn impurity in product as hydroxides, phosphates, other salt Appearance of K impurity in product as hydroxides, phosphates, other salts Appearance of Al impurity in product as hydroxides, phosphates, other salts Inhibition effect is pH dependent Inhibition is reduced with increase in Mg/P ratio Coprecipitation of humic substance reducing purity Change in crystal morphology Juice of Citrus medica Linn and the herbal extracts of Commiphora wightii, Boerhaavia diffusa Linn, and Rotula aquatica Lour were tested and found to be potent inhibitors

Reference Kozik et al. (2013) Kozik et al. (2013) Kozik et al. (2013) Zhou et al. (2015)

Chauhan and Joshi (2013)

Chemical Precipitation Chemical precipitation is the most common method (>95% literature) of struvite precipitation. In these types of processes, struvite is crystallized in the reactor by the addition of chemicals, to reach the optimum molar ratio Mg/P. The pH required to set off the nucleation is typically adjusted by alkaline addition, while a propeller is used to mix the solutions and favor the occurrence of struvite crystals. A settling zone is integrated to the reactor to allow for the accumulation of particles. To provide adequate mixing energy, stirred batch reactors are most frequently used, particularly in smallscale laboratory investigations as they are simple in operation and installation (Kabdasli et al. 2000; Kim et al. 2006; Zhang et al. 2009; Xu et al. 2012; Foletto et al. 2013). At larger scale, fluidized bed reactors are commonly used as the design gives provision for sufficient reactive surface area and solution turbulence (Seckler et al. 1996). The main advantage of chemical precipitation methods is their operational simplicity; however, production of non-recoverable, fine struvite particles is a common problem in these reactors (Adnan et al. 2003). Moreover, coprecipitation of other salts (such as calcium phosphate) due to the presence of foreign ions is another issue (Capdevielle et al. 2013). Other limitations include low solubility of chemicals like MgO and KOH, energyintensive step CO2 stripping method, loss of ammonia from aeration, and inputs accounting for a large share of the total production cost (Cusick et al. 2014).

Electrochemical Methods This method is based on the principle of electrochemical precipitation, where an electrochemical cell is used with an anode made of platinum, graphite, or carbon-felt disks and a cathode of nickel, a platinum-carbon, or a steel plate. Deposition of struvite

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Table 7 Comparative overview of commercial struvite recovery processes (Modified from Kataki et al. 2016b) Electrochemical deposition Deposition of struvite on a cathode in a solution containing Mg, PO43, and NH4+ through electrochemical reaction pH Chemical additive Self(NaOH, KOH) establishment of alkalinity Mixing Stirring, fluidizing Stirring, fluidizing Reactor Stirred batch, Electrochemical fluidized bed cell 1.1.Limitation 1. Coprecipitation 1. Coprecipitation of impurity as salt of impurity 2. Ineffective 2. Use of costly crystallization for material like Pt not meeting suitable conditions 3. Production of 3. Scale formation fines on cathode

Specifications Principle

Advantage

Chemical precipitation Precipitation of P and NH4+ in solution with addition of Mg and mixing

Easy to install and operate Does not employ the use of sophisticated equipment Installation At commercial/ laboratory scale Demonstration In real (manure, sludge, ash) and synthetic wastes Large share of Alkali source, Mg cost source

No need of alkali addition, concurrent production of potential H2 fuel

Ion exchange P and NH4+ are exchanged in ion exchangers and precipitated as struvite upon Mg addition

Biomineralization Precipitation through biomineralization in medium containing PO43 and Mg, utilizing NH4+ from N metabolism by microbes Chemical additive Self-established (NaOH) alkalinity Not applicable Ion exchange column 1. Coprecipitation of impurities 2. Regeneration of resin at regular intervals

3. Limited availability of specific anion exchangers for PO43 sorption Fast precipitation

Not applicable Batch culture 1. Coprecipitation of impurity 1.2. Slow precipitation

No external addition of alkali

At laboratory At laboratory scale At laboratory scale scale In real (sludge, In synthetic waste In synthetic waste digestate)/ synthetic waste Cathode material, Ion exchange resin Mg source electrical energy

takes place on the cathode from an analyte solution containing Mg, PO43, and NH4+ ions. In this process chemicals are not required for pH adjustment as hydrogen gas is concurrently released during the electrolytic reduction of water at the cathode during the process (Moussa et al. 2006; Wang et al. 2010). In this process, hydrogen recovery for other uses could offset the operational costs of the process (Cusick and Logan 2012).

Enhanced biological phosphorus removal process is based on the ability of P-accumulating microorganisms to uptake P into their cells from a surrounding medium

Microwave digestion of sources under specific temperature and duration solubilizing particulate P

Microwave treatment

Principle Acidification causes protonation of phosphate ions from bound phosphates (Ca/Mg/Fe phosphate), lowers their ionic product below their equilibrium solubility product which results in dissolution of particulate P into solution The ligands react with the calcium [Ca-PO4] particulates to form soluble [Ca-EDTA or Ca-oxalate] complexes and eventually PO43 is released (Zhang et al. (2010))

Enhanced biological P removal

Chelating agent treatment

Acid/base leaching

Table 8 Overview of pretreatments used in struvite recovery

Microwave, oxidants, acids, and bases

P-accumulating microbial strains

Ethylenediaminetetraacetic acid (EDTA) Oxalic acid

Materials used Acid/base (HCl, H2SO4)

Temperature, microbial population Forms of P present, microwave operating temperature, and duration of heating

Waste characteristics

Type of source

Factors affecting Type and composition of material Type of acid/base, pH, temperature

Dairy manure

Sewage sludge

Dairy manure

Used in Dairy, poultry

Process needs optimization, as excess chelate might subsequently bind to Mg limiting Mg availability Concerns regarding environmental toxicity and expense Vulnerability of accumulated P to get released from microbial mass during sludge handling. Less efficient for wastes with higher Ca contents Temperature and time should be effective

Limitation Higher metal contamination

846 S. Kataki and D. C. Baruah

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In recent works, microbial electrochemical cells have been used to help in phosphate mobilization during struvite recovery (Fischer et al. 2011). Where waste source has a high inorganic phosphate (FePO4, Al(PO4)3) content, such as in digested sewage sludge, microbial fuel cells mobilize phosphate by reducing the inorganic phosphate using energy from microbes, liberating the phosphate into solution (Fischer et al. 2011). Retention of unwanted heavy metals that are retained in sludge matrix in immobilized form is an advantage of the process (Fischer et al. 2011). Further, where a high dose of Mg is required, Mg can itself be used as anode (Hug and Udert 2013; Kruk et al. 2014) making the process cost effective (Hug and Udert 2013). The use of precious metals like platinum and cathode deterioration by struvite accumulation are the limitations of the process (Hirooka and Ichihashi 2013; Cusick et al. 2014).

Ion Exchange Methods This method is based on the principle that nutrients from wastewaters are selectively exchanged in ion exchangers and struvite is precipitated after addition of Mg2+ at controlled pH (Liberti et al. 1986, 2001; Mijangos et al. 2004, 2013; Ortueta et al. 2015). The important factors influencing the process are the eluent concentration and selection of ion exchange resin (functional group of resin and matrix of ion exchanger) (Mijangos et al. 2013; Ortueta et al. 2015). Commercial-level application of this method is not very extensive. Availability of specific anion exchangers for phosphate sorption is the main limitation of this process (Petruzzeli et al. 2004). Moreover, the high suspended solid content of regenerated effluent may cause fouling of the exchange columns (Gonder et al. 2006).

Microbial Biomineralization Methods Certain bacterial strains (e.g., Myxococcus xanthus, Staphylococcus aureus) can precipitate struvite in a medium containing PO43 and Mg through the process of biomineralization. NH4+ required for precipitation is produced from microbial metabolism of the nitrogenous compounds present in the medium or precipitating solution (Omar et al. 1998; Gonzalez-Munoz et al. 1996; Omar et al. 1998). Apart from living microbial cells, dead cells, disrupted cells, and isolated bacterial structures (cell membrane) can also induce struvite crystallization by acting as substrates for heterogeneous nucleation for crystallization (Gonzalez-Munoz et al. 1996; Omar et al. 1998). Organic matrix of disrupted bacterial cells has been reported to be rich in negatively charged multimolecular complexes (proteolipids, phospholipids, glycoprotein, proteoglycan) and attracts positive ions like Mg resulting in struvite precipitation (Gonzalez-Munoz et al. 1996; Omar et al. 1998). The effectiveness of microbial production of struvite depends upon the microbial growth phase (Lopez et al. 2007). Parameters related to culture medium (pH, total phosphorus, and total nitrogen) used for microbial growth have direct impact on the amount of struvite produced (Da Silva et al. 2000) (Fig. 15).

848

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S. Kataki and D. C. Baruah

b

Nutrient rich waste source Base Mg2+

Mg2+ Stirrer

PO43–

NH4+

PO43–

NH4+

PO43–

STRUVITE

pH N mineralization

PO43– NH4+ Organic N PO43–

Struvite

Organic N Organic N 3– PO 3– PO4

PO43– B

PO43–

B

B

B

4

B B

PO43–

B

Bacteria

Struvite

c

d

NaCI

CI–

Na+

Nutrient rich waste source e

e Mg

2+

Anion exchanger

Na+

Na+

CI–

CI–

Na+

Na+

CI–

CI–

Na+

Na+

PO43–

2– HPO4 PO43– 3– 2– PO4 HPO4 3– PO4

Mg2+

HPO42–

NaOH

H2O+e

Cation exchanger

ANODE

Resin bed

CI–

–

1/2 H2+OH– pH –

O2+2H2O+4e

–

4OH

CATHODE

e

CI–

pH

PO43–

PO43–

NH + NH4+ 3–4 PO 4

NH4+

Struvite

PO43–

Regenerate Elute Struvite

Fig. 15 Schematic diagrams showing (a) chemical, (b) microbial, (c) electrochemical, and (d) ion exchange method used for struvite precipitation (Adapted from Kataki et al. 2016b)

Use of Different Pretreatments for Maximizing Struvite Recovery In many conventional and prospective sources (farm wastes, municipal wastes and wastewater, and industrial wastes), P remaining in available or recoverable form is often at a minimum because of its fixation with other ions. This makes the recovery and crystallization process challenging. Therefore, such sources may require additional pretreatment to improve the initial phosphate solubility to achieve maximum P recovery, which are discussed below.

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pH Modification Using Acid/Base Leaching Acidification (using hydrochloric acid, sulfuric acid) dissolves P into solution enhancing P availability for struvite formation (Zhang et al. 2010). Pretreatment by acidification has been used in dairy manure (P availability increased by 500%) (Zhang et al. 2010), dairy manure (43–100% of total P released) (Shen et al. 2001), poultry manure (increase of 60–80% of total P) (Szogi et al. 2008), and sewage sludge (release of 80–100% of total P) (Stark 2005). The degree of P leaching is affected by the type and composition of material under treatment, the acid/base added, pH, and temperature (Stark 2005). It has been reported that higher amount of base is required compared to acid to leach same amount of P (Mrowiec et al. 2003). However, acid leaching gives higher metal contamination in waste effluents as acids can leach out other metal ions along with P (Stark 2005).

Chelating Agent Treatment Use of chelating agents like ethylenediaminetetraacetic acid (EDTA) and oxalic acid has been reported as a pretreatment method prior to struvite formation. Chelating agents help to release particulate PO43 by sequestering calcium from it and thereby suppressing the formation of calcium phosphate compounds (Shen et al. 2001). Zhang et al. (2012) reported increase in dissolved P of up to 93% by EDTA treatment in digested dairy manure, and Zhang et al. (2015) achieved removal of 97% of soluble Ca by calcium oxalate precipitation after using oxalic acid in dairy manure. Such treatments are suitable for waste sources with rich Ca such as farm manure; however, the quantity of chelates needs to be optimized as it might limit Mg availability by binding to it (Zhang et al. 2010; Shen et al. 2001). Further, concerns about environmental toxicity remain with some chelates.

Microwave Heating A number of studies reported the use of microwave irradiation pretreatment method for enhanced struvite recovery (Liao et al. 2005; Pan et al. 2006; Chan et al. 2007; Qureshi et al. 2008; Kenge et al. 2009; Lo and Liao 2011; Xiao et al. 2015). The factors affecting the degree of P release are forms of P present, microwave operating temperature, and duration of heating (Liao et al. 2005; Pan et al. 2006). An optimum temperature of 120  C has been suggested as optimum by Chan et al. (2007) in sewage sludge. Uniform heating throughout the material with precise control over process temperature, achievement of high heating rates, and no direct contact between heating source and material under treatment are some of the advantages of microwave treatment over conventional heating (Lo and Liao 2011). It has been reported that the efficiency of the treatment method can be enhanced by addition of chemical-assisted (oxidants, acids, and bases) microwave digestion (Pan et al. 2006; Chan et al. 2007; Qiao et al. 2008; Qureshi et al. 2008).

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Enhanced Biological Phosphorus Removal (EBPR) Enhanced biological phosphorus removal process can accumulate 80–90% of P present in surrounding medium into microbial biomass. This results in up to 12% of P concentration in EBPR sludges, whereas for conventional sludge, it remains 2–3% (Liao et al. 2005). The use of EBPR sludge as struvite source has been demonstrated in previous literature (Britton et al. 2005; Pastor et al. 2008; Shen et al. 2001) reporting 58–94% of P recovery (Munch and Barr 2001; Britton et al. 2005; Marti et al. 2008). However, the recovery efficiency varies depending upon sludge characteristics, temperature, and microbial population (Pastor et al. 2008).

Use of Seed in Struvite Crystallization Seed acts as template by providing surface area and thus reduces induction period for crystal development. Among the reported seed materials in struvite recovery, struvite is the most widely investigated (Table 9). Effect of seeding on struvite was reported to have no significant effect to significant effect. Earlier investigations by Regy et al. (2002) and Rahaman et al. (2008) found no significant effect of seed on struvite crystallization. The studies concluded that the newly formed crystal nuclei provide greater surface area for new crystal development than the seed crystals resulting in no effect under seeded condition (Regy et al. 2002). However, the use of struvite seed was found to enhance recovery by 5% and crystallization rate up to 21% compared to unseeded crystallization in studies by Zhang et al. (2009) and Yu et al. (2013). An isomorphic crystal plane of struvite seed promotes adhesion and integration of growing molecules and clusters of struvite without the need of nucleation, which makes the process energetically favorable resulting in enhanced P recovery efficiency. Further, isomorphic crystal of seed struvite intensifies crystallization minimizing induction time (Liu et al. 2011a). In struvite recovery from fertilizer wastewater, induction time is reduced by 75 min using struvite seed compared to unseeded condition (Liu et al. 2011a). Though, in recent literature, some alternative seeds such as sand, stainless steel, pumice stone, borosilicate glass, etc. are used, studies reported lower crystallization rates and higher induction time under non-struvite seed (Ali 2005). The reason could be changes in the type of nucleation from homogeneous to heterogeneous under non-struvite seed (Ohlinger et al. 1999; Ali 2005). The factors affecting seed’s effectiveness are its surface roughness, size, dosing of the seed, and supersaturation of crystallizing (Ohlinger et al. 1999). Increasing in dose and grain size of struvite seed usually has a positive effect on struvite formation (Huang et al. 2010a). In saponification wastewater, ammonium removal efficiency is increased through formation of struvite from 84% to 91% when struvite seed size is increased from 0.05–0.098 to 0.098–0.150 mm. The efficiency is further enhanced to 92% when the dose is increased to 60 g/l (Huang et al. 2010a). Liu et al. (2011a) showed that, at a seed dosing of 0.422 g l1, the induction time is less by 49 min in a solution with higher supersaturation compared to a solution with lower

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Table 9 Different seed material used in struvite precipitation and their effects on recovery (Modified and expanded from Kataki et al. 2016b) Seed used Struvite

Seed Size (μm) 1000

45–63 75–150 NR

250–500 NR

NR 30–50 NR

Effects on struvite production Production of struvite fine as product Seeding is insignificant (process appeared “selfseeding”) Crystals have similar shape with seed (no phase transformation during growth) Effectiveness of seed requires consideration of pH (pH 9 being optimum) Enhancement of crystallization by 19% at low P concentration Increased crystal size, settle ability No enhancement of P recovery and reduction in induction time Increase in recovery by approximately 5% (at pH 9.5) No effect of overdosing of seed on recovery (pH 9.5) Reduction in induction time up to 75 min depending upon supersaturation Similar shape of struvite with seed

Coarse sand

200–300

Increase in rate of crystallization (by 21%) and size of crystal (from 1.72 nm to 2.08 nm) No fixation of struvite on sand surface

Fine sand

150–200

Strong primary nucleation and formation of fine

Borosilicate glass Sand grain/ quartz particle

45–63

Slower reaction rate compared to struvite seed

210–350

Recovery of 80% of P onto seed bed

Phosphate rock Stainless steel mesh

45–63 NR

Slower reaction rate compared to struvite seed No effect mentioned on crystal

1000 um hole

Pumice stone

NR

No significant increase in crystallization Reduction in struvite fine particle No effect of seed dosing on recovery Coprecipitation of Ca and silica on seed

Reference Regy et al. (2002) Ali (2005) Kim et al. (2006 ) Liu et al. (2008a) Rahaman et al. (2008) Zhang et al. (2009)

Liu et al. (2011a) Mehta et al. (2013) Yu et al. (2013) Regy et al. (2002) Regy et al. (2002) Ali (2005) Battistoni et al. (2000) Ali (2005) Massey et al. (2007) Le Corre et al. (2007) Pakdil and Filibeli (2008)

supersaturation. However, struvite crystallization with seeding material is an energyand cost-intensive process as amounts of seed requirement are often high to make crystallization effective. Although tested in synthetic liquor, feasibility of the use of most of the novel seeds is yet to be tested with real waste.

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Alternative Magnesium Sources for Struvite Precipitation Struvite recovery needs input of chemical and indispensible consumption of Mg source make the precipitation often expensive (Chimenos et al. 2003; Lee et al. 2003). In majority of recovery sources, addition of Mg is routinely required due to their low inherent content. The most used Mg sources are salts of Mg, such as MgCl2, MgSO4, MgO, and MgOH. It has been reported that high cost of high-grade Mg compounds has limited its full-scale implementation, contributing up to 75% of overall production costs (Dockhorn 2009). Some studies investigate the effectiveness of some alternative and recyclable Mg sources to minimize the process cost (Table 10). Mainly the aspects of Mg availability, reagent solubility, and reactivity of the Mg source play a significant role on the feasibility of its use. Bittern and seawater are two novel Mg sources reported in struvite recovery. Struvite recovery efficiency in coke manufacturing wastewater using seawater and bittern is reported to be 95% and 99% of total P, respectively (Shin and Lee 1997), which is comparable with recovery by conventional Mg salts. However, because of the presence of insoluble Mg in bittern or seawater, process needs high Mg dose (>1.5:1) for efficient P recovery (Matsumiya et al. 2000; Kumashiro et al. 2001; Quintana et al. 2004). However, Ca ion prevalent in seawater could interfere in recovery process. Seawater and bittern have been shown effective for struvite precipitation in swine wastewater, coke wastewater, urine, landfill leachate, and municipal wastewater. Magnesite (MgCO3), a by-product of MgO production (Chimenos et al. 2003; Quintana et al. 2004), is another prospective Mg source. However, magnesite has low solubility in water, requiring high dose for struvite precipitation. Acid dissolution and thermally decomposed magnesite can be more effective as Mg source (Gunay et al. 2008; Huang et al. 2010b). Thermal decomposition or pyrolysis product of struvite, viz., MgHPO4 and Mg2P2O7, can also be used as recycled Mg source (Zhang et al. 2009; Huang et al. 2009; Yu et al. 2012). Process is further modified by using struvite pyrolysis under alkali condition (Türker and Çelen 2007; He et al. 2007; Zhang et al. 2009; Yu et al. 2012), which produces MgNaPO4 as per reaction 1 (Huang et al. 2011b). It is estimated that the use of struvite pyrolysate can save up to 44–48% of process cost (He et al. 2007; Huang et al. 2009). Struvite recovery by struvite pyrolysate in landfill leachate (maximum 96% NH4+ removal) was achieved under optimum conditions of OH/NH4+ of 1:1, temperature of 90  C, and time of 2 h as reported by He et al. (2007); however, the conditions vary with recovery source. It has been reported that, in the subsequent recycling cycles of struvite pyrolysate, the NH4+ removal ratio decreases (He et al. 2007; Huang et al. 2011b) because of accumulation of inactive Mg2P2O7 and Mg3(PO4)2 (Sugiyamaa et al. 2005; Yu et al. 2012), which can be improved by acidolysis of pyrolysate (Zhang et al. 2004; Yu et al. 2012).

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Prospects and Issues of Phosphorus Recovery as Struvite from Waste Streams

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Table 10 Alternative Mg sources used in struvite recovery (Adapted and expanded from Kataki et al. 2016b) Mg source Bittern

Mg content 31,390 mg l1 9220–24,900 mg l1 32,000 mg l1 27,500

Seawater

1136 mg l1 1250 mg l1 1250 mg l1 1200 mg l1 676.7 g kg1

Thermally decomposed magnesite Brucite

650 g kg1

Magnesite

300 g kg1 940 g kg1 244 g kg1

Struvite pyrolysate 530 g kg1 Wood ash

34 g kg1

Synthetic nanofiltration of brine from seawater Desalinated reject water

146 mmol l1

Mg(II) solution from seawater Calcined dolomite (MgO/CaCO3 suspension)

1555–2795 mg 1 8000 mg 1 36%

Key findings Similar recovery efficiency as MgCl2

Reference Shin and Lee (1997) Comparable struvite precipitation Li and Zhao efficiency with MgCl2 and MgSO4 (2002) Bittern is more effective in P recovery Lee et al. (2003) than NH4+ recovery More cost effective in coastal areas Etter et al. (2011) Same P recovery efficiency as MgCl2 Shin and Lee (1997) Higher Mg/PO43 (>1.5:1) necessary Matsumiya for more than 70% P recovery et al. (2000) Higher Mg/PO43 (>1.5:1) necessary Kumashiro for stabilized and easy P recovery et al. (2001) Similar recovery efficiency as MgCl2 Lee et al. (75%) (2003) Need higher Mg/PO43 molar ratio for Quintana effective recovery et al. (2004) Brucite can be used as liquid and solid Mg source; reuse of brucite is possible Acid dissolution of magnesite increase struvite recovery by 50% Thermal decomposed magnesite is cost effective than acid-dissolved magnesite Cheaper than bittern and MgSO4 Similar effects on recovery as with Mg salt Presence of impurity such as calcite, heavy metal in product Effective as Mg source but pH and organic matter influence purity of product Presence of other ions (Ca, Na) in reject water reduces recovery efficiency Higher dose of Mg/PO43 (>1:1) has no effect on recovery 89.7% of NH4+ recovery The competition between K and NH4+also led to formation of K-struvite (MgKPO4.6H2O) decreasing NH4+ removal rate

Huang et al. (2011a) Gunay et al. (2008) Huang et al. (2010) Etter et al. (2011) Huang et al. (2011b) Sakthivel et al. (2011) Zewuhn et al. (2012)

Fattah and Ahmed (2013) Lahav et al. (2013) Chen et al. (2017)

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Application of Struvite as Soil Fertilizer Fertilizer Properties of Struvite As fertilizer, the N, P, K, and Mg content of struvite are 5.7:29:0:16.4, respectively (Westerman 2009). The P content of struvite remains in the range of 11–26% (Johnston and Richards 2003) depending upon source and method of production, of which 1–2% is water soluble and rest is citrate soluble (Bridger et al. 1962). It is used by fertilizer companies as additive or as a substitute raw material in standard fertilizer production technology (Li and Zhao 2002; Rafie et al. 2013). The cost of such commercial product largely depends upon processing (drying, storage, creation of a blended product) and transportation (Westerman 2009). The most desirable properties of struvite as fertilizer is its slow nutrient release characteristics. Struvite is sparingly soluble in water with a solubility of 0.02 g 100 ml1 of water at 0  C, rendering its slow assimilation into soil solution (Li and Zhao 2002; Negrea et al. 2010). The N leaching rate of struvite is threefold lower compared to commercial N fertilizer (Rahman et al. 2011). Slow nutrient release from struvite allows its high rate of application without damaging plant roots (Li and Zhao 2002; Rafie et al. 2013). Due to its prolonged release of nutrients throughout the growing season, often a single application is sufficient to meet crop nutrient demand. However, literature indicated two concerns in this context: (i) sometimes the limited availability of N because of low N/P2O5 ratio of struvite makes N insufficient for optimal plant growth (Miso 2009; Gell et al. 2011), and (ii) when the application dose is increased to fulfill N requirement, it is shown to increase soil pH compared to other P fertilizers, which might affect nutrient availability and uptake (Rahman et al. 2011). The fertilizing effect of struvite varies with soil type due to differences in solubility and sorption properties of the soils. Struvite is most effective in soil of moderate or low pH, but its effectiveness is limited in soils with marginal fertility and high pH. Struvite solubility is minimum (0.040 mM) at pH 8.2–8.8 (Le Corre et al. 2009), which can rise up to 1–10 mM at pH below 5 (Borgerding 1972; Abbona et al. 1982). Because of similar solubility of struvite with triple superphosphate, struvite was found to have similar effect as in acid and neutral soil (Cabeza et al. 2011). However, in alkaline calcareous soil also, struvite is found to be more soluble, making it a recommended P fertilizer (Massey et al. 2007). It has been reported that, compared to chemical fertilizer-treated soil, leaching of N from struvite-treated soil is lower, but there is no significant variation in P leaching (Rahaman et al. 2011). Quality standard of struvite can be described in terms of its crushing strength, composition, and purity which is influenced by source, processing (drying), and Mg addition (Antonini et al. 2012). Multicomponent raw material of struvite like sewage sludge usually contains a number of heavy metals (Cd, Cu, Cr, Ni, Pb, and Zn) and organic micro-pollutants. However, there is a lack of information on residual impurity in recovered struvite from various sources (Wollmann and Moller 2015).

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Prospects and Issues of Phosphorus Recovery as Struvite from Waste Streams

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Previous literature reports impurity content of struvite from municipal wastewater to comply with fertilizer regulatory limit in different countries like Germany and Turkey (Uysal et al. 2010; Antonini et al. 2012; Forrest et al. 2008; Latifian et al. 2012). Crushing strength is a quality parameter for struvite as fertilizer which predicts its handling and storage properties. The average size of commercial struvite crystals generated is 2–3.3 mm in size depending upon reactor conditions, viz., upflow velocity, pH, and supersaturation ratio (Forrest et al. 2008).

Effect of Struvite Application on Crop Recently a reasonable number of studies have come up to establish the fertilizer value of struvite on a range of crops, considering its properties, composition, effect on soil and plant, and comparative performance with chemical fertilizer. These studies reported the use of struvite on diverse crops including vegetables, ornamental, forest outplanting, turf, orchards, and potted plants. It is evident from previous studies that there is no significant difference between P in struvite and water-soluble P in other phosphate fertilizers, as most of the studies reported comparable effect of struvite as fertilizer with chemical fertilizer (Ghosh et al. 1996; Johnston and Richards. 2003; Li and Zhao. 2003; Plaza et al. 2007; Massey et al. 2009; Perez et al. 2009; Liu et al. 2011b; Gell et al. 2011; Dalecha et al. 2012). Urine-derived struvite was found to produce yield and phosphate uptake comparable to those induced by phosphate fertilizer when tested with ryegrass, Zea mays L., and red clover (Simons 2008; Antonini et al. 2012). Instead, in some literatures, struvite has been reported to yield better results in comparison with some conventional fertilizers such as ammonium phosphate, diammonium phosphate, and single superphosphate (Barak and Stafford 2006; Gonzalez-Ponce et al. 2009; Yetilmezsoy et al. 2013). When compared with single superphosphate, urine derived struvite is found to be more effective in lettuce yield with enhanced P uptake, which was probably related to its higher Mg content and the synergistic effect of Mg on P uptake (Gonzalez-Ponce et al. 2009). It has also been reported that P concentration is higher in plants treated with struvite than in plants treated with other P fertilizers (Li and Zhao 2003; Gonzalez-Ponce and Garcia-Lopez 2007). While comparing the effect of struvite with phosphate rock, monoammonium phosphate, and calcium superphosphate on potted ryegrass, P accumulation is found to be highest in struvite-treated perennial ryegrass (Gonzalez-Ponce and Garcia-Lopez 2007). Three-year field experiments with MAP showed no statistically significant benefit in terms of increased grain yield over unfertilized soil or phosphate rock (Weinfurtner et al. 2009), and pot experiments suggest struvite has a relative fertilizer efficiency of 64–134% of TSP (Perez et al. 2009). However, there are some studies which reported lower yield in struvite-treated plants because of lower availability of nutrients compared to chemical fertilizer (Ganrot et al. 2007; Ackerman et al. 2013). Therefore, supplementation of chemical fertilizer along with struvite has been recommended for

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Table 11 Effect of application of struvite as fertilizer on various crops (Adapted and modified from Kataki et al. 2016b) Crop/plant Gram (Cicer arietinum L.)

Struvite source Synthetic water

Chinese flowering cabbage (Brassica parachinensis) Water convolvulus (Ipomea aquatica, I. reptans)

Municipal landfill leachate

Water spinach (Ipomoea aquatica) Chinese chard (Brassica rapa var. chinensis) Maize

Municipal landfill leachate

Municipal landfill leachate

Municipal landfill leachate NR

Swine wastewater Urine

Black water Urine Corn fiber processing wastewater

Wheat (Triticum aestivum L.)

Urine Dairy wastewater

Remark Variation of P uptake with level of P application Superior or equally effective as chemical fertilizer Similar vegetable growth and more Mg and P uptake compared to chemical fertilizer No significant difference in growth and no burning effect with increase in struvite dose Higher dose of struvite does not affect plant Similar vegetable growth and more Mg and P uptake compared to chemical fertilizer Similar vegetable growth and more Mg and P uptake compared to chemical fertilizer P uptake efficiency for struvite is 117%, and residual P availability is 178% Higher efficiency compared to chemical fertilizer Similar plant height, higher biomass, and less N2O emission as with chemical fertilizer No significant difference in dry yield compared to chemical fertilizer No significant difference in dry yield compared chemical fertilizer Similar leaf diameter and height with chemical fertilizer Plant P uptake is higher than chemical fertilizer by 4–21% depending upon application rate Nutrient availability is similar to chemical fertilizer Lower dry weight (by 50%) than chemical fertilizer Low availability of nutrients (N) than chemical fertilizer Increase in total P uptake in basic soil

Reference Ghosh et al. (1996)

Li and Zhao (2003) Li and Zhao (2003)

Li and Zhao (2003) Li and Zhao (2003) Barak and Stafford (2006)

Liu et al. (2011b) Gell et al. (2011)

Gell et al. (2011) Dalecha et al. (2012) Thompson (2013)

Ganrot et al. (2007) Massey et al. (2009)

(continued)

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Table 11 (continued) Crop/plant Perennial ryegrass (Lolium perenne)

Struvite source Municipal sewage sludge Synthetic liquor and municipal sewage sludge Poultry manure

White lupin (Lupinus albus L.)

Wastewater

Oilseed rape

Municipal sewage sludge

Purslane (Portulaca oleracea)

Poultry manure

Garden cress (Lepidium sativum) Winter barley

Poultry manure

Lettuce (Lactuca sativa L.)

Garden rocket (Eruca sativa) Dill (Anethum graveolens) Fennel (Foeniculum vulgare) Parsley (Petroselinum crispum) Canola

Municipal sewage sludge Anaerobically digested municipal sludge liquor Anaerobic sludge of poultry manure Anaerobic sludge of poultry manure Anaerobic sludge of poultry manure Anaerobic sludge of poultry manure Swine manure

Remark Similar increase in dry matter and P uptake compared to chemical fertilizer Similar dry matter yield and P uptake as with chemical fertilizer

Reference Plaza et al. (2007)

Increase in fresh and dry weight by 76% compared to 60% in control, faster growth than with control Increase in weight and rate of increase is dependent upon type of plant and soil media Equal P uptake and higher Mg uptake compared to P fertilizer

Yetilmezsoy and Zengin (2009)

Higher P uptake and grain yield compared to synthetic fertilizer/ rock phosphate Increase in fresh and dry weight by 150% compared to 207% in control, faster growth than with control dependent upon type of plant, soil media Increase in fresh and dry weight by 28% compared to 115% in control, faster growth than with control Similar P uptake and grain yield compared to chemical fertilizer/ rock phosphate More efficient than chemical fertilizer in increasing yield and P uptake

Johnston and Richards (2003)

Gonzalez-Ponce and GarcıaLopez-de-Sa (2008) Perez et al. (2009) Yetilmezsoy and Zengin (2009)

Yetilmezsoy and Zengin (2009) Perez et al. (2009) Gonzalez-Ponce et al. 2009

More gain in plant wet, dry weight, and height compared to chemical fertilizer Increase in dry weight by 191% compared to no fertilizer

Yetilmezsoy et al. (2013)

Increase in dry weight by 208% compared to no fertilizer

Yetilmezsoy et al. (2013)

Increase in dry weight by 379% compared to no fertilizer

Yetilmezsoy et al. (2013)

Similar P uptake but lower biomass yield/unit of P uptake

Ackerman et al. (2013)

Yetilmezsoy et al. (2013)

(continued)

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Table 11 (continued) Crop/plant

Tomato

Struvite source

Yeast industry wastewater

Remark compared to chemical fertilizer because of lower solubility at basic soil Supplement with chemical fertilizer might need depending upon soil type Higher N, P, and Mg uptake of plant for double/triple and quadruple dose of struvite compared to NPK treatment Higher (more than double) dose of struvite is essential for optimum effect

Reference

Uysal et al. (2014)

better results (Ackerman et al. 2013). Further, chemical P fertilizer-treated crop is shown to have better yield compared to struvite, which has been attributed to a potential potassium deficiency as reflected in crops (Hammond and White 2005). Its application enhances P uptake, as the Mg present in it has a synergistic effect on P absorption (Gonzalez-Ponce et al. 2009). From the above discussion, it is seen in previous research reports that there is variation in the fertilizing effect of struvite with no significant impact on plants but with significant effect on P and Mg uptake and biomass yield. The findings are subjected to various factors such as soil type, plant type, and climate behavior, keeping further scopes on its application at larger scale corresponding to these factors. However, considering the associated benefits of struvite recovery process, struvite recovery appears to be an attractive and feasible alternative (Table 11).

Conclusions Here we have analyzed and reviewed outcomes of different studies addressing suitability and prospects of waste streams of various origins for struvite production, methods of recovery, commercial struvite recovery, strategies for enhanced and effective recovery (pretreatments, alternative Mg source, seed), and their application prospects for comprehensive understanding of evolution in this field. Struvite recovery, as a method of nutrient recycling, has expansive multifaceted benefits and is receiving support from R&D. Potential of struvite recovery from farm wastes, municipal and industrial wastes remain highly attractive with a high P recovery efficiencies (~85 to 95%). In recent years, development of technoeconomically efficient method from alternative potential struvite sources with incorporation of more efficient method, Mg source, and seed material remains the mainstream research focus. However, while a range of laboratory-scale struvite

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recovery experiments were found to be successful, development of cost-effective, targeted, and environmentally friendly full-scale installations is still a challenge due to inherently heterogeneous nature of the sources and unfavorable economics, which requires further research and development. On the other hand, study on its application aspect identifies its favorable impact, though more intense research on field trial is required with region-specific focus for long-term effects and with a wide range of test crops. Further, strategy for struvite market development should focus on a holistic approach considering pricing, purity, quantities, size, storage, transportation, distribution, and infrastructure, in view of the legal framework of contaminants, eco-toxicity, and hygiene, so that value-added products can be developed to recycle P into the nutrient cycle and can be used as a supplement to prevailing nutrient supply system. The developmental impacts of such technological successes would be profound for global food security in terms of alternative and sustainable fertilizers for crop.

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Global Status of Nitrate Contamination in Groundwater: Its Occurrence, Health Impacts, and Mitigation Measures

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Saurabh Shukla and Abhishek Saxena

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Occurrence of Nitrate in Various Ecosystems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Anthropogenic Sources . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Geogenic Sources . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Status of Nitrate Contamination in Groundwater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Global Status . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Status in India . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Health Impacts . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Methemoglobinemia . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cancer . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Thyroid Problems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Impacts on Livestock Health and Aquatic Life . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Removal Technologies of Nitrate from Drinking Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Removal Through Carbon-Based Adsorbents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Removal Through Natural Adsorbents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Removal Through Sorbents Obtained from Agricultural Waste . . . . . . . . . . . . . . . . . . . . . . . . . . . . Removal Through Sorbents Obtained from Industrial Waste . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

870 871 873 875 876 876 878 879 880 881 881 881 882 882 883 883 883 884 885

Abstract

Nitrate has emerged as one of the most alarming and widespread contaminant of groundwater and surface water resources reported around the globe. Nitrate formation is an integral part of nitrogen cycle and is added either by the natural processes (atmospheric fixation, lightning storms) or through anthropogenic activities (fertilizer applications, septic tanks). Nitrate enters the hydrosphere

S. Shukla (*) · A. Saxena Faculty of Civil Engineering, Shri Ramswaroop Memorial University, Lucknow, India e-mail: [email protected] © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_20

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easily, and its ingestion causes various health risks such as methemoglobinemia, cancer, diabetes, etc. on humans and to some extent on livestock populations as well. Agricultural practices and subsequent fertilizer application along with other anthropogenic activities are assumed to be the primary reason behind elevated levels of nitrate in groundwater. However, even under the similar ecological conditions, the reported occurrence of nitrate in groundwater is sporadic in nature, indicating the possible interference from other complex factors (including geogenic factors) and a dynamic release mechanism. Various concepts for the different sources and occurrence of nitrate in groundwater along with its health impacts are discussed in this chapter. Several existing and upcoming technologies are there to remove excess nitrate from potable water; these are also discussed in the chapter. Keywords

Nitrate contamination · Sporadic occurrence · Geogenic sources · Methemoglobinemia · Groundwater pollution · Indo-Gangetic alluvium · Anthropogenic sources · BATs · Removal technologies · Adsorption · Activated carbon · Industrial waste · Agricultural waste

Introduction Groundwater, which serves as the only drinking source in many parts of the world (Zekster and Everett 2004), is one of the most precious natural resources. It is also the world’s most extracted raw material with withdrawal rates in the range of 982 km3/year (Margat and Gun 2013). The total volume of fresh groundwater stored on earth is believed to be in the region of 8–10 million km3, which is more than 2000 times the current annual withdrawal of surface water and groundwater combined (Gun 2012). Generally, fluctuating water levels in shallow groundwater, within a relatively short time scale, can be used as an indicator of land use stresses that may affect deep aquifer systems as well, with passage of time. The groundwater quality deterioration can be attributed broadly to two mechanisms, (i) anthropogenic and (ii) geogenic (Lapworth et al. 2017), and during the last two decades, nitrate (NO3) contamination in groundwater has become highly discussed worldwide, and it is the second most common threat after pesticides (Spalding and Exner 1993). Nitrate, nitrite (NO2), ammonia (NH3), and organically bound forms of nitrogen (Org-N) are the species of concern for water resource management (Anayah and Almasri 2009; CGWB 2014) around the globe. The background nitrate concentration in groundwater shall not exceed 10 mg/L as nitrate or 2 mg/L as nitrate-nitrogen, NO3-N (Mueller and Helsel 1996). The levels exceeding this background limit are an indicator of nitrate contamination, primarily through various anthropogenic activities (Wongsanit et al. 2015). Although this background concentration cannot be considered of pristine state, the contribution from geogenic sources of nitrate cannot be ignored in comprehensive analysis.

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Table 1 Permissible limit of nitrate in drinking water by different agencies Country/organization WHO United States Environmental Protection Agency BIS

Concentration as NO3 - N (mg/L) 11

Concentration as NO3 (mg/L) 50

10

45

–

45

Source WHO (2011) EPA (2012) IS: 10500 (2012)

The guideline value for nitrate in drinking water, prescribed by various agencies, is summarized in Table 1 to protect against methemoglobinemia in bottle-fed infants (WHO 2011). The WHO guideline value of 50 mg/L is based on epidemiological evidence for methemoglobinemia in infants, which results from short-term exposure and is protective for bottle-fed infants and, consequently, other population groups (WHO 2011). Various studies have been conducted around the world, and the presence of geogenic factors along with anthropogenic ones makes nitrate contamination of groundwater an intricate challenge. The aim of this chapter is to present a review on the nature of occurrence of nitrate in groundwater, global status, its dynamic release mechanism, various health impacts, and the existing and upcoming removal technologies.

Occurrence of Nitrate in Various Ecosystems Formation of nitrates is an indispensable part of the environmental nitrogen cycle. Globally, approximately 260 million tons of atmospheric nitrogen is being fixed. Biological fixation accounts for 67.25% of nitrogen fixed per year (through land (legume and nonlegume) and sea (Table 2). The remaining 32.75% is fixed through non-biological processes such as atmospheric lightning and industrial discharges (Miyamoto et al. 2008). This nitrogen fixed in the soil undergoes the process of ammonification (Fig. 1) and gets converted to ammonium, which is further converted first to nitrite (undergoing the process of nitrification) and then to nitrate (through the action of nitrifying bacteria). Nitrate are also formed when aerobic and anaerobic bacteria along with other microorganisms present in soil break down decaying dead plants and animals, fertilizers, and other organic residues and convert these to ammonium ion, which is subsequently converted to nitrate (Fig. 1). This converted nitrate undergoes assimilation and is utilized by plants to satisfy nutrient requirements and may accumulate nitrate in their leaves and stems. Sometimes rain or irrigation water can leach them into groundwater as nitrate-containing compounds are generally soluble and readily migrate into groundwater (Self and Waskom 2014).

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Table 2 Estimated amounts of nitrogen fixed annually from different sources (Source: Miyamoto et al. 2008)

Source of N fixation Land Legume Nonlegume Others Sea Total biological Lightning Industry Total non-biological

Sedimentary Organic Matter

Nitrogen fixed (106 tons per year) 153 39 10 104 40 193 9 85 94

N2

Mineralization

Organic N

N2O

Denitrification

tion

Fixa

Assimilation NH4+

NH4+ mineral

NO2− Nitrification

NO3−

NO3− mineral

Fig. 1 Geologic nitrogen cycle (Adapted from Holloway and Dahlgren 2002)

Some of the major factors controlling nitrate concentrations in saturated zone groundwater can be characterized using land-use type, chalk formation, borehole depth, effective precipitation, and groundwater level (Roy et al. 2007). Various ecosystems have the nitrogen accumulation and transfer potential (Fig. 2). The accumulation potential is the ability to retain the nitrogen within the respective ecosystem, and transfer potential is the ability of the respective ecosystem to allow the migration of nitrogen into another ecosystem. The nitrogen accumulation and transfer potential varies from one ecosystem to another and can be used as an invariably good indicator of risk of nitrite and nitrate contamination in these ecosystems. The increasing nitrate concentration has adverse effects on various ecosystems and may cause algal bloom and subsequent eutrophication, increase in NOx emissions (Galloway et al. 2003; Zhou et al. 2011), etc. The comparison of this accumulation and transfer potential for various ecosystems is presented in Fig. 2. It may be seen here that, although atmosphere and agro-systems have low accumulation potential, the high transfer potential presents a risk of contamination

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Fig. 2 Nitrogen accumulation and transfer potential of different ecosystems (After Galloway et al. 2003)

in other ecosystems especially transferring into the groundwater. Similarly, high accumulation and high transfer potential (grasslands and forests) present the risk of nitrate contamination in groundwater through anthropogenic interventions. Similar to other pollutants, nitrate also reaches the groundwater through anthropogenic or geogenic sources. Although it is perceived that the contribution toward elevated levels of nitrate in groundwater is primarily caused through anthropogenic sources, which is true up to some extent, still the contribution from geogenic sources also plays an important role in the nitrate contamination of groundwater resources.

Anthropogenic Sources A larger fraction of groundwater nitrates is derived from various anthropogenic sources, which are discussed below.

Fertilizer Application Nitrogen is one of the major components of all the fertilizers, and with the advent of various inorganic fertilizers, usage has increased to escalate the crop yield. Overapplication and improper timing of applying the fertilizers cause the nitrates to leach into the groundwater (Vinod et al. 2015). Fertilizer application and subsequent leaching is reported to have the highest contribution toward nitrate leaching into the groundwater (Fig. 3).

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Fig. 3 Anthropogenic sources of nitrate contamination in groundwater

13% 20%

7% 60%

Cropland

Deforestation

Industrial Waste

Domestic Wastewater & Septic Tanks

Human and Animal Waste In areas of large human and animal populations, wastes generated by them are the major source of nitrate pollution. The majority of urban population is provided with sewerage systems; still many developing countries are using septic tanks/leaching pits as decentralized sewage treatment. These in situ treatment systems have customary leakage problems and contribute toward nitrate leaching into the groundwater (Vijay et al. 2011). Open defecation and other such unhygienic sanitation practices are also responsible for nitrate contamination in groundwater (Tambekar and Neware 2012). Generally, our concern toward the animal waste is limited to problems of odor, flies, and impacts to surface water and surrounding areas; however, waste from dairies, open feedlots, and other such facilities are a potential source of nitrogen and other inputs to groundwater (CGWB 2014). Fecal waste from cattle population and livestock farming also contributes toward nitrate pollution in groundwater (Sahoo et al. 2016). Industrial Uses of Nitrates Several industries such as dye manufacturing units and metal processing industries use various nitrogen-containing compounds such as (i) anhydrous ammonia, (ii) nitric acid, (iii) urea, etc. The nitrogen in these compounds is converted to nitrite and nitrate. Improper disposal of waste material generated from these industries have high probability of adding the nitrate concentration in the groundwater (Akwensioge 2012). Explosives Manufacturing of explosives involves use of ammonium nitrate and diesel as fuel. Also, nitrogen is one of the chief constituent of explosives. The adversely affected disposal streams having high concentration of ammonium nitrate and other converted forms of nitrogen also add to high nitrate level in groundwater (Resende et al. 2014).

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Deforestation The natural tendency of forest is to conserve and accumulate nitrogen in the soil pool (Fig. 2); however, the transfer potential of forests is also high. The nitrogen is easily converted to nitrate, and various anthropogenic interventions such as deforestation lead to leaching of nitrate. This nitrate easily migrates into the groundwater because of high nitrogen transfer potential of forests. Feichtinger et al. (2002) found that surface runoff from forests have nitrate in the range of 0.1–20 mg/L. However, in certain Sierra forests (USA), Miller et al. (2005) reported that the runoff immediately after rain had 150–400 mg/L of nitrate for very brief period of time. The high nitrate concentration in these runoff water has high chance of contributing elevated nitrate levels in groundwater. Among the anthropogenic sources, leaching from croplands and other agricultural practices constitute the major contribution toward high nitrate concentration in groundwater, followed by leaching from septic tanks and contribution from domestic wastewater (Azizullah et al. 2011). The contribution from other factors such as leaching from industrial waste and deforestation is marginally lower (Dahan et al. 2014). The approximate percentage contribution of all these anthropogenic factors is illustrated in Fig. 3.

Geogenic Sources The global mean nitrogen use efficiency by plants is approximately 50%; the remaining nitrogen is lost into the environment (Rao and Puttanna 2006). The extent of denitrification controls the leaching of nitrate into the groundwater through natural processes (Stüeken et al. 2016).

Lightning Storms Atmospheric nitrogen is first converted to ammonia and subsequently to nitrate and is deposited to the soil generally through rainfall and other modes of precipitation, during lightning storms (Galloway et al. 2004). Geological Origin Although often neglected in the past, geologic sources have been reported in recent studies to comprise a large pool of nitrogen. There are several minerals like tobelite, niter, and nitratine, etc., which have nitrogen in their lattice; this nitrogen is converted to nitrate when it is released from the lattice during the process of weathering (Holloway and Dahlgren 2002). The nitrate formed during weathering process may leach down the aquifer strata and contaminate the groundwater (Holloway and Smith 2005). The geological nitrogen cycle is depicted in Fig. 1. In addition, nitrate from mineral undergoes either assimilation (by biota) or reduction to nitrogen gases (N2 or N2O) through denitrification (Holloway and Dahlgren 2002). A study by Gupta et al. (2015) found high nitrate concentrations in the “river bank filtration well” as well as in the groundwater samples in several zones of Srinagar and Srikot area, where water in the adversely affected sites was exposed to zones of bedrock. The isotopic composition of nitrate in water samples and leachates obtained from these rocks (containing phyllites and quartzites) was found similar, confirming that weathering and subsequent leaching of nitrate from these bedrocks have been

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Anthropogenic

Dynamics of Nitrate Pollution

Geogenic

Fig. 4 Sources of nitrate in groundwater

the reason for elevated concentrations nitrate in water samples. Moreover, Power et al. (1974) found that approximately 200 kg/ha of ammonium existed in a 1 m thickness of the Paleocene Fort Union Shale in North Dakota and eastern Montana, USA, which was found to be rapidly nitrified to nitrate under favorable moisture, temperature, and oxygen concentration.

Other Geogenic Factors There are several other factors, which contributes to the soil nitrates and subsequent leaching into the groundwater; however, these might be very small in terms of concentration. Precipitation brings down the nitrates and ammonium forms present in the atmosphere, which are generally formed during combustion activities, volatilization of fertilizers or animal wastes, etc. (Akwensioge 2012). In addition, factors like nature and thickness of surface deposits, rainfall quantity and distribution, depth to the groundwater level, distribution of vegetation types, and presence of nitrogen-fixing vegetation (Lorenzen et al. 2012) also contribute toward the nitrate contamination in groundwater. It is very well established through the reported studies that contribution of geogenic sources of nitrate in groundwater is smaller when compared to anthropogenic sources (Fig. 4). Although anthropogenic activities alone cannot be exclusively attributed toward the high nitrate concentration in soil, challenging analytical methods and varying forms (organic/inorganic) make it difficult for quantification of nitrate in rocks and groundwater (Holloway and Dahlgren 2002). This might be one of the reasons behind the fewer number of studies investigating the geogenic origins behind elevated levels of nitrate in groundwater.

Status of Nitrate Contamination in Groundwater Global Status Nitrate concentration above permissible limits in drinking water sources is a common global phenomenon (Gong et al. 2013; Jiang et al. 2013). The association between nitrate-contaminated well water and blue baby syndrome was first

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described by Hunter Comly, who treated two infants for symptoms of cyanosis during the early 1940s, in Iowa City, USA (Comly 1945). Both infants became ill as they were fed with formulas that were diluted with water from shallow wells. Nitrate concentrations in the related wells were 90 and 150 mg/L. Globally, mean nitrate levels have risen by an estimated 36% in global waterways since 1990 with the highest increase seen in the Eastern Mediterranean and Africa, where nitrate contamination has more than doubled. Several developed and developing countries have been reported with elevated levels of nitrate in groundwater. The developed countries have lower population density and hence different land use patterns to those of developing countries; still similar trends of nitrate contamination in groundwater resources have been reported around the world. Thorburn et al. (2003)) reported that 3% of the sampled wells were having nitrate above 50 mg/L in northeastern Australia. A total of 5101 wells were sampled from 1991 to 2003 in 51 study areas throughout the USA, and more than 4% of the sampled wells were having nitrate levels higher than EPA limit of 10 mg/L as nitratenitrogen (Burrow et al. 2010). Shomar et al. (2008) found that more than 90% of the groundwater wells sampled between 2001 and 2007 showed nitrate concentrations up to eight times higher than the WHO permissible limit in Gaza strip. Similarly, most of the local and regional aquifers in Europe are about to reach the guideline value (50 mg/L) of nitrate as reported by Otero et al. (2009). A study in Afghanistan (Houben et al. 2009) revealed that approximately 42% of the shallow groundwater samples collected in Kabul had nitrate concentration more than 50 mg/L. Tagma et al. (2009) found that more than 36% of the sampled wells in Chtouka-Massa basin, South Morocco, exceed the value of 50 mg/L, and suggested the need of further research for identification of the source. Cerdeira et al. (2008) reported higher nitrate concentration (above MCL of 6 mg/L, given by São Paulo State law in Brazil) in several urban downtown areas of Brazil. Fianko et al. (2009) reported that more than 50% of the borehole samples had elevated levels of nitrate as nitrogen (NO3-N). A total of 148 groundwater samples were collected from three different zones in Osona, Spain (declared “vulnerable to nitrate pollution” by Catalon government), where nitrate concentration was found to be in the range of 10–529 mg/L with a median value of 127 mg/L (Otero et al. 2009). Moore and Matalon (2011) found that San Joaquin Valley is the epicenter of the nitrate contamination and more than 75% of the samples collected from surface and groundwater sources located in the valley were having nitrate concentration above WHO permissible limit, impacting over 275,000 people. Several groundwater samples collected from Gimpo agricultural area in South Korea had the average nitrate concentration of 79.4 mg/L, exceeding the Korean guideline value of 44.3 mg/L of nitrate (Cheong et al. 2012). In some parts of Serbia, it is reported by Devic et al. (2014) that groundwater pollution is caused by varying levels of nitrate, causing the water quality falling to category III/IV which represents contaminated water, unsuitable for human ingestion. Four alluvial floodplains in Europe have been reported with high levels of nitrate by Antiguedad et al. (2016). Janjevic (2017) reported that water from more than 28% of the 700 wells sampled in Germany had nitrate “exceeding

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the legal limit of 50 mg/L.” Nitrate concentrations more than the 10 mg/L as NO3-N are most common in eastern alluvial fans subregion in the Central Valley of California, USA (Beutel et al. 2017).

Status in India India is one of the topmost country when it comes to groundwater abstraction with an estimated withdrawal rate of 251 km3/year reported in 2010 (Gun 2012). As per the studies conducted by Rai (2003), approximately 117.93 million people in India are drinking water with nitrates level 45–100 mg/L and 108.2 million people with levels more than 100 mg/L of nitrate. A study revealed that more than 367 Indian districts are affected by nitrate contamination above the BIS permissible limit in shallow aquifers as shown in Fig. 5 (CGWB 2012). Jacks and Sharma (1983) found nitrate concentrations as high as 1500 mg/L in Southern India, where higher concentrations were more common in village wells than in the field irrigation wells. In Kanpur district, more than 42% of the groundwater samples collected from shallow aquifers and 26% collected from deep aquifers had been reported by Singh et al. (2006) to have nitrate concentration exceeding the BIS limit. In Rajasthan, Suthar et al. (2009) found out that the average nitrate was 60.6  33.6 (SD) mg/L in some agro-economy-based rural habitation of northern parts of the state. Also, Shyam and Kalwania (2011) found that nearly two-thirds of total samples had more than 50 mg/L of nitrates in Sikar, Rajasthan. Dar et al. (2009) stated that 85% of samples during summers and 67% during winters were observed to have nitrates exceeding the permissible limit in Sopore town, Kashmir, and attributed it toward the nitrogenous fertilizers use in the area. Rai (2003) estimated that around 10.71 million people are drinking water with nitrates level 45–100 mg/L and 3.47 million people with levels more than 100 mg/L in Uttar Pradesh alone. Reddy et al. (2009) revealed that 65% of the samples were found to be unfit for drinking purposes in the pre-monsoon season and 45% in postmonsoon due to excess nitrate in north eastern part of Anantapur District, Andhra Pradesh, where intense agricultural practices, improper sanitation, and organic waste disposal methods were observed to contribute nitrate to the shallow and moderately deep aquifers. A report by Central Ground Water Board (CGWB) reported that groundwater in 42 district in Uttar Pradesh have more than 45 mg/L of nitrates with mean value of 152 mg/L ranging between 47 and 1162 mg/L (CGWB 2012). Lokesh (2013) found that more than 43% samples had higher nitrates than prescribed value in Delhi and attributed it to fast urban growth and subsequent interventions. Nitrate levels in groundwater over vast agricultural areas can be correlated with intensive irrigated agriculture, corresponding use of nitrogenous fertilizers and groundwater development, and consequent diffuse agricultural pollution has already endangered the safety of potable groundwater for future generations in both rural and urban areas. Most of the countries reportedly affected with nitrate contamination in groundwater have almost 100% cropland and affected by heavy fertilizer application; this is evident in Fig. 6, where areas with heavy fertilizer consumption and having 100% cropland are marked with dark red color.

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Fig. 5 Locations with nitrate >45 mg/L in shallow aquifers across India (Source: CGWB 2014)

Health Impacts Nitrate assimilation in humans takes place via drinking water and having food, and it becomes toxic when it is reduced to nitrite in the oral cavity (Lundberg et al. 2009). Nitrate may enter the human body through endogenous and exogenous pathways,

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Fertilizer use Cropland

N W

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Fig. 6 Global comparison of fertilizer application and availability of agricultural areas (Source Potter et al. (2010, 2011))

where endogenous nitrate results from the circulation of nitrite and other nitrogen oxides and exogenous nitrate enters via ingestion through food, drinking water, inhalation, etc. (Fathmawati et al. 2017). Customarily, human ingestion of nitrate is mostly from vegetables consumption, but more than 50% of nitrate intake may come from drinking water sources when levels are above the maximum contaminant level (DeRoos et al. 2003). Several studies have been conducted around the world attempting to correlate the nitrate consumption and its various health impacts. Ingestion of nitrate-contaminated water (NO3 >45 mg/L) may cause one or more of the mentioned health impacts. This additive health impact depends upon the exposure duration and ingested concentration of nitrate, and as the concentration increases, the health risk also increases. These health impacts are discussed in the following sections.

Methemoglobinemia The most threatening of all the health impacts is “methemoglobinemia” also known as “blue baby syndrome” which is caused when water with high nitrate concentration is ingested in infants specially in “bottle-fed neonates” (WHO 2011). When ferrous iron, in the blood, loses an electron and oxidizes to ferric iron, hemoglobin is converted to methemoglobin (MetHb). MetHb reduces the oxygen carrying capacity of the blood. There are various effects (Fig. 7) of formation of methemoglobin in the human body because of ingestion of elevated levels of nitrate, viz., cyanosis (decoloration of blood) and clubbing (appearance of blue color in nails, tongues, etc.). A study by Sadeq et al. (2008) also reported that older children (aged 1–7 years) drinking water containing more than 50 mg/L of nitrate

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Fig. 7 Effects of methemoglobin production in human body, (a) central cyanosis, (b) clubbing, and (c) change in blood color (Source: RamanaMurty 2013)

were significantly more likely to be affected by blue baby syndrome than those drinking less than 50 mg/L of nitrate in drinking water.

Cancer Nitrate is a precursor in the formation of nitrosamines. The International Agency for Research on Cancer (IARC) classified nitrate as a probable human carcinogen (class 2A) that may form N-nitroso compounds (NOCs) through endogenous nitrosation. Likeliness of strong carcinogenic effect of N-nitroso compounds (NOC) in humans based on animal evidence of carcinogenicity in every species tested is also there (Ashok and Hait 2015).

Thyroid Problems Nitrate is a goitrogenic substance, and study by Eskiocak et al. (2005) suggested that histomorphological changes in the thyroid are observed at 250 and 500 mg/L of nitrate dosage. Many experimental studies suggest that nitrates impair thyroid function involving the hypothalamo–hypophysio–thyroid axis.

Impacts on Livestock Health and Aquatic Life Several studies suggested that, at levels between 100 and 200 mg/L, nitrate–nitrogen in water causes decrease in appetite in livestock and asks for better management practices to prevent livestock loss (Sahoo et al. 2016). It has been reported that nitrate concentrations lower than the drinking water standard cause substantial egg and fry mortality in some salmonid fish species (Kincheloe et al. 1979), whereas Alonso and Camargo (2013) found that nitrate did not cause mortality, but it reduced the velocity of movement (at 44.9, 81.8 and 156.1 mg NO3-N/L) and number of live newborns (in all tested concentrations) of aquatic snails.

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Fig. 8 Different adsorbents for nitrate removal (Source: Bhatnagar and Sillanpaa 2011)

Removal Technologies of Nitrate from Drinking Water Globally the most commonly accepted treatment technologies for nitrate removal from drinking water includes chemical denitrification using zero valent iron and magnesium (Ahn et al. 2008), ion exchange (IE) (Samatya et al. 2006), reverse osmosis (RO) (Schoeman and Steyn 2003), electrodialysis (ED), and biological denitrification (Soares 2000). WHO has suggested the use of ion exchange (IX) process which requires waste brine disposal, and posttreatment is also required to avoid corrosion of effluent and biological denitrification as nitrate removal methods. While US EPA has suggested IE, RO (requiring high TDS disposal and has a high operational cost and large volume of reject water), electrodialysis, as Best Available Technologies (BATs) to treat nitrate-contaminated water (Bhatnagar and Sillanpaa 2011). Various adsorbents used for nitrate removal are summarized in Fig. 8 and are discussed in the following sections.

Removal Through Carbon-Based Adsorbents Activated carbon (AC) is considered to be a universal adsorbent, and it removes diverse types of aquatic pollutants. Powdered activated carbon (PAC) and carbon nanotubes (CNTs) were also used for the removal of NO3, where removal capacity of CNTs was found to be higher than PAC (Khani and Mirzaei 2008). Granular activated carbon (GAC) obtained from coconut shells have also been examined for NO3 removal, using steam activation process (Bhatnagar et al. 2008). The adsorption efficiency of bamboo powder charcoal (BPC) in nitrate removal from water has

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been investigated by Mizuta et al. (2004), and he found that uptake values of BPC and commercial activated carbon (CAC) at 10  C were 1.25 and 1.09 mg/g, respectively, showing that the adsorption effectiveness of BPC was higher than that of CAC regardless of the concentration of nitrate and test temperature.

Removal Through Natural Adsorbents Clay being a natural sorbent can also be used for NO3 removal. Calcium bentonite (CB) and calcium montmorillonite (CM) have been used, where CM, activated by hydrochloric acid, showed better removal capacity (Mena-Duran et al. 2007). Xi et al. (2010) found that poor and untreated clays such as Queensland bentonite, kaolinite, and halloysite have shown poor adsorption capacity for nitrate removal. These clay minerals when modified with hexadecyltrimethylammonium bromide showed improved adsorption capacities. The chitosan-coated zeolite (Ch-Z), protonated with either sulfuric or hydrochloric acid, has been tested for its ability to remove nitrate from potable water, where it was found that protonation with hydrochloric acid resulted in a higher nitrate removal capacity as compared to sulfuric acid (Arora et al. 2010).

Removal Through Sorbents Obtained from Agricultural Waste Use of agricultural waste material is another option for nitrate removal and various materials like wheat straw, raw wheat residue, sugarcane bagasse, and rice hull being used for effective NO3 removal. Orlando et al. (2002) investigated the feasibility of lignocellulosic agricultural waste materials (LCM), sugarcane bagasse (BG), and rice hull (RH) for NO3 removal from water. The maximum removal capacity was obtained from PL (1.8 mmol/g and 412.5%), followed by BG (1.41 mmol/g and 300%), PC (1.34 mmol/g and 166%), and RH (1.32 mmol/g and 180%). A study compared the removal capacity of raw wheat straw (RWR) and modified wheat straw (MWR) and found that MWR had greater nitrate adsorbing capacity than RWR (Wang et al. 2007).

Removal Through Sorbents Obtained from Industrial Waste A limited number of industrial wastes have been considered and tested for their suitability to be used as nitrate removal adsorbents. Cengeloglu et al. (2006) compared the nitrate adsorption efficiency of original and activated red mud (produced in alumina industries) and reported that removal capacity of activated mud was higher than original mud. Several other industrial wastes such as slag, fly ash, etc. have also been examined with other adsorbents and found to be useful in removal of nitrate from water. Selection of a suitable technology for nitrate removal from water generally depends on several factors including (a) initial nitrate concentrations, (b) presence

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Fig. 9 Comparison of some nitrate removal technologies (Source: Bhatnagar and Sillanpaa 2011)

and concentration of other ions, (c) adsorbent dose, and (d) variation of pH in water. Thus, the selection of the appropriate technology/sorbent media can become a tedious task. A comparative summary of some of these nitrate removal technologies are shown in Fig. 9 (Shams 2010; Bhatnagar and Sillanpaa 2011).

Conclusion Nitrate is one of the major contaminant of drinking water in most of the areas around the world. Interestingly, the occurrence of elevated nitrate concentration in groundwater is sporadic in nature, and even though having various advanced technologies around the world, this problem is not addressed precisely in terms of its source as well as the release mechanism. It is also clearly not evident that whether the nitrate is leaching from fertilizers or minerals and the reason behind its release mechanism are dynamic and complex. It is very well established that the low-cost filters and BATs can be effectively used for the treatment of nitrate-contaminated waters. But the problem doesn’t end with mitigation as it is an end of the pipe solution, and these technologies create a stream of concentrated nitrates and also disposal problems. There is a dire need of the hour to assess the extent of nitrate contamination in groundwater on case-to-case basis and suitable control and remedial measures shall be adopted.

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Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . The Aim and Objective of this Study . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sources of Classification of Industrial and Automobile Exhausts and Disposal . . . . . . . . . . . . . . Toxicity of Exhausts and their Release to Environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Possibilities of Remediation and Deactivation of Exhausts . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Significant Scientific Endeavor in the Field of Air Pollution Control of Industrial and Automobile Exhausts and Industrial Wastewater Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Visionary Scientific Endeavor in the Field of Environmental Management . . . . . . . . . . . . . . . . . . . Environmental and Energy Sustainability and the Vast Vision for the Future . . . . . . . . . . . . . . . . . Future Research Trends and Future Recommendations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion and Future Perspectives . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

The world of environmental engineering science and environmental protection is witnessing immense challenges today. Environmental management today stands in the midst of immense scientific vision and vast introspection. Air pollution control and industrial wastewater treatment in the similar manner are today in the path of new scientific regeneration. In this paper, the authors deeply address the vast scientific success of remediation of industrial and automobile exhausts for environmental management. Human civilization’s immense scientific prowess, S. Palit (*) Department of Chemical Engineering, University of Petroleum and Energy Studies, Energy Acres, Dehradun, Uttarakhand, India e-mail: [email protected] C. M. Hussain Department of Chemistry and Environmental Science, New Jersey Institute of Technology, Newark, NJ, USA e-mail: [email protected] © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_21

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the vast futuristic vision, and the technological vision of environmental engineering tools will all lead a long and effective way in the true emancipation of environmental management today. Environmental and energy sustainability are the other sides of the visionary coin of environmental engineering today. The global climate change, the loss of ecological biodiversity, and the frequent environmental disasters are urging the scientists and engineers to gear forward toward a more visionary future with respect to innovations and challenges. In this paper the author deeply addresses the various technologies of remediation of industrial and automobile exhausts for environmental management. Global scientific orders are in a state of deep challenges, unending vision, and deep comprehension. The future of environment is immensely dismal. This paper opens up new windows of scientific innovation and scientific instinct in the field of environmental management in decades to come. Keywords

Industrial · Automobile · Exhausts · Vision · Environment · Sustainability

Introduction Environmental management and environmental engineering science are today in the path of immense scientific vision and deep regeneration. Scientific fortitude and scientific ingenuity in application of environmental engineering tools to tackle global air pollution issue is in a dismal state. Industrial and automobile exhausts are the major pollutants of modern civilization today. Human scientific vision in air and water pollution control is witnessing immense and drastic challenges. Scientific endeavor needs to be readdressed and revamped as science and engineering of environmental protection surges forward. Technology and engineering science has today few answers to the marauding crisis of global climate change. In the similar vein, industrial wastewater treatment and drinking water treatment are the burning issues and scientific enigma of modern human civilization. In this paper, the authors pointedly focus on the scientific success and the scientific subtleties of the application of environmental engineering tools to protection of global environment. The advancement of engineering science, the needs of human society, and the futuristic vision of environmental engineering science will all lead a long and visionary way in the true emancipation of environmental sustainability. The success of science of environmental protection is enumerated in this chapter. Environmental management and true realization of environmental sustainability are the two opposite sides of the visionary coin today. Today, global environment is faced with the monstrous issue of water shortage and groundwater heavy metal contamination. Arsenic groundwater contamination is a veritable curse to modern day human civilization and the vast scientific research pursuit. This chapter also is a veritable eye-opener to the scientific difficulties, the scientific barriers, and the vast scientific profundity in the control of industrial and automobile exhausts. This entire chapter opens up new doors of innovation in the field of environmental management also. Today environmental

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management is the need of the hour. Human scientific endeavor in environmental management today stands in the midst of vision and deep scientific fortitude. The challenge and the vision of the science of environmental management are today surpassing vast and versatile visionary frontiers. This paper opens up new chapters in the scientific research pursuit in both environmental engineering and environmental management.

The Aim and Objective of this Study Science and technology are vast avenues of vision and emancipation today. Environmental engineering and industrial wastewater treatment are in the midst of deep scientific introspection and vast scientific vision. The aim and objective of this study goes beyond scientific imagination and replete with scientific profundity and vast scientific acuity. Today the scientific world is faced with immense environmental issues such as global climate change and frequent environmental disasters. Air pollution and industrial exhausts are the causes of grave concern as science and engineering surges forward toward a newer visionary realm. In the similar vein, environmental management should be the hallmark of deep scientific profundity in modern day human civilization. In this paper, the authors rigorously point toward the need of environmental engineering tools to confront the emission of industrial and automobile exhausts with the sole objective of furtherance of science and engineering of environmental protection. Environmental sustainability and environmental engineering science are moving toward newer knowledge dimensions as science and engineering surges forward. Industrial wastewater treatment and drinking water purification are the other cornerstones of this well-researched treatise. Today arsenic and heavy metal groundwater contamination are destroying the scientific landscape of immense vision and scientific might. In this chapter, the authors clearly and poignantly depict the vast scientific vision behind the application of engineering tools in tackling industrial and automobile exhausts (Najjar 2011). Human scientific ingenuity and vast scientific foresight are the necessities of modern science and research pursuit today. This chapter veritably opens up newer scientific innovations and newer scientific insights in the field of air and water pollution control with the sole purpose of furtherance of science and engineering (Najjar 2011).

Sources of Classification of Industrial and Automobile Exhausts and Disposal Environmental pollution control and air pollution control today needs to be re-envisioned and restructured with the passage of scientific history and time. Transportation is a major component of modern civilization’s scientific progress. Transportation is also the heart of modern industrial systems which involves resources and means for their transformation, pumps out products, and provides an inseparable link between the different parts of industrial systems as a vital and

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veritable part of commercial activities. Transportation is also the mobile source of major anthropogenic contribution to environmental protection (Najjar 2011). The sources of industrial air pollution are calcium carbide industry; carbon black industry; cement industry; brick kilns; copper, lead, and zinc smelting units; coal mines; diesel generator sets; vehicular exhaust emission; and a host of other pollutants (Najjar 2011). Exhaust gas is a major component of motor vehicle emissions (and from stationary internal combustion engines) which can include (1) crankcase blowby and (2) evaporation of unused gasoline (Najjar 2011). Motor vehicle emissions contribute to air pollution and are a major ingredient in the creation of smog in large cities. Scientific vision and vast scientific acuity should be the targets of scientific endeavor in air pollution control today. Science of environmental protection should start a newer beginning and new technologies, and newer innovations should be deeply envisioned (Najjar 2011).

Toxicity of Exhausts and their Release to Environment Generally an exhaust gas is a gas emitted through a combustion process. The exhaust gases comprise of many different gases: nitrogen, carbon dioxide, water, and oxygen (Najjar 2011). Though some are harmless, there are a few that are harmful and are considered major pollutants. Ecological engineering and environmental protection are today linked by an unsevered umbilical cord. Urban smog is a dark yellow or brown haze that builds up in a large stagnant air mass and hangs over populated areas on calm hot summer days (Najjar 2011). Smog is made of mainly ground-level ozone, but it also contains numerous other chemicals including carbon monoxide, particulate matter such as soot and dust, and volatile organic compounds such as benzene, butane, and other hydrocarbons. The main source of both nitrogen oxides and hydrocarbons is the motor vehicles. Hydrocarbons and nitrogen oxides react in the presence of sunlight in hot calm days to form ground-level ozone, which is the primary component of smog. Ozone irritates the eyes and damages the air sacs in the lungs where oxygen and carbon dioxide are veritably exchanged, causing eventual hardening of soft and spongy tissues. Technology and engineering science of environmental protection should thus be re-envisioned as regards health effects and more robust pollution control strategies (Najjar 2011).

Possibilities of Remediation and Deactivation of Exhausts Catalytic converters are used to reduce emissions in cars (Najjar 2011). Actually an automotive catalyst uses precious metals: platinum, rhodium, and palladium. When toxic substances (carbon monoxide, hydrocarbons, and nitrogen oxides) are in contact with these metals, they are immediately converted. Today ecological engineering and water process engineering are on the path of newer scientific rejuvenation (Najjar 2011). Many nations throughout the world depend on oil for energy uses and sustainable development, hence import all of its oil and gas needs.

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Thereby, consequent harmful and disastrous effects on the environment, such as global warming, ozone, smog, and acid rain, may result from the gases emanating from fossil fuel combustion (Najjar 2011). Biofuel technologies are another avenue of reducing exhaust gas emissions. Increased use of biomass-based renewable fuel can be another visionary opportunity for minimizing air polluting emissions (Najjar 2011). While conventional biofuel is derived from corn starch, advanced biofuel is produced from other renewable biomasses such as (1) cellulose, hemicelluloses, and lignin, (2) sugar and nonsugar starch, (3) waste material such as agricultural crop, (4) planted trees and tree residues, (5) animal wastes, (6) algae, and (7) separated food wastes (Najjar 2011). This is a major avenue of scientific research pursuit today. Plasma-assisted catalyst for nitrogen oxides remediation from lean gas exhausts is a major scientific endeavor today (Najjar 2011).

Significant Scientific Endeavor in the Field of Air Pollution Control of Industrial and Automobile Exhausts and Industrial Wastewater Treatment Air pollution control and industrial wastewater treatment today stand in the midst of vast vision and scientific fortitude. Environmental calamities and global climate change are today challenging the scientific firmament and veritably opening up new doors of scientific innovation and scientific instinct in decades to come. Research and development initiatives in the field of water science and air pollution control are ushering in a new era in the field of scientific emancipation and technological validation. Validation of science and engineering are the utmost need of the hour as human civilization surges forward. Drinking water purification, desalination, and industrial wastewater treatment are the needs of human society today. In this chapter, the author pointedly focuses on the vast scientific success, the scientific subtleties, and the deep scientific ingenuity in the field of air pollution control, remediation of industrial and automobile exhausts, and the vast world of scientific validation. Rajendran et al. (2003) deeply discussed with cogent and lucid insight microbes in heavy metal remediation. Technological candor, the deep scientific vision and the vast world of scientific validation are the pillars of science today. This treatise is a vast scientific discernment and a deep scientific vision in the field of heavy metal remediation. Heavy metal contamination due to natural and anthropogenic sources is a global environmental concern. Drinking water treatment and industrial wastewater treatment today stand in the midst of vision and deep scientific ingenuity. Release of heavy metal without proper treatment poses a deep and significant threat to public health and hygiene because of its persistence, biomagnification, and accumulation in food chain (Rajendran et al. 2003). Human scientific endeavor and research and development initiatives in water technology are today ushering in a new era in science and engineering. Non-biodegradability and sludge production stand as a major scientific impediment in human civilization today (Rajendran et al. 2003). Recent advances have been made in understanding metal-microbe interaction and

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their application for metal accumulation. Human scientific progress in biochemical engineering and environmental science thus needs to be re-envisioned as engineering science and technology moves from one visionary paradigm toward another (Rajendran et al. 2003). There is a greater need of metal bioremediation. The technology of bioremediation is highly advanced today and surpassing vast and versatile scientific boundaries. Greater awareness of the ecological effects of toxic metals and their biomagnifications through food chain as well as environmental disasters such as mercury pollution in Minamata, Japan, has vastly prompted a demand for decontamination of heavy metals in the aquatic environments (Rajendran et al. 2003). The challenge and vision of global research and development initiatives in water pollution control are immense and pathbreaking. Essential metals are required for enzyme catalysis, nutrient transport, protein structure, charge neutralization, and the need for control of osmotic pressure. A number of heavy metals are required as micronutrients in plants (Rajendran et al. 2003). They act as cofactors as part of prosthetic groups of enzymes which are involved in a wide variety of metabolic and scientific pathways (Rajendran et al. 2003). Metal concentrations have been linked to birth defects, cancer, skin lesions, retardation leading to disabilities, liver and kidney damage, and a host of other health issues. This paper pointedly focuses on the scientific success and the scientific ingenuity in the application of microbes in bioremediation. Human scientific and technological innovations in bioengineering are today vast and versatile scientific boundaries. The world of science and engineering is today in the midst of vision and vast scientific articulation. The authors in this paper discuss conventional methods of heavy metal removal from soil and water, microbes for metal remediation, bioremediation techniques, mechanisms for metal tolerance, genetic engineering for metal remediation, biosorption, immobilization for metal remediation, and plant-microbe interaction in metal remediation (Rajendran et al. 2003). The authors in this treatise deeply comprehend the success of biochemical engineering and biochemistry in metal remediation (Rajendran et al. 2003). Air pollution control is a vast area of scientific endeavor today. Human scientific challenges and the vast scientific genre in bioengineering and water technology are deeply addressed in this paper. Shannon et al. (2008) deeply discussed with lucid and cogent insight science and technology for water purification in the coming decades. One of the enigmatic issues of scientific endeavor and scientific civilization today is inadequate access to clean water and sanitation. Science and engineering of water purification is today ushering in a new era in the field of global water shortage issues. Problems with water are expected to grow worse in the coming decades, with water shortage occurring globally, even in water-rich regions throughout the world (Shannon et al. 2008). Addressing these issues calls out for a tremendous effort to be conducted to identify robust new innovations of purifying water at a low cost and with less energy while at the same time minimizing the use of chemicals and preventing the impact on environment. The vast many problems worldwide associated with the lack of clean, fresh water are vastly well known: 1.2 billion people lack proper access to safe drinking water, 2.6 billion have little or no sanitation, and millions of people die annually – 3900 children a day – from diseases transmitted through impure and

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unsafe water (Shannon et al. 2008). Human mankind’s immense scientific prowess, the futuristic vision of water technology, and the imminent needs of pure drinking water will go a long and visionary way in the true emancipation of water process engineering today. Today chemical process engineering and environmental engineering science are in the avenue of vast and versatile scientific ingenuity. Technology and engineering have few answers to the monstrous issue of global water scarcity and groundwater contamination. The authors in this paper deeply discussed the concepts of disinfection, decontamination, reuse, and reclamation and desalination. The need for global water research and development initiatives is elucidated in details in this paper (Shannon et al. 2008). The overarching goal of this research paper is to target the needs of water process engineering, the technological vision, and the vast scientific profundity (Shannon et al. 2008). The US Environmental Protection Agency Report (2010) deeply discussed, with immense scientific conscience, green remediation best management practices and clean fuel and emission technologies for site cleanup (United States Environmental Protection Agency Report 2010). Mankind’s immense scientific prowess, man’s vast scientific vision, and the futuristic vision of basic human needs will lead a long and visionary way in the true emancipation of environmental management and environmental sustainability (United States Environmental Protection Agency Report 2010). This report outlines the Agency’s policy for evaluating and minimizing the environmental footprint of activities undertaken when cleaning up the contaminated site. Human scientific regeneration and vision today stand in the midst of deep crisis. Cleanup of hazardous waste sites can involve significant consumption of gasoline, diesel, or other fuels by mobile and stationary sources. This report with deep scientific conscience discussed operations and maintenance, advanced diesel technologies, alternative fuels and additives, and the wide domain of alternative vehicles. The need of remediation technologies in industrial and automobile exhausts is today immense and pathbreaking. Mankind’s immense scientific challenges in air pollution technologies and the innovations behind it are deeply discussed in this report (United States Environmental Protection Agency Report 2010). Bioremediation of contaminated soils and its in situ and ex situ techniques are the challenges of science and engineering today. For the treatment of contaminated soils, the application of biotreatment is growing at a rapid pace today. Scientific progress, scientific acuity, and deep scientific farsightedness are the imminent necessities of bioremediation today. In this chapter, the authors deeply discussed with cogent insight the success, the ingenuity, and the fortitude behind scientific emancipation of remediation of industrial exhausts. The US Environmental Protection Agency Report (1998) deeply discussed with vast insight an analysis of composting technologies as an environmental remediation technique. The composting process is currently viewed primarily as a waste management method to stabilize organic waste such as manure, yard trimmings, municipal biosolids, and organic urban wastes (United States Environmental Protection Agency Report 1998). Human civilization and human scientific endeavor today stand in the midst of deep scientific vision and vast forbearance. This report summarizes the available information on the use of compost for managing hazardous

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waste streams and widely investigates possible future recommendations of remediation technologies (United States Environmental Protection Agency Report 1998). A deep scientific discernment on cross-media transfer of contaminants during the implementation of various bioremediation technologies is addressed in details in this report. Groundwater remediation and bioremediation are the absolute needs of the human civilization and the human scientific endeavor today. Composting is a managed system that uses microbial activity to degrade raw organic materials, such as yard trimmings, so that the end product is relatively stable, reduced in quantity, and free from offensive odors (United States Environmental Protection Agency Report 1998). Technology and engineering science of bioremediation is today attaining visionary heights as global concerns for environment aggravate. This report opens up new ideas of scientific innovation and scientific instinct in the field of composting technology and bioremediation (United States Environmental Protection Agency Report 1998). The Central Pollution Control Board Report (2010), India, discussed with deep and cogent insight the study of the exhaust gases from different fuel-based vehicles for carbonyls and methane emissions. Air pollution is caused by a number of pollutants evolved from various sources. Under air pollution control strategy, the detailed analysis and assessment of all the pollutants having detrimental effects on human health and environment are required (Central Pollution Control Board, India Report 2010). Thus the need of this vast scientific vision and scientific endeavor. A developing country like India today is in the midst of a vast scientific crisis and technological vision. The authors firstly did a comprehensive literature review and then did a detailed methodology and after that delineated and discussed future recommendations. Scientific progress, the needs for air pollution control strategy, and the visionary technologies are the veritable torchbearers of environmental sustainability today. Environmental sustainability and infrastructural development both today stand in dire straits and vast scientific catastrophe. Technology and engineering science needs to be re-envisioned and thoroughly addressed with the passage of scientific history and time. The success of human scientific ingenuity in air pollution control strategies is elaborated deeply in this report (Central Pollution Control Board, India Report 2010). The International Energy Agency Report (2008) did a comprehensive study of carbon dioxide capture in the cement industry. The knowledge dimensions, the vast futuristic vision of air pollution control, and the human scientific needs will all lead a long and visionary way in the true emancipation of environmental sustainability today. Today air pollution control is a visionary parameter of scientific endeavor in environmental protection. The cement industry is one of the world’s largest industrial sources of carbon dioxide emissions accounting for more than 6% of global emissions from the use of fossil fuels. Scientific validation and the vast technological vision are the necessities of research and development initiatives in air pollution control today. Human mankind today stands in the midst of deep scientific discernment and definite vision (International Energy Agency Report 2008). Over the years, the cement industry has substantially reduced emissions of carbon dioxide per tonne of cement by improving energy efficiency, replacing fossil fuels with wastes which

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can sometimes be regarded as “carbon neutral,” and by increasing the use of additives in the cement product (International Energy Agency Report 2008). Scientific rejuvenation and deep scientific understanding are the hallmarks of human civilization today. In this report, the author deeply comprehends the effect of greenhouse gas emissions in cement industry. The report describes cement plants and the global cement industry, reviews the carbon dioxide capture processes, evaluates the performance and economics of cement plants, discusses retrofitting carbon dioxide capture, and vastly identifies vast research and development needs. Today the world of engineering and technology is in the avenue of a vast scientific regeneration. This report is a case study of the human scientific progress in air pollution control strategies in cement industries (International Energy Agency Report 2008). Environmental management today stands in the midst of deep scientific introspection and vast scientific foresight. Today is the age of nanotechnology and nanomaterials. Nanomaterials or smart materials are the technological innovations of modern science and modern civilization. Environmental analysis and environmental protection are today linked by an unsevered umbilical cord. Aghabozorg and Hassani (2017) discussed with vast insight the removal of pollutants from the environment using sorbents and nanocatalysts. The authors in this well-researched chapter discussed the removal of sulfur compounds from fuels, elimination of heavy metals from wastewater, and separation of the dangerous radionuclides from liquid nuclear wastes (Aghabozorg and Hassani2017). Pollutants are generally waste materials of different forms that veritably effect the balance in the environment. Industrial modernization and progress over the years have resulted in a burgeoning environmental impact. Chemical and petroleum industries introduce many pollutants into the surrounding environment (Aghabozorg and Hassani 2017). The authors in this chapter delineated with precision the removal of sulfur compounds from fuels. There are different kinds of sulfur compounds in hydrocarbon fuels, owing to new environmental regulations. Thus the status of the environment is immensely dismal. Human civilization thus is in a state of immense distress (Aghabozorg and Hassani 2017). Many research and development forays have been performed about the removal of sulfur compounds from the middle distillate fraction, such as gasoline, jet and diesel fuels, and kerosene, to produce clean fuels. In this aspect, hydrodesulfurization is an important catalytic process that has been developed for decreasing the sulfur contents in engine fuels to less than 2 ppm (Aghabozorg and Hassani 2017). So here the authors stressed on the importance of technological innovations in sulfur removal. The other upshot of this process is the elimination of heavy metals from wastewater. Heavy metals, such as mercury, arsenic, lead, cadmium, and chromium, can contaminate the environment owing to their toxicity. The unit operation in chemical engineering touched upon is adsorption (Aghabozorg and Hassani 2017). The other areas of this scientific endeavor are the separation of the dangerous radionuclides from liquid nuclear wastes (Aghabozorg and Hassani 2017). Radiological contamination is the presence of radioactive materials within solids, liquids, or gases where their presence is not desired. This contamination rises to an immense health hazard and a burning issue of

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environmental engineering science. Human scientific endeavor in environmental protection today is in a state of immense catastrophe and deep comprehension (Aghabozorg and Hassani 2017). Accidental release of radioactive elements (such as uranium, cesium, and strontium, which are serious pollutants) to the air leads to a massive deposition of these metals into the soil, vegetation, and water. Thus environmental protection is the utmost need of the hour. In this chapter, the author with a deep precision and insight delineates pollutant removal technologies with a clear view of furtherance of science and engineering (Aghabozorg and Hassani 2017). Azzaza et al. (2017) in a phenomenal research endeavor discussed nanomaterials for heavy metal removal. Toxic metals (called heavy metals) are a part of the earth’s crust. They occur in rocks and are predominantly distributed in the environment (soil, water, and air) (Azzaza et al. 2017). The challenge and the vision of engineering science and technology are today far-reaching and surpassing the vast and versatile scientific frontiers. Environmental engineering is today a frontier science today. Human scientific progress in environmental engineering science needs to be re-envisioned and revitalized as civilization surges forward. In order to make the environment healthier for human beings, contaminated water streams and contaminated lands need to be rectified to free them from heavy metals and free elements. The types of adsorbents used in such a case are carbon-based nanomaterials, activated carbon, carbon nanotubes, graphenes, metal-based nanomaterials, and nanosized metal oxides (Azzaza et al. 2017). Human scientific fortitude and human scientific farsightedness in the field of research forays in environmental engineering are today ushering in a new eon of vision and might. The authors in this chapter deeply ponder upon the newer applications of nanomaterials in heavy metal removal and environmental protection in a broader sense (Azzaza et al. 2017).

Visionary Scientific Endeavor in the Field of Environmental Management Environmental management is today the utmost need of human civilization today. The state of global environment is in a state of immense scientific distress and replete with scientific barriers. Yet science and engineering of environmental protection is today surpassing vast and versatile scientific frontiers. In the similar manner, environmental management is a frontier science today. The need for environmental management is immense for human society today. Human civilization today stands in the midst of immense scientific barriers and deep scientific vision. Environmental sustainability and environmental protection are in the process of vast scientific regeneration. In this section the author deeply ponders on the vast scientific research and vast scientific ingenuity in the application of environmental management in human scientific progress. Technology and science are too ebullient and visionary in the path toward scientific emancipation today. Scientific endeavor in the field of environmental management needs to be re-envisioned and restructured as science moves toward a newer era of environmental engineering emancipation.

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Rusko and Prochazkova (2011) discussed with deep and cogent insight the solution to the problems of the sustainable development management. Human scientific insight in environmental management is today in the path of immense rejuvenation. This paper shows that environment is one of the basic public assets of a human system, and it must be therefore specially protected and re-envisioned. According to the present scientific world, the sustainability science is necessary for all human systems, and it is absolutely necessary to invoke the sustainable development principles in all human assets and all human genres (Rusko and Prochazkova 2011). The vision of Dr. Gro Harlem Brundtland, former Prime Minister of Norway today, needs to be restructured and re-envisaged as civilization moves forward. Sustainable development is understood as a development that veritably does not erode ecological, social, or political system on which it depends, but it explicitly approves ecological limitation under the economic activity frame, and it has full comprehension for support of vast human needs. This well-researched paper summarizes the results of the systematic study of the environment in the last 30 years (Rusko and Prochazkova 2011). Social sustainability and economic sustainability are today the cornerstones of scientific research pursuit in environmental sustainability and environmental protection today. The vision of science and technology is today groundbreaking and too effervescent with the march of environmental management. This paper outlines the tools, methods, and techniques used to solve the environment problems and the tasks of executive governance in the environmental protection segment (Rusko and Prochazkova 2011). The environment itself is a system of systems that, from the viewpoint of human existence and development, is a part of the superior system of systems, the human system. Based on the recent cognition, sustainable development is not only related to the environment but also to the entire human system and its basic assets (i.e., public assets) on which the human lives are dependent (Rusko and Prochazkova 2011). Sustainable development, infrastructural development, and environmental sustainability are the necessities of human civilization today. This paper opened a new chapter in the field of sustainable development management along with a deep comprehension of environmental sustainability.(Rusko and Prochazkova 2011) The US Department of Energy (2016) delved deep into the basic research needs for environmental management. The upshot of the paper detailed waste stream characterization, transformation, and separations, waste forms, contaminant fate and transport in geological environments, exploitation of complex speciation and reactivity far from equilibrium, understanding of chemical and physical processes at interfaces, harnessing of physical and chemical processes to revolutionize separations, mechanics of materials degradation in environment, predictive understanding of subsurface system behavior and response to perturbations, and transformative research capabilities (Office of Science, Department of Energy, USA Report 2016). Human scientific progress in environmental management is in the process of new scientific regeneration. Technology and engineering science have few answers to the scientific difficulties and scientific travails of environmental management (Office of Science, Department of Energy, USA Report 2016). The authors deeply delve into the success of the science of environmental management. Human civilization today

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stands in the midst of deep scientific comprehension and vast vision of its own. This report vastly opens a new chapter in the field of environmental management and environmental sustainability with a clear scientific conscience and clear scientific profundity. Scientific discernment and deep scientific vision are all leading a long and visionary way toward the true emancipation of environmental management. Sustainable development is showing a new path in the domain of environmental engineering science. The management of environment and the vast research needs in its true realization are the hallmarks of this research endeavor (Office of Science, Department of Energy, USA Report 2016). Ahmad et al. (2009), with immense lucidity, discussed environmental management system and sustainability. Technological validation and deep scientific validation are the research needs in environmental management systems today (Ahmad et al. 2009). To check rapidly the deteriorating environmental conditions, many management tools are being used in different industries. Technology and engineering science are the two huge pillars with a definite vision of its own. Today the corporate world is highly interested in the application of environmental system in its industries. Environmental science and environmental sustainability are the other areas of research need today. There is a strong need to check how these environmental management systems are rendering environmental management services. This paper opens a new window in the framework for strategic sustainable development and strategic life cycle management (Ahmad et al. 2009). Scientific candor, vast scientific profundity, and the needs of human society are today ushering in a new era in the field of strategic life cycle management. The use of management tools in the application of sustainability to modern society is deeply elaborated in this chapter (Ahmad et al. 2009). Kumar and Katoria (2013) discussed with deep and cogent farsightedness air pollution and its control measures. An important parameter in environmental management system is the vast domain of air pollution control and industrial water pollution control. In this paper, the authors deeply pronounce the success of scientific vision and vast scientific ingenuity in the application of air pollution control measures. Air pollution is basically the foreign material in the air, can be man-made or occur naturally, and is concentrated in area where people are predominantly present. Pollution is injurious to health and its prevention places an economic burden on the human mankind. The aim of this study is to understand the innovation activity in technological and scientific domain and the different ways to observe patterns in relation to diffusion of innovation in different jurisdictions. The authors emphasize on prominent Indian company active in air pollution control measures in innovative technology business and research and development initiatives. Scientific progress in air pollution control is today in the state of immense scientific regeneration and deep vision (Kumar and Katoria 2013). This paper highlights the various countries on priority list of innovative technologies for the protection and exploitation of developed technologies. Human scientific endeavor is today in the state of immense scientific regeneration and deep revival. Air pollution control stands in the midst of vast and wide scientific revival (Kumar and Katoria 2013). The World Bank Report (1996) vastly comprehends the standards and technologies for controlling emissions. The report deeply delineates emission standards

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and regulations in many developed and developing countries, quantification of vehicle emissions, vehicle technology for controlling emissions, controlling of emissions from in-use vehicles, and fuel options for controlling emissions. Because of their versatility, flexibility, and low initial cost, motorized road vehicles overwhelmingly dominate the markets for passenger and freight transport throughout the world (The World Bank Report 1996). Thus the need of stringent emission control regulations and stringent restrictions. The futuristic vision of air pollution control, the vast technological vision, and the deep scientific profundity will all lead a long and effective way in the true emancipation of environmental legislations today. This is a comprehensive report which elucidates on the present and future of environmental legislations globally. In all but the poorest developing countries around the world, economic growth has ushered in an era of increase in the number and use of motor vehicles (The World Bank Report 1996). Technological subtleties and deep engineering succor are the cornerstones of scientific validation today. A deep scientific thought and a vast scientific understanding are the utmost need of the hour as air pollution control technologies move from one visionary paradigm over another (The World Bank Report 1996). Apart from the other vehicles, motorized road vehicles will retain their overwhelming presence in road transport. This handbook presents a state-of-the-art review of vehicle emission standards and testing procedures and vastly delineates worldwide experience with vehicle emission control innovations and its application in both industrialized, developed and developing countries throughout the world. The overarching goal of this handbook and report is to underpin the World Bank’s overall objective of promoting transport development that is vastly environmentally sustainable and totally replete with scientific vision and scientific clarity (The World Bank Report 1996). Seetharam (2014) discussed with immense foresight automobile exhaust pollution and its effective control. With increase of standard of living and growing need for faster road communication, the number of automobiles has been steadily increasing and also pollution from them. The exhausts from them are causing serious threats to public health and hygiene and the quality of life. In addition poor maintenance of roads, adulteration of fuels, improper maintenance of vehicles, and heavy vehicular inflow also vastly aggravate the problem (Seetharam 2014). The pace of urbanization resulting in an enormous increase in population and industrial activities has resulted a huge demand of traffic. The factors influencing emissions are air-fuel ratio, spark timing, carburetion, and surface-to-volume ratio of the combustion chamber. Human scientific research pursuit and scientific acuity are the needs of modern day human civilization. Remediation of industrial and automobile exhausts is the immediate need of scientific vision and profundity today (Seetharam 2014). The sources of automobile emissions are exhaust emissions, evaporative emissions, carburetor emissions, and crankcase emissions. Thus the need of intense scientific innovations (Seetharam 2014). The pollution control measures are fuel change, engine modification, stratified charge engine, reduction of carbon monoxide emissions, decrease of NOx emissions, and reduction of traffic jams in metropolitan cities around the world. The author in this paper

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pointedly focuses on the vast scientific success, the immediate needs of scientific farsightedness, and the vast scientific vision of air pollution control (Seetharam 2014). Human scientific vision and deep scientific fortitude are today in the path of newer regeneration. Global concerns for environment are ever growing due to the severities of climate change, loss of ecological biodiversity, and frequent environmental catastrophes. Technology, engineering, and science have few answers to the marauding issue of water shortage, air pollution, and industrial wastewater treatment. Thus here comes the need of environmental management and environmental sustainability. In this chapter, the authors rigorously delineate the vision of environmental management systems and its application to human society and human scientific progress (Palit 2017a, b; Hashim et al. 2011).

Environmental and Energy Sustainability and the Vast Vision for the Future Environmental and energy sustainability are the needs of human society and modern science today. Sustainable development, infrastructural development, and the futuristic vision of scientific validation will all lead a long and visionary way in the true emancipation and the true realization of environmental engineering techniques and environmental sustainability. The vision for the future in the implementation of sustainability science to human society needs to be redrafted and reorganized with the passage of scientific history and visionary timeframe. Air pollution control techniques and control of industrial and automobile exhausts are the utmost needs of human society today. This paper deeply comprehends the scientific success, the vast scientific profundity, and the imminent scientific needs of human civilization today. The perspectives of science and engineering are vast and versatile today. In the similar vein, scientific discernment and scientific perception in the field of environmental protection need to be re-envision with human scientific progress. Water purification technologies and industrial wastewater treatment are the other needs of human society today. Zero-discharge norms, vast environmental regulations, and the deep scientific acuity are the veritable torchbearers toward a true realization of environmental sustainability. The glowing vision of Dr. Gro Harlem Brundtland, former Prime Minister of Norway on the domain of “sustainability science,” needs to be readdressed and re-envisioned as global climate change crisis is tackled. Scientific succor and deep scientific subtleties are the necessities of inventions and vast innovations today. The authors in this treatise deeply enumerate the scientific success, the scientific needs, and the vast scientific discernment in the field of air and water pollution control technologies today. Man’s immense scientific prowess, human mankind’s vast scientific needs, and the need for scientific validation will be the torchbearers of the future of science and technology. This paper goes beyond scientific imagination and scientific barriers in addressing the success of scientific and technological validation today (Palit 2015, 2016a, b).

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Future Research Trends and Future Recommendations The science of environmental engineering and applied science are today in the midst of deep scientific discernment and vast scientific foresight. Scientific research forays in industrial and automobile exhaust remediation need to be readdressed and revamped as human civilization moves forward. Today environmental science and engineering are huge colossus with a vast and definite vision of its own. Technology has few answers to the burning issue of groundwater heavy metal contamination. Arsenic groundwater contamination is a scientific enigma in developed and developing countries throughout the world (Cheryan 1998). Big cities and metropolis throughout the world are faced with the difficult situation of automobile and industrial exhausts which are destroying the environment (Cheryan 1998). The vision of science, the needs of environmental engineering, and the futuristic vision of air pollution control will go a long and effective way in the true realization and true emancipation of environmental sustainability. Future research trends should be targeted toward scientific emancipation in environmental management and sustainable development. Today nanotechnology and nanomaterials are integrated with every branch of scientific endeavor. This also should be the future recommendations of this study. Human scientific vision will only be realized if the basic human needs such as water and energy are met. Here comes the importance of energy and environmental sustainability. Human mankind’s immense scientific determination, the innumerable technological challenges, and the needs of sustainable development will surely open new windows of scientific innovation in decades to come (Cheryan 1998).

Conclusion and Future Perspectives Science and technology are the two huge colossuses with a definite and vast vision of its own. Environmental engineering science and air pollution control today stand in the midst of deep scientific comprehension and acuity. Future perspectives in scientific research pursuit in environmental pollution control should be targeted toward innovations and scientific instinct. Global climate change is today challenging the scientific firmament of vision and foresight. The science of sustainability today stands in the midst of deep scientific introspection and vast scientific challenges. Human scientific struggle in the field of environmental protection is highly challenged today. Technology and engineering science in such a crucial juncture of human history needs to be restructured and readdressed as human mankind moves from one visionary paradigm to another. Global water research and development initiatives in the similar manner need to be revamped as science and technology surges forward. Industrial wastewater treatment and drinking water treatment should be equally revitalized and reorganized with the march of science. In this paper, the author elucidates with deep and cogent insight the need for environmental sustainability and environmental protection in tackling the intricacies of scientific research pursuit. The challenge and vision of environmental engineering, chemical process engineering, and applied science are in the process of newer scientific regeneration.

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The authors in this treatise deeply fulfill the utmost need of scientific literature and scientific initiatives in tackling global environmental engineering issues. Future recommendations and future flow of scientific thoughts should be vastly targeted toward the provision of basic human needs, the needs of energy security, and the vast world of research and development initiatives in water science and technology. Energy sustainability is another vast avenue of scientific endeavor today. Human scientific challenge in this area is immense and groundbreaking. Energy security and energy sustainability are today connected by an unsevered umbilical cord. This treatise opens up new scientific understanding and the vast scientific vision in energy sustainability issues of human society. Environmental management is today another big and enigmatic scientific issue. Academic and scientific rigors in environmental management are deeply elucidated in this paper. The authors in this paper vastly comprehend the basic needs of human society today such as water, energy, and food. The challenge and the vision of scientific research pursuit in environmental management and environmental pollution control stand as the major pillars of this entire paper.

References Aghabozorg HR, Hassani SS (2017) Chapter 4, Removal of pollutants from the environment using sorbents and nanocatalysts. In: Hussein CM, Kharisov B (eds) Advanced environmental analysis: application of nanomaterials, vol 1. Royal Society of Chemistry Detection Science, London, United Kingdom, pp 74–89 Ahmad SS, Saha PK, Abbasi A, Khan M (2009) Environmental management systems and sustainability: Integrating sustainability in environmental management systems. Master of Strategic Leadership towards Sustainability Thesis, Blekinge Institute of Technology, Karlskrona, Sweden Azzaza S, Thinesh Kumar R, Vijaya JJ, Bououdina M (2017) Chapter 7, Nanomaterials for heavy metal removal. In: Hussein CM, Kharisov B (eds) Advanced environmental analysis: application of nanomaterials, vol 1. Royal Society of Chemistry Detection Science, London, United Kingdom, pp 139–166 Central Pollution Control Board, India Report (2010) Study of the exhaust gases from different fuel based vehicles for carbonyls and methane emissions (Control of Urban Pollution Series) Cheryan M (1998) Ultrafiltration and microfiltration handbook. Technomic Publishing Company Inc, Lancaster, USA Hashim MA, Mukhopadhyay S, Sahu JN, Sengupta B (2011) Remediation technologies for heavy metal contaminated groundwater. J Environ Manag 92:2355–2388 International Energy Agency Report (2008) CO2 capture in the cement industry, Report No.-2008/3 Kumar S, Katoria D (2013) Air pollution and its control measures. Int J Environ Eng Manage 4(5):445–450 Najjar YSH (2011) Gaseous pollutants formation and their harmful effects on human health and environment. In: Innovative energy policies, vol 1. Ashdin Publishing, Honnelles, Belgium, pp 1–8 Office of Science, Department of Energy, USA Report (2016) Basic Research needs for environmental management. Report of the Office of Science Workshop on Environmental Management, 8–11 July 2016 Palit S (2015) Advanced oxidation processes, nanofiltration, and application of bubble column reactor. In: Kharisov BI, Kharissova OV, Rasika Dias HV (eds) Nanomaterials for environmental protection. Wiley, New York, USA, pp 207–215

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Palit S (2016a) Nanofiltration and ultrafiltration- the next generation environmental engineering tool and a vision for the future. International Journal of Chem Tech Research 9(5):848–856 Palit S (2016b) Filtration: Frontiers of the engineering and science of nanofiltration-a far-reaching review. In: Ortiz-Mendez U, Kharissova OV, Kharisov BI (eds) CRC concise Encyclopedia of nanotechnology. Taylor and Francis, Boca Raton, pp 205–214 Palit S (2017a) Chapter 14, Advanced environmental engineering separation processes, environmental analysis and application of nanotechnology – a far-reaching review. In: Hussain CM, Kharisov B (eds) Advanced environmental analysis: application of nanomaterials, vol 1. Royal Society of Chemistry, London, United Kingdom, pp 377–416 Palit S (2017b) Chapter 17, Application of nanotechnology, nanofiltration and drinking and wastewater treatment – a vision for the future. In: Grumezescu AM (ed) Water purification. Academic Press, New York, USA, pp 587–620 Rajendran P, Muthukrishnan J, Gunasekaran P (2003) Microbes in heavy metal remediation. Indian J Exp Biol 41:935–944 Rusko M, Prochazkova D (2011) Solutions to the problems of the sustainable development management. Research Papers, Faculty of Materials Science and Technology in TRNAVA, Slovak University of Technology in Bratislava, Slovakia Seetharam KA (2014) Automobile exhaust pollution. J Chem Pharm Sci Special Issue:73–74 Shannon MA, Bohn PW, Elimelech M, Georgiadis JG, Marinas BJ, Mayes AM (2008) Science and technology for water purification in the coming decades, vol 452. Nature Publishing Group, London, United Kingdom, pp 301–310 The World Bank Report (1996) Air pollution from motor vehicles: standards and technologies for controlling emissions (Faiz A, Weaver CS, Walsh MP) United States Environmental Protection Agency Report (1998) An analysis of composting as an environmental remediation technology. Office of the Solid Waste and Emergency Response, Washington D.C., USA United States Environmental Protection Agency Report (2010) Green remediation best management practices: clean fuel and emission technologies for site clean-up. Office of Solid Waste and Emergency Response, Washington D.C., USA

Management of Residues from Air Pollution Control Systems

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Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Air Pollution Control . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . General . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Air Pollution Control Tasks and Residues Thereof . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Residues from Air Pollution Control in Major Industry Sectors . . . . . . . . . . . . . . . . . . . . . . . . . . . Enrichment of Volatile Elements in the Off-Gas . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Ways of Utilization of Air Pollution Control Residues . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Use as Products . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Direct Recycling . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Treatment of Residues for Recycling or Utilization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Use of Residues in Another Process . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . New Ways of Utilization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Landfill . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . General . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Backfilling in Mines . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Solidification . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

This chapter gives an overview of the residues generated in various industrial air pollution control processes. After summarizing the amount of resulting residues in various industries the current ways of utilization, treatment, and disposal of these residues as well as actual developments and future prospects are reviewed.

C. Lanzerstorfer (*) University of Applied Sciences Upper Austria, Wels, Austria e-mail: [email protected] © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_153

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Keywords

Air pollution control · Waste hierarchy · Reduce · Reuse · Recycle · Energy use · Disposal · Economical solution · Emission abatement

Introduction Air pollution control (APC) systems have become important in all industrial production plants. The pollutants previously emitted to the atmosphere are now separated from the off-gas and remain as residues from the APC systems. In the selection of an APC process the most important requirement is to achieve compliance with the legal emission limits. Another important target is to find the most economical solution for the given emission abatement task. In the economic evaluation of different processes, the costs incurred for the treatment or disposal of the remaining residues can play an important role (Quina et al. 2011; Karatepe 2000). From an environmental point of view, the different levels of the waste hierarchy, reduce, reuse, recycle, energy use, and disposal in descending order should also be taken into account.

Air Pollution Control General APC processes are typically classified into two main categories, wet processes and dry processes. In dry processes no liquid effluent from the process exists, whereas in wet processes a wastewater stream is usually generated. Under certain conditions some wet processes can be operated without discharging wastewater. Table 1 provides an overview of important APC processes for the main pollutants as well as the types of residues produced. Besides the type of pollutants to be separated and the off-gas parameters the selection of an APC process has also to consider several circumstances with respect to residues: • • • •

The possibility of discharging wastewater Local proximity of potential customers for residual products Local proximity of potential treatment facilities for APC residues Cost and local proximity of landfills

Air Pollution Control Tasks and Residues Thereof De-dusting In a dry de-dusting process, the composition of the separated dust is usually the same as the composition of the emitted dust. However, there are also combined

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Table 1 Important off-gas cleaning processes and remaining residues Pollutant Particles, dust

SO2, SO3

H2S and/or CS2

HCl, HF

NOx

NH3 CxHy or CxHyOz (volatile organic compounds)

Chlorinated hydrocarbons

a

Process/equipment Cyclone, inertia separator Electrostatic precipitator (ESP) Filter (pulse jet filter, baghouse, filter with ceramic filter candles, . . .) Scrubber (Venturi scrubber, nozzle scrubber, packed tower, . . .) Semidry and dry chemisorption processes Limestone flue gas desulphurization (FGD) Regenerative sorption processes (active coke, sodium sulfite, organic solvents) Absorption in basic solutions (NaOH, aqueous ammonia) Regenerative sorption processes (activated carbon, organic solvents) Absorption (H2O2, NaOH) Semidry and dry chemisorption processes Scrubber Selective catalytic reduction (SCR) Selective non-catalytic reduction (SNCR) Absorption Condensation Post-combustion (thermal postcombustion, catalytic postcombustion) Adsorption processes (activated carbon, . . .) Absorption processes (poly(ethylene glycol) dimethyl ether, . . .) Adsorption processes without regeneration Piniost-combustion under certain conditions (temperature, residence time)

Residues Dry dust Dry dust Dry dust Wastewater and sludge separated from the wastewater Dry dust containing sulfites, sulfates, and unreacted sorbent Gypsum, wastewatera SO2 gas (concentrated)b

Sodium sulfate, ammonium sulfate, wastewatera Elementary sulfur and/or CS2 (concentrated) Sulfuric acid or sodium sulfate Dry dust containing chlorides, fluorides, and unreacted sorbent Wastewaterc N2 and H2O in the off-gas N2 and H2O in the off-gas Aqueous ammonia solution Liquid hydrocarbons CO2 and H2O in the off-gas

Hydrocarbons (concentrated) Hydrocarbons (concentrated) Contaminated sorbent CO2, H2O, and HCl in the off-gas

Can be operated without wastewater under certain conditions Can be used for production of sulfuric acid, elementary sulfur, or liquid SO2 c Under certain circumstances hydrochloric acid can be produced

b

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dry de-dusting and gas cleaning processes where an adsorbent or a reactant is added to the off-gas which allows for a simultaneous separation of dust and gaseous pollutants in the subsequent de-dusting system. In this case the resulting residue is a mixture of the original dust, the adsorbent, or the reaction products and some unused reactant. In the case of a wet de-dusting process the insoluble fraction of the dust becomes suspended in the scrubber water and water-soluble components are dissolved. The suspended solids are usually separated from the scrubber water by sedimentation or filtration. The water content of the resulting sludge or filter cake depends on the characteristics of the solids, especially the particle size distribution, and on the separation process itself. Most of the scrubber water is reused in the scrubber. However, some discharge of scrubber water is usually required in order to control the concentration of dissolved ions (chlorides, sodium, potassium, . . .). Treatment of this wastewater before discharge usually generates additional solid residues.

Desulfurization The residues from separation of sulfur dioxide and sulfur trioxide from off-gas can vary substantially depending on the type of desulfurization process applied (Table 1). In the selection of a desulfurization process the off-gas parameters are important, especially when a sellable product shall be produced. Semidry and dry chemisorption processes produce a residue which is a mixture of sulfites, sulfates, and unreacted sorbent. In contrast, aqueous sorption processes usually deliver sulfate because sulfite can be oxidized more easily in the liquid phase. The resulting products depend on the base added to the scrubber water. When lime-based sorbets are used (limestone FGD) gypsum is precipitated. In the Walther process ammonia is added to the scrubber water. The ammonium sulfate is recovered by evaporation of the water. In the case of sodium hydroxide, wastewater rich in sodium sulfate is discharged. In regenerative sorption processes, for example the Wellman-Lord process (Neumann 1991) or the active coke process (Richter 1991) concentrated sulfur dioxide gas is produced in the regeneration step. This can be processed further to elemental sulfur (Claus process) and sulfuric acid or it is compressed to liquid sulfur dioxide (Elsner et al. 2003). Hydrogen sulfide and carbon disulfide can be separated from gas streams by adsorption on activated carbon or by absorption. In the regeneration step concentrated carbon disulfide can be produced. Regenerated hydrogen sulfide from adsorption is usually converted to elemental sulfur, while in absorption processes using hydrogen peroxide or sodium hydroxide sulfuric acid or sodium sulfate is produced (Schultes 1996). Separation of Hydrogen Chloride and Hydrogen Fluoride In scrubbers, hydrogen chloride and hydrogen fluoride can be separated using water as sorbent. Some of the scrubber water has to be discharged after neutralization. To avoid high loads of salts in the discharged water the fluoride concentration can be limited by precipitation of calcium fluoride using calcium hydroxide. Precipitation of chlorides is not feasible on a technical scale. In the case of high concentrations of hydrogen chloride, the production of a technical grade hydrochloric acid can be considered (Joint Research Centre 2016a).

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Reduction of Nitrogen Oxide Emissions Nitrogen oxides are usually removed from off-gas by selective catalytic reduction (SCR) or selective non-catalytic reduction (SNCR). In these processes the nitrogen oxides react with ammonia or urea to form nitrogen and water vapor. The reaction products are released to the atmosphere together with the off-gas. Reduction of Ammonia Emissions Ammonia can be removed from off-gas by absorption in water or acids. When no other pollutants are contained in the gas and deionized water is used as absorbent, aqueous ammonia can be produced. Reduction of the Emissions of Volatile Organic Compounds The most common method to reduce the emission of volatile organic compounds (VOCs) is the oxidation of the components to carbon dioxide and water vapor. Both reaction products are released to the atmosphere together with the off-gas. Various designs for such a post-combustion process have been realized. When valuable VOCs are present in a gas stream at higher concentration, condensation of these VOCs at low temperature can be a feasible process. In this case the residue from the gas cleaning process is a liquid mixture of organic components. Another process for recovery of VOCs is adsorption on activated carbon. Alternatively, VOCs can be absorbed by an organic fluid having a low vapor pressure. In the regeneration step concentrated hydrocarbons can be recovered.

Residues from Air Pollution Control in Major Industry Sectors Cement Industry APC in cement production is usually performed by de-dusting with ESPs or fabric filters. The amount of dust separated from the off-gas of a cement kiln (cement kiln dust, CKD) depends on the mode of operation of the plant: compound operation or direct operation. In compound operation the amount of filter dust is in the range of 50–150 kg/t clinker produced, while in direct operation it is 80–200 kg/t clinker (Schorcht et al. 2013). The dust is characterized by a composition similar to the cement clinker with increased fractions of carbonate, chloride, and sulfate. In the kiln system an internal circulation between the kiln and the preheater acts as an enrichment cycle for volatile compounds. An increased input of chlorine, sulfur, and alkalis with the raw materials or fuels leads to a high circulation of these components. This enforces the use of a gas bypass at the kiln inlet. By discharging part of the process gas, chlorine, sulfur and alkalis are removed from the cycle. Typical bypass rates are up to 15% for chlorine bypass and up to 70% for a sulfur bypass. De-dusting of the bypass gas produces the bypass dust which is characterized by an increased content of alkali chlorides or sulfates. Iron and Steel Industry For the production of steel from iron ore in an integrated steel mill several emissionintensive process steps are required.

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In a first step sinter is produced from fine-grained iron ore in the sinter plant. APC in these plants is usually performed by de-dusting with ESPs or fabric filters. The amount of dust separated is in the range of 0.2–3.6 kg/t sinter (Remus et al. 2013). In some sinter plants additionally a dry off-gas desulfurization process is applied for gas cleaning. In the blast furnace (BF) pig iron (hot metal) is produced from sinter, iron ore pellets, and coke. The top-gas from the BF is usually de-dusted in two steps. In the first stage, a dust-catcher or cyclone, the coarse dust is separated (BF dust) at a rate of 3.4–18 kg/t hot metal (Remus et al. 2013). In the second stage for the separation of the fine dust a scrubber is usually applied. Most of the scrubber water is recirculated after sedimentation of the particles which are discharged as sludge. The amount of BF sludge is in the range of 2.0–22 kg/t hot metal (Remus et al. 2013). Some water has to be discharged from the scrubber system to limit the concentration of dissolved ions (chloride, . . .). The amount of wastewater is in the range of 0.1–13.7 m3/t hot metal. In some BFs, especially in China, second-stage de-dusting is performed by a fabric filter or an ESP. In this case the residue is a fine-grained BF dust. The cleaned BF gas is utilized as a fuel. During tapping of the hot metal a considerable amount of dust is produced in the cast house. In most BFs the cast house ventilation air is de-dusted by an ESP or a fabric filter. The separated amount of cast house dust is in the range of 0.6–5.1 kg/t hot metal (Remus et al. 2013). In a further process step, the basic oxygen furnace (BOF), the hot metal is converted into steel. The BOF off-gas can be de-dusted in a wet process (scrubber) or a dry process (ESP). Subsequently, the clean BOF gas is used as a fuel. The amount of solid residue remaining from de-dusting is 0.75–24 kg/t liquid steel (Remus et al. 2013). In the case of a wet de-dusting system the amount of wastewater is up to 6 m3/t liquid steel (Remus et al. 2013). In the coke oven plant coal is converted to coke which is required in the BF process. The off-gas from the coke ovens (coke oven gas, COG) is first cooled. In this step high-boiling hydrocarbons are condensed and discharged as tar (26–48 kg/t coke (Remus et al. 2013)). In a series of different downstream absorbers hydrogen sulfide, ammonia, and light hydrocarbons are separated from the coke oven gas. The light oil is recovered in the regeneration of the absorbent (9.1–14 kg/t coke (Remus et al. 2013)). From the separated hydrogen sulfide and ammonia various products can be produced (sulfuric acid, elemental sulfur, anhydrous ammonia, ammonium sulfate) depending on the applied treatment processes. Finally, the clean COG is used as a fuel. Steel can be produced in a single step using iron scrap in an electric arc furnace (EAF). The off-gas from EAF steelmaking is usually de-dusted in a fabric filter. The amount of EAF dust is in the range of 10–30 kg/t of liquid steel (Remus et al. 2013).

Combustion and Incineration Plants In the EU15 coal-fired power plants generated 48 million tons of combustion residues in 2010. Most of these residues resulted from flue gas cleaning: 65% was fly ash separated in de-dusting systems and 21% was gypsum from flue gas desulfurization (Joint Research Centre 2016b).

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The ash content of heavy fuel oil is usually well below 0.2% while the sulfur content is between 0.5 and 2.2% (Joint Research Centre 2016b). Thus, the overwhelming share of combustion residues in oil combustion results from off-gas desulfurization. In biomass combustion plants various types of biomass (peat, wood, straw, . . .) are used as fuel. The amount of fly ash separated in the de-dusting system depends on the ash content of the used fuel and the type of combustion system. In grate-fired combustion the fraction of fly ash is typically 30% of the total ash (Narodolawsky and Obernberger 1996), while in fluidized-bed combustion this fraction is higher than 80% (Joint Research Centre 2016b). Due to the potentially high emission levels of waste incineration plants waste gas cleaning is extensive in these plants. Thus various residues are generated depending on the type of waste gas cleaning system. Typical amounts of APC residues in municipal solid waste incinerators (MSWI) are in the case of a wet gas cleaning system 20–40 kg boiler dust and 15–40 kg fly ash per ton of waste. In semidry and dry gas cleaning systems the amount of fly ash is somewhat higher (20–50 kg/t waste) (Joint Research Centre 2016a).

Oil Refineries In oil refineries the main residue from APC processes is sulfur components from the regeneration of absorptive desulfurization processes. These components are usually converted into elemental sulfur in a Claus process. The sulfur recovery in modern refineries is higher than 99%. Thus, the amount of residue depends mainly on the sulfur input with the crude oil (Barthe et al. 2015).

Enrichment of Volatile Elements in the Off-Gas In the off-gas from processes operated at high temperatures like combustion processes or melting processes volatile elements are enriched. This leads to an enrichment of these elements in the residues from off-gas cleaning. When the residues from APC are recycled back into the process where they have been generated the volatile components are usually cycled up in the process (Table 2). The volatility of elements depends on the temperature in the process but also on the presence or absence of other components like oxygen and chlorine. Important volatile components with respect to up-cycling in APC systems are alkali chlorides (potassium chloride, sodium chloride), some heavy metals (lead, zinc, cadmium, . . .), and sulfur components. Components volatilized in a process at high temperature usually recondense on dust particles when the temperature of the off-gas decreases. As the specific surface area of particles depends on the particle size also the concentration of the recondensed material usually varies with the particle size. This dependence can be described approximately by the relation c ~ 1/dN (Linak and Wendt 1993). The exponent N results from the mechanism of condensation or reaction of the components on the particles. When the concentration of a component has been measured in different size fractions of a dust the exponent can be obtained by linear regression.

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Table 2 Important volatile elements and components in some industrial high-temperature processes KCl, NaCl

Process plants Iron ore sinter plant Cement kiln

Zn

As, Cd, Pb, Zn, . . . S

Biomass combustion plants using fuels like wheat straw or Miscanthus Steel mill (electric arc furnace) Integrated steel mill (basic oxygen furnace) Integrated steel mill (blast furnace) Copper smelter Biomass combustion plants

Iron ore sinter plant

Residue ESP dust By-pass dust Fly ash

Main problems caused Separation efficiency of ESP

EAF dust BOF dust BF dust

Energy demand for Zn reduction and volatilization Energy demand for Zn reduction and volatilization Operation problems of furnace

Dust Fly ash

Operation problems of furnace Concentration above limit concentration makes use as soil conditioner impossible Reemission of sulfur components

Dry DeSOx residue

Kiln operation problems Ash melting behavior

For several processes significantly increased concentrations of volatile components were found in the finest size fractions of the dusts. This effect is used in some APC systems consisting of more than one separation stage. In the first de-dusting stage, e.g., a cyclone, the coarse dust is separated, while in the second stage, for example an ESP, the fine dust is separated. As a consequence, the concentration of the volatilized and recondensed components is lower in the dust from the first de-dusting stage. Two- or three-stage du-dusting is applied, e.g., in BFs, for steelmaking off-gas or for the off-gas from biomass combustion plants. Alternatively, classification of the separated dust can be applied to split a separated dust into two fractions: a fine fraction with increased concentration of the volatile components and a coarse fraction with lower concentrations.

Ways of Utilization of Air Pollution Control Residues Use as Products General Some APC processes are designed in a way that the remaining residue is a product that can be sold. Some additional processing of the residue is often required to achieve this goal.

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Gypsum In the limestone FGD process the final product is gypsum. The sulfur dioxide is absorbed in the scrubber suspension which contains limestone and some dissolved oxygen. The sum equation of the reaction is: SO2 þ CaCO3 þ ½ O2 þ 2 H2 O ! CaSO4 :2 H2 O þ CO2 Small gypsum crystals are also contained in the suspension as crystallization nuclei for the produced gypsum. Classification in a hydro-cyclone separates the coarse crystals for discharge, while the small crystals remain in the recirculated suspension. The coarse gypsum crystals are further dewatered and sold as granular material or briquetted after agglomeration. The gypsum is used in the building material industry, for example for wallboard production, as plaster or as set retarder in Portland cement. It can also be used in agriculture for acid soil amelioration or in mining applications (Poullikkas 2015).

Sulfur Dioxide, Elemental Sulfur, and Sulfuric Acid In regenerative sorption processes for off-gas desulfurization, for example in the Wellman-Lord process, the active coke process or the Solinox process, concentrated sulfur dioxide gas is produced in the regeneration of the sorbent (Poullikkas 2015). The sulfur dioxide can be compressed and liquefied. Alternatively, it can be converted into elementary sulfur in a Claus process (Elsner et al. 2003): SO2 þ 2 H2 S ! 3=2 S2 þ 2H2 O The required hydrogen sulfide is produced from part of the sulfur dioxide by reduction with hydrogen. Another option is the production of sulfuric acid. In regenerative adsorption processes for the adsorption of hydrogen sulfide the concentrated hydrogen sulfide gas obtained in the regeneration step is also converted in a Claus process to elementary sulfur. The required sulfur dioxide is produced by oxidation of part of the hydrogen sulfide.

Hydrochloric Acid When large amounts of hydrochloric acid are separated from an off-gas by absorption in an aqueous solution, for example in the APC system of a waste incineration plant, technical grade hydrochloric acid can be produced (Menke et al. 2003). When the hydrochloric acid concentration in the off-gas is not extremely high only diluted acid can be produced in the absorption process. Thus, a rectification step is required. Simple rectification is only possible starting at a hydrochloric acid concentration from the absorption process above the azeotropic point (20.2%w/ w). By adding additives (calcium chloride, magnesium chloride, . . .) to the diluted hydrochloric acid, the phase equilibrium can be shifted in a manner that the azeotropic point occurs at lower concentrations or disappears totally (Schultes 1996).

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Volatile Organic Compounds Volatile organic compounds separated from the off-gas by condensation, by regenerative absorption, or by adsorption processes (Gupta and Verma 2002) can be used in appropriate applications. CKD as Binder CKD can be used directly to some extent by blending it with the cement produced (Schorcht et al. 2013). The partial replacement of cement in concrete by CKD has been investigated in numerous studies. The strength of the cured concrete decreased somewhat (Shoaib et al. 2000) or even slightly improved (Maslehuddin et al. 2008). Another possibility for utilization of CKD is the addition to lime mortar (Siddique 2006). In soil stabilization CKD can be used as an alternative to lime. In this case the conventional compaction 2 days after the introduction is not necessary (Siddique 2006). Fly Ash as Pozzolan Fly ash from coal combustion plants is an artificial pozzolan (Ahmaruzzaman 2010). A pozzolanic reaction is the volume-neutral chemical reaction of calcium hydroxide and silicon dioxide forming calcium silicate hydrates: x CaðOHÞ2 þ y SiO2 þ z H2 O ! x CaO • y SiO2 • ðx þ zÞ H2 O Due to its pozzolanic properties fly ash contributes to structural and strength formation of concrete. In DIN EN 206 (2014) fly ash is classified as additive type II (pozzolanic additives). The fly ash may be used as a concrete additive if it meets the requirements of DIN EN 450 (2012) which are met by most fly ashes from the combustion of hard coal. Fly ash from the combustion of lignite fits to a much lesser extent (Lutze and vom Berg 2004). According to DIN EN 197 (2011) the fly ash can also be used directly in cement as a main component.

Fly Ash as Soil Ameliorant In many countries the ashes from the combustion of chemically untreated biomass are utilized as soil conditioner on agricultural land and forests. The recycling of biomass ashes to the soil is proposed to help to close the nutrient cycles for the soil where the biomass was grown. However, country-specific limit concentrations for various heavy metals in the ash (e.g., arsenic, cadmium, chromium, copper, mercury, nickel, lead, zinc) have to be observed (Lanzerstorfer 2014). Often the fly ash is stabilized before it is spread onto the soil (Steenari and Lindqvist 1997). Fly ash from the combustion of coal is also a useful ameliorant that may improve the physical, chemical, and biological properties of problem soils and is a source of readily available plant macro- and micronutrients (Jala and Goyal 2006; Basu et al. 2009).

Direct Recycling Direct recycling of residues back into the process where the cleaned off-gas originated is possible in some cases. However, very often the residues contain elements or

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components which are unwanted in the process. In such cases no recycling or only partial recycling of the untreated residues is possible. The removal of the unwanted elements or components before the recycling can be an alternative. For this treatment process, e.g., leaching, thermal treatment or classification can be applied.

Volatile Organic Compounds When volatile organic compounds, for example solvents, are separated from the off-gas by condensation these VOCs are often reused in the same process. Dust When the enrichment of unwanted elements or components is small, dusts can often be recycled, at least partially, back into the process from which they originate. Another limitation for direct recycling can be the fine grain size of a dust. An example for direct dust recycling is the use of the iron-rich ESP dust of iron ore sinter plants as feed material of the sintering process (Remus et al. 2013). In cement plants the composition of CKD is similar to cement. Thus, it can be recycled back to the kiln (Schorcht et al. 2013). However, in both cases the content of alkali salts can be a limiting factor. EAF dust can also be recycled back into the furnace. To avoid increased reemission of dust it is injected into the liquid bath (Jensen and Wolf 1997). Also dust from dry de-dusting of BOFs can be recycled by being injected into the furnace (Forsthuber and Enkner 1998).

Treatment of Residues for Recycling or Utilization In many cases the dusts have to be treated before they can be recycled. Unwanted components in dust can be removed by leaching (hydrometallurgical) processes and by thermal (pyrometallurgical) processes. Because of the fine grain size, the recycled dust often has to be agglomerated before recycling to limit immediate re-entrainment of the recycled dust.

Agglomeration of Dust Recycling of a dust back into a process can increase the rate of dust generation substantially. Because of the small grain size, the particles are re-discharged easily by the produced off-gas. In such cases agglomeration of the dust prior to recycling is required. The two main processes used for agglomeration of dust are pelletizing and press agglomeration (briquetting). Pelletizing is a wet nonpressure agglomeration method which produces round pellets. It requires the use of a binder to foster particle size enlargement. The finegrained material is fed onto a disc pelletizer or a drum pelletizer where the pellets are grown by tumbling the fines against themselves in the presence of the binder. Once the pellets have reached the desired size they exit the pelletizer. In the case of press agglomeration, external forces are applied to the dry dust by means of pressing tools. Many contacts are being formed between the particles. The cavity area is reduced, brittle primary particles can break, and plastic particles are

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deformed. In press agglomeration, binders may be added to increase the agglomerate strength. Typical binders used are molasses, water glass, or bentonite. EAF dust is often agglomerated before it is recycled to reduce reemission. Carbonaceous material can be added to the mixture to compensate for the increased energy demand of the recycled material (Bacinschi et al. 2012; Lopez and LopezDelgado 2002; Wu et al. 2017). In a similar way the sludge from BOF gas wet de-dusting can be agglomerated and recycled into the furnace (Su et al. 2004). Dust from BOF dry de-dusting can be briquetted for recycling at a material temperature of 500–700  C because of its metallic iron content. During hot briquetting sintering bridges occur between the particles (Pietsch 1994). Recently, the agglomeration of BF sludge with starch binder for recycling has been investigated (Drobíková et al. 2016). Granulation of a dust can be used to increase the dust bulk density for transport or landfill. Such granules, which are matured by intermediate storage, exhibit high mechanical stability and abrasion resistance as well as improved dust behavior (Salihoglu and Pinarli 2009).

Leaching Leaching can be applied for the extraction of some species from a residue, aiming to improve the quality of the residue for further utilization and/or to recover the species in question. Leaching of residues which are hazardous waste aims to allow disposal on nonhazardous waste landfill sites (Amutha Rani et al. 2008). Besides water, various acidic and basic leaching agents have been investigated in different applications depending on the solubility of the respective components. Leaching processes have been investigated widely for example for EAF dust (Jha et al. 2001) and MSWI fly ash (Quina et al. 2008). A reduction of the content of critical heavy metals in wood combustion fly ash can also be achieved by leaching (Lanzerstorfer and Kröppl 2018). Leaching with water can be applied to recover potassium chloride from cement kiln bypass dust (Sturm and Galichet 2012) or biomass combustion fly ash. Thermal Treatment Thermal treatment can be applied for the volatilization of some species from the dust to improve the quality of the residue for further utilization and/or to recover the volatilized components. Thermal treatment has been investigated for example for MSWI fly ash (Quina et al. 2008; Nowak et al. 2010). The treatment is capable of removing the more volatile heavy metals like cadmium and lead. Thermal treatment can be used to obtain residues with improved leaching characteristics as well as a treated material that is suitable for reuse. In a vitrification process the residues are mixed with glass precursor materials and then combined at a temperature of 1000–1500  C into a single-phase amorphous glassy product. In melting processes similar temperatures are applied but without addition of glass materials resulting in a multiple-phased product. In a sintering process the temperature is around 900  C. Bonding of particles occurs and chemical phases in the residues reconfigure (Sabbas et al. 2003). These techniques show the potential of stabilizing hazardous waste and producing reusable materials.

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Classification Classification of the sludge obtained from the wet cleaning of BF off-gas has become a common treatment procedure (Remus et al. 2013). Thereby, the sludge is separated by hydro-cyclones into two size fractions. The coarse fraction is less contaminated with zinc and therefore can be added to the feed material of the sinter plant. Residues from dry de-dusting can be classified in a similar way using a shifter or air classifier. Classification of the residue from dry BF off-gas de-dusting for zinc enrichment and depletion has been suggested (Lanzerstorfer and Kröppl 2014). In a similar way air classification could also be applied for steelmaking dust (Lanzerstorfer 2018).

Use of Residues in Another Process Iron Ore Sinter Plant In integrated steel mills the sintering process is applied to agglomerate the fine ore before it can be processed in the BF. The sinter strand is also used to recycle the finegrained iron-bearing residues generated at various process steps in steel production. Thus, a lot of the residues from off-gas de-dusting processes are also recycled via this process (Hansmann et al. 2008). However, there are some limitations for the inclusion of dusts in sintering. These are the alkali chloride content of the dust and the content of VOCs, lead, and zinc. The limit for alkali chlorides arises from the operating conditions for the ESP used for APC because high concentrations of alkali chlorides increase the specific dust resistivity above the limit for reasonable ESP separation efficiency. A high load of VOCs can cause the risk of fire in the ESP because the specific combustion conditions in the sintering process cause the volatilization of a large fraction of the VOCs in the process. The recondensed VOCs stick to the dust particles. When the content of VOCs in the separated dust is high the sparks occurring in the ESP can cause a fire. The limit for lead and zinc does not arise from the sintering process itself but from the quality requirements for the sinter produced. A high amount of lead and zinc fed to the BF causes operational problems in BF operation and the sintering process is not very effective in volatilization of zinc (Lanzerstorfer et al. 2015). Cement Kiln The correct proportions of calcium, silicon, aluminum, and iron in Portland cement are specified within rather strict limits. To provide each of the main elements of the cement also residues from APC, especially fly ash from coal combustion and some steel mill dusts are used as part of the feed (Barthe et al. 2015; Koros 2003). Pyrometallurgical Zn Recovery Dusts from steelmaking with an increased zinc content are often processed for zinc recovery. Several processes have been developed to achieve this goal but only the Waelz kiln reached widespread application. It accounts for about 75% of the total treated EAF dust globally (Lin et al. 2017). In these processes the zinc is volatilized by reduction to metallic zinc. Crude zinc oxide is recovered from the process off-gas

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which is a valuable raw material for zinc production. In the Waelz process the remaining slag is sent to landfill, while in other processes (rotary hearth furnace, the PRIMUS® process, and OXYCUP ® process) the iron-rich residue is use in the iron production step (Antrekowitsch et al. 2015; Holtzer et al. 2015). As still no optimal solution for recovery of zinc from ironmaking and steelmaking dusts is available research in this area is continuing (Lin et al. 2017; Antrekowitsch et al. 2015).

Recovery of Alumina The recovery of alumina from coal fly ash was pioneered by Grzymek in Poland. In recent years several plants for alumina recovery from coal fly ash have been installed in China (Yao et al. 2014). The feasibility of alumina recovery from fly ash mainly depends on its alumina content. Usually fly ash with an alumina content of more than 30% can be considered viable for the recovery of alumina.

New Ways of Utilization Production of Potassium Chloride for Fertilizer Potassium can be enriched in several dusts: in sinter plant dust, in cement kiln bypass dust, and in fly ash from biomass combustion. In recent years, increasing research efforts in recovery of potassium chloride for fertilizer production are evident (Maeda et al. 2017; Lanzerstorfer 2016; Peng et al. 2009; Zhan and Guo 2013; Yu et al. 2017). A first plant for production of potassium chloride from cement mill bypass dust on an industrial scale was built in 2011 (Sturm and Galichet 2012). In air classification tests with the dust from sinter plant de-dusting (Lanzerstorfer 2015a), bypass dust from a cement kiln (Lanzerstorfer 2016) and biomass combustion fly ash (Lanzerstorfer 2011, 2015b), it has been demonstrated that a fine dust fraction with a significantly higher potassium chloride concentration can be separated from the bulk. Processing dust with a higher potassium chloride content should improve the efficiency and economy of the winning process. In the case of sintering dust and bypass dust the remaining coarse fraction depleted in potassium chloride could be recycled in the respective process. Glass-Ceramics The production of glass-ceramic from various silicate-rich air pollution control residues (waste incineration fly ash, coal fly ash, EAF dust, CKD, . . .) has been investigated for a long time. Although several studies have shown promising results, widespread industrial production of useful glass-ceramics including APC residues is not yet achieved (Rawlings et al. 2006; Barbieri et al. 1999). For example, fly ashes originating from APC at MSWIs can be mixed with different amounts of inert materials such as glass cullet and feldspar waste to vitrifiable mixtures. The glasses obtained by means of the vitrification process were chemically stable showing a low leachability of contaminants. The material properties were comparable to those of commercial soda-lime glasses (Andreola et al. 2008).

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Ceramics The use of EAF dust as raw material for ceramic building products has been investigated. Use of up to 20% of EAF dust showed acceptable properties of the produced material. In leaching tests with the material produced the heavy metal concentrations were below the acceptable limits (Sikalidis and Mitrakas 2006; Machado et al. 2011). Production of Zeolites Several studies have shown that coal combustion fly ash can be used to synthesize zeolites (Querol et al. 2002). Zeolites are crystalline aluminum silicates, with group I or II elements as counterions. As a consequence of the structural properties of zeolites, they have a wide range of industrial applications mainly based on ion exchange, gas adsorption, and water adsorption. Current research is into the processing of the fly ash targets to improve zeolite yield and transformation efficiency (Yao et al. 2015).

Landfill General Disposal of material ranks lowest in the waste hierarchy. However, safe disposal of non-recoverable residues, especially hazardous material, is still an important part of APC residue management.

Backfilling in Mines Backfilling means the back stowing of mines for safety reasons. Mine cavities that have resulted from the exploitation of natural resources by mining are filled with appropriate industrial wastes. Backfilling has been approved by the European Court as a form of recycling on condition that by means of a necessary measure natural resources are being substituted by industrial wastes and resources are protected (Marx et al. 2005). For environmental protection limiting values of harmful substances within waste and the leaching behavior have to be considered. However, these limiting values are not valid for salt mines which completely enclose the harmful substances, thus permanently closing them from the biosphere. Hazardous APC residues can be dumped safely in underground salt caverns. Thereby, different methods for backfilling are applied: hydraulic backfilling and big bag backfilling. In hydraulic backfilling the fine-grained APC residues are usually processed into a flowable backfilling mixture by adding a concentrated salt solution as a transporting medium on the surface and remain in a reaction drum to reduce exothermal effects and possibly existing gas production potential. After a residence time of few hours the mixture is transported into the recoveries to be backfilled by

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means of a piping system. The excess transporting solution drains the backfilling material by means of a barrier system. It is collected in a basin at a deeper point of the mine and subsequently transported back into the process on the surface (circulation of the transporting solution). By means of hydraulic backfilling a nearly complete backfilling (>90 Vol. %) of the mining sites is possible (Marx et al. 2005). In big bag backfilling the big bags are transported into the underground via the shaft and big vehicles transport them into the backfilling chambers. Forklift trucks install the big bags into the backfilling chambers in layers. Alternating storage with covering salts is applied to reduce the pores.

Solidification Solidification is a pre-landfill waste treatment that aims to make hazardous APC safe for disposal. This process involves mixing APC into a binder system. A wide range of binders (cementitious binders, lime, bitumen) have been used for solidification. The aim is to incorporate the APC into the binder system and produce a monolithic solid with structural integrity and long-term stability. Thus, solidification should also inhibit leaching (Amutha Rani et al. 2008).

APC Residues from Waste Incineration Plants MSWI fly ash can be stabilized by mixing with cement to reduce leaching of heavy metals (Shi and Kann 2009). By leaching of the fly ash with water for the removal of soluble salts the amount of fly ash that may be incorporated into the cementitious matrices is significantly increased (Mangialardi et al. 1999). By-Product from Semidry and Dry Sorption Processes for Acid Gases When calcium-based sorbets are used the by-product from semidry and dry chemisorption processes typically consists of calcium sulfate, calcium sulfite, and calcium chloride also including carbonates and unreacted lime. The content of fly ash varies between 1 and 80% (by mass) depending on the degree of pre-de-dusting. Normally, the by-product is disposed of. However, industrial utilization may be possible in the future (Poullikkas 2015). In some plants mixed sorbents—hydrated lime and activated carbon—are used for integrated separation of acid gases and toxic organic compounds, for example PCDD/F. In such plants the by-product also contains these toxic components. EAF Dust EAF dust can be stabilized before dumping to minimize leaching of heavy metals. Stabilization with lime and/or Portland cement has been investigated (Salihoglu and Pinarli 2009). It was found that the best results were achieved with a mixture of lime and Portland cement of 1:1, since the pH of the suspension was shifted to the slightly basic range. With this additive mixture, total mixtures with 30% EAF dust could be stabilized.

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Part VI Environmental Analysis

Tailor-Made Molecular Traps for the Treatment of Environmental Samples

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Rüstem Keçili, Özlem Biçen Ünlüer, and Chaudhery Mustansar Hussain

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . MIP-Based Molecular Traps for Solid Phase Extraction of Pharmaceuticals from Environmental Samples . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Ion Imprinted Traps for Solid Phase Extraction of Metals from Environmental Samples . . . . MIP-Based Molecular Traps for Solid Phase Extraction of Dye Compounds from Environmental Samples . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

Environmental pollution is one of the serious global problems today. Many hazardous compounds in environmental samples such as heavy metals, drugs, and dye compounds phenolics, herbicides and pesticides, etc. are main sources of environmental pollution, and they may cause serious diseases in humans, animals, and plants. Thus, analysis and efficient removal of these undesired compounds from environmental samples such as wastewater and soil is a crucial task. Traditional materials such as activated carbon are commonly used for the treatment of environmental samples. However, it has some disadvantages such as regeneration problems and high cost. These challenges can be overcome by using R. Keçili (*) Yunus Emre Vocational School of Health Services, Department of Medical Services and Techniques, Anadolu University, Eskişehir, Turkey e-mail: [email protected] Ö. Biçen Ünlüer Faculty of Science, Department of Chemistry, Anadolu University, Eskişehir, Turkey C. M. Hussain Department of Chemistry and Environmental Science, New Jersey Institute of Technology, Newark, NJ, USA © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_24

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tailor-made traps called molecularly imprinted polymers (MIPs). MIPs are highly cross-linked resins having selective binding groups and cavities toward the target compound. In this chapter, we provide the recent progresses of new MIP materials for the removal of pollutants such as pharmaceuticals, metal ions, and dye compounds from environmental samples. Keywords

Molecularly imprinted polymers (MIPs) · Molecular traps · Extraction · Treatment · Environmental samples · Wastewater

Introduction Molecularly imprinted polymers (MIPs) are engineered molecular traps that possess specific recognition sites with high affinity and selectivity toward the target compound (Fig. 1). Therefore, MIPs can be used in many applications where selective binding phenomena is crucial, such as solid phase extraction, biosensors, and catalysis (Sellergren 2001; Kryscioa and Peppas 2012; Cheong et al. 2013; Vasapollo et al. 2011; Chen et al. 2012a, b; Guerreiro et al. 2011; Lai and Feng 2003; Wang and Wei 2017; Keçili et al. 2011, 2012; Kupai et al. 2017; Erdem et al. 2010; Cui et al. 2015; Prasad et al. 2010). Environmental pollution is one of the significant international concerns. Various chemicals in water such as pharmaceuticals, heavy metals, dye compounds, etc. exhibit harmful effects on human body, animals, and plants. Thus, selective recognition and extraction of these compounds from environmental samples are crucial.

Fig. 1 The principle of molecular imprinting (Reproduced with permission from Lofgreen and Ozin 2014)

Y

Functional monomer

imprint

imprint

Y Polymerization with cross-linker

Removal of target

imprint or target

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Traditional approaches such as gas chromatography (GC) and high-performance liquid chromatography (HPLC) were used for the trace analysis of these undesired chemicals in environmental matrices (Bagheri et al. 2004; Bonwick et al. 1995; Hladik et al. 2008). But these techniques are time-consuming and expensive. These drawbacks of traditional techniques can be overcome by preparation of MIP-based micro- and nanomaterials which exhibit high selectivity and affinity toward target compound. In this chapter, recent developments and environmental applications of MIP-based selective materials in micro- and nanoscale were provided.

MIP-Based Molecular Traps for Solid Phase Extraction of Pharmaceuticals from Environmental Samples Solid-phase extraction (SPE) is one of the powerful tools for sample preparation in analytical chemistry. The researchers experimentally applied this technique in the late of 1940s (Liska 2000), and the developments which lead to use of SPE in the current analytical applications were started in the early 1970s. It is now widely used in many different areas in analytical chemistry such as preconcentration, purification, fractionation, etc. Various traditional SPE resins such as silica-based (Cazes 2009; Spivakov et al. 2006), carbon-based (Pyrzynska 2007; Rodríguez et al. 2015), and clay-based (Valdes et al. 2006) materials have been used in different SPE applications. Most of these SPE resins are now commercially available as SPE tube and pipet tip format such as Oasis-HLB (manufactured by Waters) and Omix (manufactured by Agilent), MonoTips (manufactured by GL Sciences). Even though SPE is a popular and widely used analytical tool for sample cleanup and preconcentration of target compounds from complex samples, the traditional materials used for SPE applications show lower selectivity toward desired analytes which leads to binding of interfering compounds in the sample matrix. Molecularly imprinted polymers (MIPs) as promising SPE resins which display higher selectivity and affinity toward target compound can overcome these disadvantages. MIPs are selective for a target compound or group of compounds and allow preconcentration of these target compound(s) in the presence of other interfering compounds in the sample matrix. Another superiority of the MIPs compared to conventional SPE resins is their stability under extreme conditions. Compared to various SPE materials, MIPs are stable under harsher chemical conditions such as high temperature and higher and lower pH values. MIPs are widely used as SPE materials in many different samples such as environmental samples, biological samples, and food samples. The interaction and binding between the target compound in the sample matrix and the MIP can be driven by various chemical interactions such as polar, covalent, non-covalent, and hydrophobic interactions. A typical SPE protocol using MIP resins composed of four different steps as briefly explained in the following:

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1. Conditioning step: The MIP-based SPE material in the column needs to be washed with an appropriate solvent to remove undesired impurities, and then it is conditioned with the loading solvent. The specific binding sites of the polymer should be activated by performing conditioning step to maximize the interactions with the target compound. It should be noted that the polymer should not be dried during conditioning step as this may lead to recovery of the target compound in lower values. 2. Sample loading step: Sample loading is carried out after the conditioning step. In this step, the sample is loaded onto the SPE column filled with MIP to interact target compound in the matrix with the polymer. The matrix medium affects the binding properties of the target compound to the MIP. The binding (imprinted) sites of the MIP usually interact with the target compound by H-bonding, ionic interactions, or π-π interactions. Thus, solvents that have low polarity are used in the sample loading step to obtain a selective binding of the target compound to the polymer. 3. Washing step: After sample loading, the washing step is applied. When water is used as the medium for target compound, other undesired compounds in the sample matrix are non-specifically bound to the MIP by hydrophobic interactions. Therefore, a washing step using an organic solvent should be performed to disrupt undesired non-specific hydrophobic bonds. It also is crucial to dry the polymer in the column between the loading and washing step in order to avoid miscibility problems when changing from aqueous medium to organic medium. On the other hand, the binding ability of the target compound to the MIP may be effected during the washing step. This step is more crucial in SPE using MIPs than in SPE using traditional materials. The main aim of the washing step is to increase the specific interactions between the target compound in the sample and the prepared MIP in SPE cartridge and to elute other undesired interfering compounds in the sample. In general, an organic solvent that has low polarity such as DCM and CHCl3 is used as washing solvent. 4. Elution step: The last step of a typical SPE is elution step which is sometimes composed of one or more steps, where the specific interactions between target compound and the MIP are disrupted. MeCN or MeOH is commonly used as the elution solvent in MIP-based SPE. During the elution step, small volumes of elution solvent are used to get high enrichment factors. In some cases, the specific interactions between target compound and the MIP are very strong. Therefore, a weak acid such as acetic acid or a weak base such as trimethylamine should be used to get high recovery values. The group of Sellergren reported the first use of MIPs in SPE applications (Sellergren 1994). In this study reported in 1994, MIP for selective binding and extraction of pentamidine which is a pharmaceutical compound used for the AIDS patients was prepared and successfully applied for SPE of target compound pentamidine. After this work, a lot of effort was put into the design and preparation of novel MIPs toward different compounds for selective SPE applications in different areas (Caro et al. 2003; Chapuis et al. 2004; Lai et al. 2004; Brüggemann et al. 2004;

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Puoci et al. 2012; Xin et al. 2013; Li et al. 2014; He et al. 2016a; Han et al. 2015; Morais et al. 2017; Vasconcelos and Fernandes 2017; Ansari and Karimi 2017). Terzopoulou and co-workers developed molecularly imprinted fiber-based materials for selective extraction of abacavir which is an antiviral compound (Terzopoulou et al. 2016). In this work, free radical polymerization of the functional monomer acrylic acid (AA) and ethylene glycol dimethacrylate (EGDMA) was carried out in the presence of the template abacavir. After removal of the template from polymeric structure by using soxhlet extraction with MeOH and EtOH, the obtained abacavir imprinted fibers were used for the extraction of abacavir in biological and environmental samples. The results indicated that the prepared molecularly imprinted fibers displayed high affinity and selectivity toward the target compound abacavir and the maximum binding capacity of the imprinted fibers was found to be as 149 mgg1 at pH 8.0. Magnetic nanoparticles have drawn significant interest in separation science due to their advantages such as low cost, low toxicity, high chemical stability, and excellent recovery capability from complex sample matrices. Combination of magnetic nanoparticles with MIPs increases selectivity and affinity toward target compound and provides a new approach for separation applications. In a study published by Chen et al. (2015), magnetic MIPs for the selective SPE of antibacterial sulfadiazine drug were prepared. The functional monomer acrylamide (AAm) and the cross-linker EGDMA were used for the preparation of the magnetic MIPs toward sulfadiazine. The sulfadiazine binding to the prepared magnetic MIPs was carried out in the presence of other competing compounds such as sulfacetamide, sulfamethazine, and sulfamethoxazole. The results obtained from these experiments confirmed that molecularly imprinted magnetic beads showed high selectivity toward the target drug sulfadiazine. The highest binding of sulfadiazine to the MIP beads was achieved within 7 min. The maximum binding capacity of the magnetic MIPs was 775 μgg1. In another reported study, MIP beads for selective SPE of enrofloxacin which is an antimicrobial veterinary drug were prepared (Benito-Peña et al. 2015). For this purpose, methacrylic acid (MAA) and 2-hydroxyethyl metharcrylate (HEMA) were used as the functional monomer and hydrophilic comonomer, respectively (Fig. 2). The prepared molecularly imprinted spherical particles enrofloxacin were applied for the selective extraction of enrofloxacin. The outcomes from the SPE experiments showed that enrofloxacin-selective MIPs exhibited high affinity and selectivity toward the target compound with a Ka value of 7558 M1. Kyzas and his colleagues prepared MIPs for selective extraction of metformin from wastewater (Kyzas et al. 2015). In their study, functional monomer AA and cross-linker EGDMA were used for the preparation of metformin selective-MIPs. The extraction performance of the prepared MIPs toward the target drug metformin was investigated under different conditions including medium pH, interaction time, and initial drug concentration. Figure 3 shows the schematic depiction of the prepared MIPs and extraction procedure. The results showed that the prepared MIPs exhibited excellent affinity and selectivity toward metformin with a binding capacity of 80 mgg1.

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O O OH O F

O −

O

N

H

O− O

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N+ O− O

Fig. 2 Molecularly imprinted beads toward enrofloxacin (Reproduced with permission from Benito-Peña et al. 2015)

Cavity

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Medium affinity site

Drug

Imprinting cavity MIPs after extraction

MIPs after sorption

Dispersion of MIP Nanoparticles To aqueous solution

MIPs Sorption of Analytes

Separation of phases

Removal of aqueous

Stirring

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phase

Elution with Organic Solvent Ultra sonication

Centrifugation Upper phase

Introduction to LC-MS

Fig. 3 Schematic depiction of the prepared MIPs and extraction procedure (Reproduced with permission from Kyzas et al. 2015)

Another interesting work was reported by Tan et al. (2013a) on the preparation of MIP nanoparticles for the removal of ofloxacin from aqueous solutions. For this purpose, surface of the mesoporous carbon nanoparticles was covalently grafted with MIP nanoparticles. MIP nanoparticles were prepared by using MAA and TRIM as functional monomer and cross-linker, respectively. The prepared MIP-based nanocomposite displayed high affinity and selectivity toward ofloxacin in the presence of other fluoroquinolone antibiotics such as balofloxacin, sarafloxacin, norfloxacin, gatifloxacin, and enrofloxacin. The maximum binding capacity of the prepared nanocomposite was 40.98 mgg1.

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In another research reported by Olcer et al. (2017), a MIP-based SPE adsorbent was prepared for the extraction of ibuprofen (IBU) from drinking and tap water samples. For this purpose, selective MIPs toward ibuprofen were prepared by polymerization of the functional monomer MAA and the cross-linker TRIM in the presence of target drug ibuprofen. The schematic representation of the preparation of MIP toward IBU is shown in Fig. 4. The prepared MIPs exhibited high affinity and selectivity toward IBU in the presence of ketoprofen and naproxen. The developed method was also validated, and the obtained recoveries were 97.2% and 97.7% for drinking and tap water, respectively. Li et al. reported the use of molecularly imprinted magnetic Fe3O4@SiO2 particles for the recognition and extraction of tadalafil from pharmaceutical samples (Li et al. 2011). Different functional monomers AA, MAA, and 2-(trifluoromethyl) acrylic acid (TFMAA) were tested for the preparation of three different MIPs and the binding energy results indicated that TFMAA exhibited strongest interaction with the target compound tadalafil. Figure 5 shows the SEM and TEM images of the

Fig. 4 TEM (small image) and SEM images of (a) Fe3O4 particles, (b) Fe3O4/SiO2 particles, and (c) Fe3O4/SiO2 particles coated with MIP film toward tadalafil (Reproduced with permission from Li et al. 2011)

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N

NH N

N

N

60

O

N

O

O

O N

80

NH

100

O O

N S

O

z

O O

O

S

120

O

N

z

N

O

140

NH

40

N

Extraction amounts (mmol g-1)

160

20 0 4

S-MIPs

S-NIPs

V-MIPs

2 V-NIPs

mo l L -1 )

T-NIPs

(m

T-MIPs

Fig. 5 Selectivity of magnetic MIP nanoparticles toward tadalafil (Reproduced with permission from Li et al. 2011)

synthesized Fe3O4 nanoparticles, Fe3O4@SiO2 nanoparticles and Fe3O4@SiO2 nanoparticles coated with MIP film toward tadalafil. The prepared molecularly imprinted magnetic Fe3O4@SiO2 nanoparticles were successfully applied for the selective extraction of tadalafil (T) in the presence of sildenafil (S) and vardenafil (V). The prepared magnetic Fe3O4@SiO2 nanoparticles coated with MIP film displayed high selectivity toward target compound tadalafil with an extraction capacity of 140 μmolg1 (Fig. 5). The control non-imprinted nanoparticles exhibited lower binding behavior toward all compounds. The group of Xu reported the preparation of molecularly imprinted magnetic cellulose microbeads for selective extraction of artesunate (Huang et al. 2016). For this purpose, MIP shell was prepared on the surface of magnetic Fe3O4-cellulose composite beads by using surface imprinting approach. The obtained results showed that the prepared molecularly imprinted magnetic cellulose microbeads exhibited high affinity toward artesunate with a maximum binding capacity of 220 mgg1. The highest artesunate extraction was achieved in 10 h.

Ion Imprinted Traps for Solid Phase Extraction of Metals from Environmental Samples Ion imprinted polymers (IIPs) as selective traps toward target ions were first introduced by Nishide and co-workers in 1976 (Nishide et al. 1976). The process for the preparation IIPs involves the following steps:

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• Complex formation between the functional monomer and target ion • Polymerization in the presence of cross-linker • Removal of the target ion from polymeric structure The obtained IIP has binding cavities that exhibit high affinity and selectivity toward target ion. In recent years, various IIPs were used for the extraction of metals from environmental matrices. For example, Zhang and his colleagues have prepared IIP-based magnetic nanocomposites for Pb2+ extraction from environmental samples (Zhang et al. 2011). 3-(2-Aminoethylamino)propyltrimethoxysilane (AAPTS) as the functional monomer, tetraethylorthosilicate (TEOS) as the cross-linker, and template ion Pb2+ were used for the preparation of nanocomposites toward Pb2+ ions. Figure 6 shows the schematic depiction of the prepared ion imprinted magnetic nanocomposites for Pb2+ ions. Different factors which effect the selective extraction of target ion such as pH and sample volume were studied. The results obtained from SPE studies for Pb2+ in the presence of other potential interfering ions such as Zn2+, Cd2+, and Hg2+ in environmental samples showed that the prepared MIP-based magnetic nanocomposites display high selectivity toward Pb2+ ions. The binding capacity of the nanocomposites for Pb2+ ions was determined as 19.61 mgg1. A clay-IIP nanocomposite for selective extraction of Fe3+ ions from aqueous solutions was prepared by Karabörk et al. (2008). In their study, the intercalation of quartamine cations was performed by an ion exchange process between the smectite host and quartamine. Fe3+ ions were complexed with N-methacryloylamido antipyrine (MAAP) as the functional monomer. The results obtained from SPE

TEOS

AAPTS, Pb(II)

NH3H2O

TEOS, NH3H2O

Adsorption Desorption

Fe3O4@SiO2

Fe3O4

HN

Desorption

N H

Adsorption

H2N

NH2 Pb2+

H2N

Fe3O4@SiO2@IIP

NH

Template

NH2 H N

Imprinted sites

Fig. 6 Schematic depiction of ion imprinted magnetic nanocomposites for Pb2+ (Reproduced with permission from Zhang et al. 2011)

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experiments showed that the prepared clay-MIP nanocomposite showed high selectivity toward Fe3+ ions. The binding capacity of the clay-MIP nanocomposite was found to be 78.5 mgg1. In another interesting study, Nd3+ imprinted polymers were prepared by Dolak et al. (2015). N-methacryloylamido folic acid (MAFol) and EGDMA were used as the functional monomer and cross-linker, respectively. The effects of various factors such as medium pH, initial metal concentration, and interaction time on Nd3+ binding to the prepared ion imprinted polymers were investigated. The equilibrium time for Nd3+ binding was obtained in 30 min. The prepared ion imprinted polymers showed high selectivity toward Nd3+ ions in the presence of other interfering ions such as Ce3+, La3+, and Eu3+, and maximum capacity of the prepared ion imprinted polymers was 14.6 mgg1. Li and co-workers prepared IIP-based carbon nanotube/chitosan composite for the selective extraction of Gd3+ from aqueous solutions (Li et al. 2015). In their study, magnetic silica nanoparticles (SiO2@Fe3O4) and Gd3+ imprinted carbon nanotube-chitosan composite were dispersed in aqueous solution of rare earth metals. The schematic depiction of the extraction process using magnetized-Gd3+ imprinted nanocomposite is given in Fig. 7. The obtained results showed that selective extraction of Gd3+ ions in the presence of other ions such as Ce3+ and La3+ were successfully achieved. The binding capacity IIP-based carbon nanotubechitosan composite toward Gd3+ ions was found to be 88 mgg1. Another IIP toward Lu3+ was prepared by Lai et al. (2012). For this purpose, 4-vinylpyridine-acetylacetone- Lu3+ and EGDMA were used as the complex functional monomer and cross-linker, respectively. The obtained results showed that the highest Lu3+ binding to the prepared IIP was achieved at pH 5.5 within 30 min and the maximum binding capacity was found as 64.2 mgg1. A novel core-shell IIPs were prepared by Liu and colleagues for the selective removal of Co2+ ions from aqueous solutions (Liu et al. 2015a). For this purpose, 3-aminopropyltriethoxysilane (APTS) and acryloyl chloride (AC) were used for the modification of silica particles, and photopolymerization was then carried out using acrylamide (AM) as the functional monomer in the presence of template ion Co2+

(SiO2@Fe3O4)

(Carbon nanotube)

Gd3+-imprinted site

: Gd3+ : Ce3+ : La3+

Fig. 7 Schematic depiction of the extraction process using Gd3+ imprinted magnetic nanocomposite (Reproduced with permission from Li et al. 2015)

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Fig. 8 Schematic depiction of the preparation of silica-based IIPs toward Co (II) ions (Reproduced with permission from Liu et al. 2015a)

(Fig. 8). The prepared core-shell IIP showed high affinity and selectivity toward Co2+ ions. The maximum Co2+ binding was achieved at pH 6.0. Yin et al. prepared thiol-modified magnetic IIPs for selective extraction of Ag+ ions from wastewater (Yin et al. 2017). The prepared magnetic IIPs exhibited high affinity and selectivity toward target Ag+ ions in the presence of other competing ions such as Li+, Ni2+, Cu2+, and Co2+. The maximum binding capacity of IIPs toward Ag+ was found as 35.47 mgg1. In another study reported by Jiang and Kim (2014), Pt4+ imprinted polymer was prepared by copolymerization of styrene and divnylbenzene in the presence of functional monomers 4-vinylpyridine (4-VP) and dimethylglyoxime. The prepared IIP displayed excellent binding selectivity toward Pt4+ ions in the presence of Ni2+, Cu2+, and Pd2+. The highest Pt4+ binding was achieved within 40 min, and the maximum binding capacity of IIP was 38.89 mgg1.

MIP-Based Molecular Traps for Solid Phase Extraction of Dye Compounds from Environmental Samples Textile, paper, rubber, food, and cosmetic industries commonly use dye compounds (Hunger 2003; Kusic et al. 2013), and these compounds may lead to generation of large volumes of contaminated effluents. The release of contaminated effluents into the environment is a main source of environmental pollution that is a serious problem all over the world. These contaminants must efficiently be removed from the wastes which is a challenging task. Therefore, facile, efficient, low-cost, and environmentally friendly methods are needed for this purpose. Chemical and biological approaches are used for the removal of industrial dyes from contaminated effluents. In biological processes, microorganisms such as fungi,

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bacteria, algae, etc. have been used (Tan et al. 2013b; Gül 2013; Kelewou et al. 2014). Although this technique is low-cost and environmentally friendly, it is challenging and not efficient. On the other hand, conventional materials such as activated carbon are widely applied for the removal of dye compounds from environmental samples (Hazzaa and Hussein 2015; Aguiar et al. 2016; Silva et al. 2016; Regti et al. 2017) due to its high adsorption capacity. However, activated carbon has some drawbacks such as high-cost and regeneration problems. MIP-based molecular traps can overcome these drawbacks. Recently, many studies have been reported in the literature on the design and preparation of MIP-based molecular traps for the removal of dye compounds from environmental samples. In one of these studies, Kyzas et al. (2013) prepared chitosan and cyclodextrin (CD)-based MIPs for the selective removal of Remazol Red 3BS (RR) dye from aqueous solutions. The preparation of CD-based MIP toward RR schematically depicted in Fig. 9. The effect of pH, time, and initial dye concentration on binding of dye were also studied to determine the best conditions. The obtained results showed that CD-based MIP exhibited higher binding behavior toward target dye compound than chitosan-based MIPs. The maximum binding capacity was found to be as 35 mgg1. The selectivity experiments were also carried out and the results confirmed that the prepared MIPs showed great selectivity toward target dye RR in the

O

O

CI SO3Na

NaO S

N

S

N

O

O N

OH HN

N

N

N H

O

O

TDI

SO3Na

NaO S O

Dye

O O S ONa

+

β-CD Dye extraction

Dye rebinding Desorption

Fig. 9 The preparation of CD-based MIP toward RR (Reproduced with permission from Kyzas et al. 2013)

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presence of other reactive dyes such as Remazol Yellow Gelb 3RS and Remazol Brilliant Blue RN. In another study, Li et al. prepared magnetic MIPs for the extraction of Rhodamine B (RhB) dye from aqueous solutions (Li et al. 2017). For this purpose, magnetic silica nanoparticles (Fe3O4@SiO2) were coated with MIP shell having fluorophore nitrobenzoxadiazole (NBD). The results showed that a clear change in fluorescent intensity was observed through fluorescence resonance energy transfer (FRET) between the functional monomer and RhB dye (Fig. 10). Therefore, the prepared MIP-coated magnetic silica nanoparticles also behave as a fluorescent nanosensor toward the target compound RhB. The prepared MIP-coated magnetic

Fig. 10 The schematic representation of the preparation of magnetic MIPs toward Rhodamine B (Reproduced with permission from Li et al. 2017)

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silica nanoparticles were successfully applied for the recognition of RhB in aqueous solutions. The maximum RhB binding was achieved in within 60 min, and the maximum binding capacity of the magnetic nanoparticles was found to be as 29.64 mgg1. Al-Degs and co-workers prepared MIPs toward Cibacron reactive red (CRR) dye (Al-Degs et al. 2009). MAA and methyl methacrylate (MMA) were used as functional monomers for the preparation of CRR imprinted polymers (MIP 1 and MIP 2). The binding capacities of the prepared MIP 1 and MIP 2 toward CRR were 79.2% and 38.2%, respectively. Furthermore, the obtained results from selectivity experiments indicated that the prepared MIPs exhibited high selectivity toward target dye compound CRR in the presence of other reactive dyes Cibacron reactive blue and Cibacron reactive yellow. In another research published by Asman et al., MIP was prepared for the extraction of methylene blue from aqueous solutions (Asman et al. 2011). MAA and EGDMA were used as the functional monomer and cross-linker, respectively. The prepared MIP was successfully applied for the selective extraction of methylene blue from aqueous solutions in the presence of methyl orange and fast green. The maximum binding of methylene blue was achieved at pH 5.0 within 30 s. Melvin et al. developed polysulfone membrane having TiO2 nanoparticles coated with MIP and successfully used for the removal of methylene blue from aqueous solutions (Melvin Ng et al. 2017). The prepared MIP-based composite membrane showed high affinity and selectivity toward target dye compound in the presence of methylene orange with a selectivity factor of 2.0. Gao and colleagues prepared Sunset yellow-imprinted SiO2 particles using acryloyloxyethyl trimethyl ammonium chloride as the functional monomer (Gao et al. 2015). The obtained results showed that the prepared MIP-based SiO2 particles exhibited excellent affinity toward the target compound Sunset yellow dye and the maximum binding capacity was found to be 416 mgg1. The selectivity experiments were also carried out in the presence of other azo dyes such as Acid red 14 and Acid red 18. The selectivity coefficients of the MIP-based SiO2 particles for Sunset yellow/Acid red 18 and Sunset yellow/Acid red 14 binary mixtures were 8.36 and 7.63, respectively. Table 1 shows the micro- and nanostructured MIP-based traps for environmental applications.

Conclusions This chapter highlights the recent applications of MIPs as tailor-made selective traps in extraction of pollutants such as such as pharmaceuticals, metal ions, and dye compounds from environmental samples. In recent years, there has been growing interest in SPE applications of MIP-based materials. The number of published examples confirmed that MIPs with high affinity and selectivity toward target compound(s) are effective and promising materials for the selective removal of

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Table 1 Micro- and nanostructured MIP-based traps for environmental applications Reference Composition of MIP-based trap Pharmaceutical compounds Piacham et al. QCM sensor surface coated with MIP (2005) layer using MAA as the functional monomer Mirmohseni QCM sensor surface coated with MIP and layer using MAA as the functional Houjaghan monomer (2013) Singh and QCM sensor surface coated with MIP Singh (2015) layer using 3-thiophene acetic acid (3-TAA) as the functional monomer Niu et al. Magnetic core-shell MIPs prepared by (2017) using functional monomers MAA and AAm Zazouli et al. MIP particles prepared by using MAA (2017) as the functional monomer Lu et al. MIP film prepared by using functional (2017) monomer MAA on the capacitive sensor surface Manzo et al. MIP particles prepared by using (2015) 1-vinylimidazole as the functional monomer He et al. Magnetic MIP nanoparticles prepared (2016b) by using functional monomer MAA Zhang et al. ZnS coated by MIP film layer using (2012) MAA as the functional monomer Bitar et al. MIP having methacrylamide as the (2017) functional monomer MIP having MAA as the functional Du et al. monomer (2014) Kubo et al. MIP having MAA as the functional (2014) monomer Chen et al. MIP having MAA as the functional (2013) monomer Metal ions Hashemi and Rb (I) imprinted nanoparticles having Shamsipur dibenzo-21-crown-7 as the ligand (2016) Monier and Au (III) imprinted polymer-chitosan Abdel-Latif composite modified with SH groups (2017) Xi et al. Cd (II) imprinted polymer having (2015) dithizone-Cd (II) complex Preetha et al. (UO2)(II) imprinted polymers having (2006) functional monomer 4-VP

Analyte

Sample

Propranolol

Aqueous solutions

Methomyl

Natural water

Melphalan

Aqueous solutions

Norfloxacin

Pharmaceutical wastewater

Sulfathiazole

Pharmaceutical wastewater Aqueous solutions

Pazufloxacin mesylate Diclofenac and mefenamic acid

Waste water

Dienestrol

Seawater

Ciprofloxacin

Difenoconazole

Aqueous solutions Aqueous solutions Tap water

Sulpiride

River water

Sulfonamides

Lake water

Rb (I)

Aqueous solutions

Au (III)

Aqueous solutions

Cd (II)

Wastewater

(UO2)(II)

Nuclear power reactor effluents

Iprodione

(continued)

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Table 1 (continued) Reference Liu et al. (2015b) Behbahani et al. (2015) Büyüktiryaki et al. (2007) Liu et al. (2015c) Kalal et al. (2013) Say et al. (2004) Fasihi et al. (2016) Khoddami and Shemirani (2016) Chen et al. (2016) Kong et al. (2014) Hou et al. (2015)

Composition of MIP-based trap Graphene-SiO2 composite having Ni (II) imprinted polymer having functional monomer acrylamide Pb (II) imprinted nanoparticles having functional monomer 4-VP CH3Hg (I) imprinted beads having MAC as the functional monomer Sr (II) imprinted polymer-chitosan composite having dithiocarbamate Ion imprinted aniline/formaldehyde resin CN imprinted polymer having MAH-Ni(II) as the complex monomer Silica particles grafted with (UO2) (II) imprinted polymer Magnetic MIPs having functional NH groups

Analyte Ni (II)

Sample Aqueous solutions

Pb (II)

Aqueous solutions Aqueous solutions Aqueous solutions Aqueous solutions Aqueous solutions Aqueous solutions Aqueous solutions

Ion imprinted magnetic chitosan composite Cr(VI) imprinted nanoparticles having functional monomer 4-VP Ag (I) imprinted hollow particles having glycidyl methacrylate as the functional monomer

Ni (II)

Dye compounds Franco et al. Magnetic molecularly imprinted (2017) polymer having acrylonitrile as the functional monomer Deng et al. MIP-chitosan composite having Ti (IV) (2017) Foguel et al. MIP having 1.vinylimizadole as the (2017) functional monomer Lu et al. Molecularly imprinted membrane (2015) having functional monomer MAA Lian and MIP having functional monomer MAA Wang (2012) Ziru and MIP having functional monomer MAA Jiangtao (2014) Kyzas et al. MIP having functional monomer AAm (2009) MIP having functional monomer MAA

CH3Hg (I) Sr (II) Ir (III) and Pd (II) CN () (UO2)(II) Co (II)

Cr (VI) Ag (I)

Aqueous solutions Aqueous solutions Aqueous solutions

Disperse Red 73

Aqueous solutions

Active Brilliant Red X-3B Acid green 16

Aqueous solutions Textile effluent and tap water Lake water

Rhodamine B Malachite green Crystal violet

Seawater

Reactive red Basic red

Aqueous solutions

Seawater

undesired compounds from environmental matrices such wastewater. Furthermore, these tailor-made materials can potentially be applied for large-scale extraction processes.

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Cheong WJ, Yang SH, Ali F (2013) Molecular imprinted polymers for separation science: a review of reviews. J Sep Sci 36:609–628 Cui M, Huang J, Wang Y, Wu Y, Luo X (2015) Molecularly imprinted electrochemical sensor for propyl gallate based on PtAu bimetallic nanoparticles modified graphene-carbon nanotube composites. Biosens Bioelectron 68:563–569 Deng H, Wei Z, Wang X (2017) Enhanced adsorption of active brilliant red X-3B dye on chitosan molecularly imprinted polymer functionalized with Ti(IV) as Lewis acid. Carbohydr Polym 157:1190–1197 Dolak I, Keçili R, Hür D, Ersöz A, Say R (2015) Ion-imprinted polymers for selective recognition of neodymium(III) in environmental samples. Ind Eng Chem Res 54:5328–5335 Du T, Cheng J, Wu M, Wang X, Zhou H, Cheng M (2014) An in situ immobilized pipette tip solid phase microextraction method based on molecularly imprinted polymer monolith for the selective determination of difenoconazole in tap water and grape juice. J Chromatogr B 951–952:104–109 Erdem M, Say R, Ersöz A, Denizli A, Türk H (2010) Imprinted polymer/organo-smectite nanocomposites for paraoxon hydrolysis. Appl Clay Sci 47:223–228 Fasihi J, Shamsipur M, Khanchi A, Mahani M, Ashtari K (2016) Imprinted polymer grafted from silica particles for on-line trace enrichment and ICP OES determination of uranyl ion. Microchem J 126:316–321 Foguel MV, Pedro NTB, Wong A, Khan S, Zanoni MVB, Sotomayor MDPT (2017) Synthesis and evaluation of a molecularly imprinted polymer for selective adsorption and quantification of acid green 16 textile dye in water samples. Talanta 170:244–251 Franco JH, Aissa AB, Bessegato GG, Fajardo LM, Zanoni MVB, Pividori MI, Sotomayor MDPT (2017) Assessment of molecularly imprinted polymers (MIPs) in the preconcentration of disperse red 73 dye prior to photoelectrocatalytic treatment. Environ Sci Pollut Res 24:4134–4143 Gao B, Chen T, Qiao Z, Fang L (2015) Preparation of acid dye molecule surface-imprinted material for effective removal of acid dyes from water and study on its molecule recognition performance. Sep Sci Technol 50:1108–1119 Guerreiro JRL, Freitas V, Sales MGF (2011) New sensing materials of molecularly-imprinted polymers for the selective recognition of chlortetracycline. Microchem J 97:173–181 Gül UD (2013) Treatment of dyeing wastewater including reactive dyes (Reactive Red RB, Reactive Black B, Remazol Blue) and Methylene Blue by fungal biomass. Water SA 39(5): 593–598 Han S, Li X, Wang Y, Chen S (2015) Multifunctional imprinted polymers based on CdTe/CdS and magnetic graphene oxide for selective recognition and separation of p-t-octylphenol. Chem Eng J 271:87–95 Hashemi B, Shamsipur M (2016) Synthesis of novel ion-imprinted polymeric nanoparticles based on dibenzo-21-crown-7 for the selective pre-concentration and recognition of rubidium ions. J Sep Sci 38:4248–4254 Hazzaa R, Hussein M (2015) Adsorption of cationic dye from aqueous solution onto activated carbon prepared from olive stones. Environ Technol Innov 4:36–51 He X, Mei X, Wang J, Lian Z, Tan L, Wu W (2016a) Determination of diethylstilbestrol in seawater by molecularly imprinted solid-phase extraction coupled with high-performance liquid chromatography. Marine Poll Bull 102:142–147 He X, Lian Z, Tana L, Wang J (2016b) Preparation and characterization of magnetic molecularly imprinted polymers for selective trace extraction of dienestrol in seawater. J Chromatogr A 1469:8–16 Hladik ML, Smalling KL, Kuivila KM (2008) A multi-residue method for the analysis of pesticides and pesticide degradates in water using HLB solid-phase extraction and gas chromatographyion trap mass spectrometry. Bull Environ Contam Toxicol 80:139–144 Hou H, Yu D, Hu G (2015) Preparation and properties of ion-imprinted hollow particles for the selective adsorption of silver ions. Langmuir 31:1376–1384

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Removal of Pharmaceutically Active Compounds from Contaminated Water and Wastewater Using Biochar as Low-Cost Adsorbents, An Overview

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Adel Al-Gheethi, Efaq Ali Noman, Radin Mohamed, Mohd Adib Mohammad Razi, and M. K. Amir Hashim Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Removal of PhACs from Water and Wastewater by Filtration Membrane . . . . . . . . . . . . . . . . . . . . Removal of PhACs from Water and Wastewater by Adsorption Process . . . . . . . . . . . . . . . . . . . . . . Biosorption of PhACs by Microorganisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cross-References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

Pharmaceutically active compounds (PhACs) have received high attention during the last few years due to their ability to persist for long time in the environment as well as their role in increasing the antimicrobial resistance among the floral bacteria in the nature. Many of the technologies have been investigated and applied to remove those compounds from the contaminated water and wastewater. The most common technology depends on the oxidation process which leads to degrade these compounds to be in inactive form. However, the oxidation process has some challenges which lie in the presence of secondary products and toxic by-products. The adsorption process is the best alternative technology

A. Al-Gheethi (*) · R. Mohamed · M. A. M. Razi · M. K. Amir Hashim Micro-Pollutant Research Centre (MPRC), Department of Water and Environment Engineering, Faculty of Civil and Environmental Engineering, University Tun Hussein Onn Malaysia (UTHM), Parit Raja, Johor, Malaysia e-mail: [email protected] E. A. Noman Faculty of Applied Sciences and Technology (FAST), Universiti Tun Hussein Onn Malaysia, Pagoh, Johor, Malaysia © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_25

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where no energy is required and no toxic by-products are generated, and the process leads to separate the pollutants from the water and wastewater. In this chapter, the adsorption of PhACs by low-cost materials such as biochar and microorganisms is discussed. Keywords

Biochar · Bacteria · Adsorptive capacity · Efficiency

Introduction One of the main pollutants in the water and wastewater which has high concerns by the researchers is the xenobiotic organic compounds (XOCs) which included endocrine-disrupting compounds (EDCs), disinfection by-products (DBPs), and pharmaceutically active compounds (PhACs), since these compounds have high persistence in the environment and can be transmitted through the water into the food and then into the animals and human. Some of these compounds have high risk health because they are synthesized in the laboratories and applied extensively as antimicrobial agents or as growth inducers. The concepts of XOCs have occurred in the 1990s due to the developments of the GC-MS and LC-MS techniques which have high accuracy to detect the concentration of XOCs even at very low concentrations (PPM). Moreover, antibiotics which are group of PhACs represent the most concern even at very low concentration due to their ability to induce the antimicrobial resistance among the bacteria in the natural water system. The high distribution of resistant bacteria in the environment and water system creates great issues especially when they are transmitted through the food chain into the human. Recently many of the pathogenic bacteria have exhibited high resistance for several types of the antibiotics. Indeed, it has to mention that most of the studies conducted on the removal of PhACs have been reported in the developed countries, with very few studies have been reported in the developing countries due to the absence of advanced technology. Moreover, there are no specific legislations and regulation for the presence of the PhACs in the wastewater disposed into the environment. The current chapter highlighted the most common removal of PHACs using adsorption process.

Removal of PhACs from Water and Wastewater by Filtration Membrane There are several technologies have been investigated to remove PhACs from the water and wastewater, the most common are the degradation process (chemical and biological oxidation) and adsorption process (natural and biological adsorbents). The filtration membrane is one of the efficient technologies to separate these compounds from the water. Kimura et al. (2003) examined the efficiency of

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nanofiltration (NF) and reverse osmosis (RO) membranes for removing PhACs based on the adsorption of these compounds on the physicochemical structure of the filter media. The study found that the filtration achieved >90% of the removal regardless of physicochemical properties of the tested compounds; the removal process takes place as a result of electrostatic exclusion. Therefore, the non-charged compounds were less removed ( CsNCB > Cs. The potential of silica zeolites (Y, mordenite, and ZSM-5) with pore opening sizes to adsorb sulfamethoxazole sulfonamide antibiotic was examined by Blasioli et al. (2014). The study investigated the removal process in the water and studied the effect of the temperature on the adsorption process. The results revealed that the zeolite Y and MOR exhibited fast and efficient removal of sulfamethoxazole (100%), while the kinetic adsorption by ZSM-5 associated with the temperature where the removal efficiency increased with the increasing of temperature indicating that the removal was a thermodynamic process. The analysis for the silica zeolites showed that the sulfamethoxazole sorption was incorporated and localized alongside the medium-weak and cooperative host-guest into the pore of each zeolite system, which means that the water molecules play a certain role only in zeolite Y and mordenite. One of the low-cost materials with high efficiency to adsorb PhACs is the biochar (BC), which is produced from the pyrolysis process for the carbon-rich biomass. It has high application as a soil amendment as well as sorption of organic compounds in the amended soil due to the high surface activity (Jha et al. 2010). The ability of biochar to adsorb the PhACs such as antibiotics has been reported in the literature; the high efficiency belongs to the interaction between the aromatic rings on the PhACs and biochar (Hurtado et al. 2017). However, the authors claimed that the efficiency of biochar to adsorb PhACs is influenced by humic acid (HA) dependent on the properties of BC, sorbate, and adsorbed HA (Guo et al. 2007; Duan et al. 2017). Lian et al. (2015) investigated the effect of loaded humic acid (HA) in two different modes including pre-coating and co-introduction with sorbate on biochar sorption for sulfonamides. The study revealed that the effect of HA on the sorption of sulfonamides by BC relied on the nature of interaction between HA fraction and sorbate as well as the HA concentrations. Nevertheless, it has to mention that the efficiency of the biochar in the adsorption process is less than the activated carbon due to the low porous available on the less surface area, which are the critical factors in the adsorption process (Falco et al. 2013). Therefore some of the authors have shifted to make a modified particles of biochar to improve the adsorption capacity. An advanced work on the biochar was conducted by Shan et al. (2016). In this study the biochar was prepared in an ultrafine magnetic particles biochar by using ball milling. The ultrafine magnetic particles of biochar were prepared as a magnetic biochar/Fe3O4 and activated carbon (AC)/Fe3O4. Thereafter, both materials were used for carbamazepine (CBZ) and

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tetracycline (TC) removal by adsorption and mechanochemical degradation. The study revealed that the biochar/Fe3O4 exhibited 62.7 mg/g for CBZ and 94.2 mg/g for TC of the adsorptive capacity, while AC/Fe3O4 achieved 135.1 mg/g for CBZ and 45.3 mg/g for TC, respectively. The investigation for the effect of pH on the adsorption process indicated that the removal of TC was dependent on the pH of the solution, while adsorption of CBZ was independent. In the study for the degradation of the adsorbed compounds, the study claimed that TC degraded by 97% within 3 h, while only 50% of the CBZ was degraded in the same period. In the study, the authors examined the addition of quartz sand during the preparation of BC and found that the presence of the sand enhanced the mechanochemical degradation of CBZ on biochar/Fe3O4 to 98.4% with 0.3 g quarts sand/g adsorbent. Zhang et al. (2016) synthesized a novel carbonaceous nanocomposite for sulfamethazine (SMT) sorption. The adsorbent was synthesized by pyrolysis of dip-coating straw biomass in carboxyl functionalized multiwalled carbon nanotube solution at 300  C and 600  C in air condition. The preparation method used here was effective for producing high sorption capacity materials. The carbonaceous nanocomposites were also mixed with the soil to understand the influence of the soil/biological and chemical aging on the adsorption process, and the results found no effect on the SMT sorption. The author explained the sorption mechanism based on the van der Waals forces as well as the hydrogen bonding and electron-donoracceptor interaction. Based on this study, it was concluded that the carbonaceous nanocomposites have potential in removal of PhACs from the contaminated water and wastewater. Zhang et al. (2011) used the corn straw biochars (CSB) as adsorbent for simazine; the CSB was prepared at temperature ranging from 100  C to 600  C. The adsorption process was explained based on the characteristics of biochars determined by BET-N2 surface area (SA), 13C NMR, and FTIR, while the isotherms of adsorption were studied with Freundlich and dual-mode models. The study revealed a positive correlation between log Koc values and aromatic C contents of the simazine, while exhibited a negative correlation with (O + N)/C ratios. These results indicated that the aromatic-rich biochars have high-binding affinity to simazine. The dual-mode model results claimed that the adsorption capacity of biochar is dependent on the carbonization degree. Therefore, the corn straw biochars prepared at high temperature can effectively adsorb the simazine. Ncibi and Sillanpää (2017) conducted several experiments on the removal of dorzolamide (DA) and carbamazepine (CZ) from artificially contaminated waters by adsorption process using as-synthesized multiwalled carbon nanotubes (intermittently MWCNTs or CNTs) and mesoporous activated carbons (Meso-AC). The optimization process was determined based on four factors which included solution pH, temperature, the application of ultrasonication, and exposure time. The study revealed that the efficiency of the adsorption process depended on the porous structure of the materials. The optimum removal process was noted at pH between 6 and 8. Moreover, the ultrasound process has enhanced the removal capacities of PhACs investigated in this study. Based on the Brouers-Sotolongo equation analysis,

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the isotherm model detected 99.5–99.9% of the removal efficiency. The maximum adsorptive capacity of MWCNT was 224.6 mg/g for CZ and 78.8 mg/g for DA; this capacity is higher than that reported for other adsorbents. However, it’s still less than that recommended by (Gadd 1990) where the minimum absorptive capacity should be more than 150 mg/g to be competitive with other technology. Lin et al. (2017) examined the effect of salts and organic matter on the efficiency of biochar and an activated carbon in removing ibuprofen and sulfamethoxazole from reclaimed water reverse osmosis (RO), synthetic and concentrate solutions spiked with selected organic compounds, as well as nontarget water constituents (trace elements, humic acid (HA), alkalinity, and bovine serum albumin (BSA)). The kinetic modeling studies were performed to explain the role of water composition on the adsorption process of PhACs. The study appeared that the adsorption capacity of biochar depended on the physicochemical properties such as ash content, specific surface area, charge, pore volume, as well as hydrophobicity, π-energy, and speciation of pharmaceuticals. The adsorption process was pH dependent; therefore, the removal efficiency associated with the elevation of pH value as a response for the formation of electrical interactions between ibuprofen and sulfamethoxazole and biochar, and for the same reason, the high salt contents improved the removal process, while HA and BSA reduced the adsorption capacity of biochar.

Biosorption of PhACs by Microorganisms The utilization of microorganism in the biosorption process has been investigated extensively by many of the authors in the literature. Most of those studies have been performed for heavy metal removal. The microorganisms which included bacteria, fungi, and microalgae have a novel cell wall structure which enables them to be attractive in the biosorption technique. The cell surface of the microorganisms has several functional groups such as hydroxyl, carboxyl, phosphate, and lipids which have a negative charge and can easily attract the positive charge pollutions such as heavy metals. However, few studies have been conducted in the using of microorganism cells to removal of PhACs based on the biosorption process. Al-Gheethi et al. (2014) investigated the potential of bacterial cell biomass (living and dead cells) to remove cephalexin antibiotic from the secondary effluents. In the study the bacterial cells were treated by autoclave (110  C for 10 min) and then used for the removal of cephalexin. The results revealed that the maximum biosorption capacity was 60 mg cephalexin g1 cells. Moreover, this efficiency has reduced by 40.83% and 82.88% (living and dead cells, respectively) in the presence of heavy metals such as Ni2+ in comparison with the control. In contrast, the biosorption by dead cell biomass was undetectable in the presence of Cd, Zn, Cu, and Pb ions, indicating the less affinity of cephalexin to the bacterial cells in comparison with the heavy metals. Moreover, in another study Al-Gheethi et al. (2017) used consortium bacterial cells to remove cephalexin from the aqueous solution. The factors affecting the

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removal process which included cephalexin concentrations, biomass concentration, pH, temperature, and time were tested. In the study the bacteria species were subjected for the adaption process to tolerate the cephalexin in order to increase the biosorption capacity. The results revealed that the maximum biosorption efficiency was recorded at low concentrations of cephalexin (94.73% vs. 92.98% for living and dead cells, respectively). The dead cells have more capacity to adsorb the cephalexin than the living cells at high concentrations of cephalexin. The optimum pH was determined between pH 4 and 6 (71.95–68.90%), while the adsorption process was more efficient at 25  C and 30  C. However, similar findings were also noted in this study where the affinity of cephalexin is reduced in the presence of heavy metals. The most common microorganisms which are used for biosorption of PhACs are presented in Table 1. Table 1 Biosorptive capacity of some types of biomass to remove antibiotics (Adopted from Al-Gheethi et al. 2015) Biomass R. arrhizus

Antibiotic Penicillin G

Biosorptive capacity (Qmax) 459.0 mg g1

Active carbon

Amoxicillin

200 mg g1

Activated sludge

Graphene oxide (GO)

Tylosin Tetracycline Cephalexin Cefixime Penicillin Ampicillin Cephalosporin Tetracyclines

7.7 mg g1 72 mg g1 1100 mg g1 820 mg g1 427.3 mg g1 164.2 mg g1 33.67 mg g1 313 mg g1

Sludge biochar

Fluoroquinolone

19.80 mg g1,

Biocomposite fibers of graphene oxide/calcium alginate B. subtilis 1612WTNC (living cells) B. cepacia 103WTC (dead cells) Mixed Gram-positive bacteria (living cells) Mixed Gram-negative bacteria (living cells) Mixed Gram-positive bacteria (dead cells) Mixed Gram-negative bacteria (dead cells) Mixed bacterial biomass (living cells)

Ciprofloxacin

18.45–39.06 mg/g

Cephalexin

35.02 mg g1 40.74 mg g1 50.91 mg g1

Multiwalled carbon nanotubes (CNTs) Granular Activated carbon (GAC)

40.44 mg g1 15.99 mg g1 25.11 mg g1 60 mg g1

References Aksu and Tunç (2005) Budyanto et al. (2008) Prado et al. (2009) Jafari et al. (2011) Liang (2011)

Gao et al. (2012) Yao et al. (2013) Wu et al. (2013) Al-Gheethi (2014)

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Conclusion It appeared that many of the low-cost materials have been used for the removal of PhACs. However, the main challenges lie in the less affinity of these compounds to be adsorbed on the surface of the adsorbent. Moreover, the preparation of biochar at high temperature might improve their ability to be much better than activated carbon in removing of PhACs. Biochar represents one of the most alternative materials for the activated carbon. The potential of microorganisms in the biosorption of PhACs has less investigations. However, the most potent microorganisms might be mixed with the biochar to increase the surface area of the biochar and then enhance the adsorption efficiency. More work is needed for getting an efficient adsorbent for removal of PhACs from the contaminated water and wastewater.

Cross-References ▶ Development In-House: A Trap Method for Pretreatment of Fat, Oil, and Grease in Kitchen Wastewater ▶ Microbial Risk Associated with Application of Biosolids in Agriculture ▶ Treatment of Domestic Gray Water by Multicomponent Filters

References Aksu Z, & Tunç, Ö (2005) Application of biosorption for penicillin G removal: comparison with activated carbon. Process biochemistry 40(2):831–847 Al-Gheethi AAS (2014) Screening of bacterial isolates from sewage treated effluent with potential to remove heavy metals and β-lactam antibiotics. Ph.D. Thesis, Environmental Technology Division, School of Industrial Technology, University Science Malaysia. Penang Al-Gheethi AA, Norli I, Lalung J, Azlan AM, Farehah ZN, Kadir MOA (2014) Biosorption of heavy metals and cephalexin from secondary effluents by tolerant bacteria. Clean Techn Environ Policy 16(1):137–148 Al-Gheethi AA, Lalung J, Noman EA, Bala JD, Norli I (2015) Removal of heavy metals and antibiotics from treated sewage effluent by bacteria. Clean Techn Environ Policy 17(8): 2101–2123 Al-Gheethi AA, Efaq AN, Mohamed RM, Norli I, Kadir MO (2017) Potential of bacterial consortium for removal of cephalexin from aqueous solution. J Assoc Arab Univ Basic Appl Sci 24:141–148 Blasioli S, Martucci A, Paul G, Gigli L, Cossi M, Johnston CT, . . ., Braschi I (2014) Removal of sulfamethoxazole sulfonamide antibiotic from water by high silica zeolites: a study of the involved host–guest interactions by a combined structural, spectroscopic, and computational approach. J Colloid Interface Sci 419:148–159 Budyanto S, Soedjono S, Irawaty W, Indraswati N (2008) Studies of adsorption equilibrium and kinetics of amoxicillin from simulated wastewater using activated carbon and natural bentonite. J Environ Prot Sci 2:72–80 Comerton AM, Andrews RC, Bagley DM, Yang P (2007) Membrane adsorption of endocrine disrupting compounds and pharmaceutically active compounds. J Membr Sci 303(1):267–277

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Duan M, Li H, Gu J, Tuo X, Sun W, Qian X, Wang X (2017) Effects of biochar on reducing the abundance of oxytetracycline, antibiotic resistance genes, and human pathogenic bacteria in soil and lettuce. Environ Pollut 224:787–795 Falco C, Marco-Lozar JP, Salinas-Torres D, Morallon E, Cazorla-Amorós D, Titirici MM, LozanoCastelló D (2013) Tailoring the porosity of chemically activated hydrothermal carbons: influence of the precursor and hydrothermal carbonization temperature. Carbon 62:346–355 Gao Y, Li Y, Zhang L, Huang H, Hu J, Shah SM, Su X (2012) Adsorption and removal of tetracycline antibiotics from aqueous solution by graphene oxide. J Colloid Interface Sci 368(1):540–546 Gadd GM (1990) Heavy metal accumulation by bacteria and other microorganisms. Experientia 46(8):834–840 Guo Y, Yadav A, Karanfil T (2007) Approaches to mitigate the impact of dissolved organic matter on the adsorption of synthetic organic contaminants by porous carbonaceous sorbents. Environ Sci Technol 41(22):7888–7894 Hurtado C, Cañameras N, Domínguez C, Price GW, Comas J, Bayona JM (2017) Effect of soil biochar concentration on the mitigation of emerging organic contaminant uptake in lettuce. J Hazard Mater 323:386–393 Jafari M, Aghamiri SF, Khaghanic G (2011) Batch adsorption of cephalosporins antibiotics from aqueous solution by means of multi-walled carbon nanotubes. World Appl Sci J 14:1642–1650 Jha P, Biswas AK, Lakaria BL, Rao AS (2010) Biochar in agriculture–prospects and related implications. Curr Sci 99:1218–1225 Kimura K, Amy G, Drewes JE, Heberer T, Kim TU, Watanabe Y (2003) Rejection of organic micropollutants (disinfection by-products, endocrine disrupting compounds, and pharmaceutically active compounds) by NF/RO membranes. J Membr Sci 227(1):113–121 Kyzas GZ, Kostoglou M, Lazaridis NK, Lambropoulou DA, Bikiaris DN (2013) Environmental friendly technology for the removal of pharmaceutical contaminants from wastewaters using modified chitosan adsorbents. Chem Eng J 222:248–258 Lian F, Sun B, Chen X, Zhu L, Liu Z, & Xing B (2015). Effect of humic acid (HA) on sulfonamide sorption by biochars. Environmental Pollution 204:306–312 Liang S (2011) Removal of antibiotics from wastewater by adsorption and biodegradation. Ph.D. Thesis, School of Civil & Environmental Engineering. Singapore Lin L, Jiang W, Xu P (2017) Comparative study on pharmaceuticals adsorption in reclaimed water desalination concentrate using biochar: impact of salts and organic matter. Sci Total Environ 601:857–864 McCallum EA, Hyung H, Do TA, Huang CH, Kim JH (2008) Adsorption, desorption, and steadystate removal of 17β-estradiol by nanofiltration membranes. J Membr Sci 319(1):38–43 Ncibi MC, Sillanpää M (2017) Optimizing the removal of pharmaceutical drugs Carbamazepine and Dorzolamide from aqueous solutions using mesoporous activated carbons and multi-walled carbon nanotubes. J Mol Liq 238:379–388 Prado N, Ochoa J, Amrane A (2009) Biodegradation and biosorption of tetracycline and tylosin antibiotics in activated sludge system. Process Biochem 44:1302–1306 Shan D, Deng S, Zhao T, Wang B, Wang Y, Huang J, . . ., Wiesner MR (2016) Preparation of ultrafine magnetic biochar and activated carbon for pharmaceutical adsorption and subsequent degradation by ball milling. J Hazard Mater 305:156–163 Wu S, Zhao X, Li Y, Zhao C, Du Q, Sun J, Xia L (2013) Adsorption of ciprofloxacin onto biocomposite fibers of graphene oxide/calcium alginate. Chem Eng J 230:389–395 Yao H, Lu J, Wu J, Lu Z, Wilson PC, Shen Y (2013) Adsorption of fluoroquinolone antibiotics by wastewater sludge biochar: role of the sludge source. Water Air Soil Pollut 224:1370 Zhang G, Zhang Q, Sun K, Liu X, Zheng W, Zhao Y (2011) Sorption of simazine to corn straw biochars prepared at different pyrolytic temperatures. Environ Pollut 159(10):2594–2601 Zhang C, Lai C, Zeng G, Huang D, Yang C, Wang Y, Cheng M (2016) Efficacy of carbonaceous nanocomposites for sorbing ionizable antibiotic sulfamethazine from aqueous solution. Water Res 95:103–112

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Adel Al-Gheethi, Efaq Ali Noman, Radin Mohamed, Abd. Halid Abdullah, and M. K. Amir Hashim

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biosolid-Borne Bacteria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Risk Assessment of Pathogenic Bacteria in Soil Fertilized with Biosolids . . . . . . . . . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cross-References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

The present chapter aimed to highlight the pathogenic risk associated with the application of biosolids in agriculture. The main pathogens and their health importance as well as their ability to survive and be transmitted into humans through the food chain have been discussed. It has appeared that the direct utilization of biosolids in agriculture represents one of the main sources of human and animal infections due to the ability of most pathogens to survive in the environment. Therefore, the biosolids should be subjected to further treatment process in order to

A. Al-Gheethi (*) · R. Mohamed · M. K. Amir Hashim Micro-Pollutant Research Centre (MPRC), Department of Water and Environment Engineering, Faculty of Civil and Environmental Engineering, University Tun Hussein Onn Malaysia (UTHM), Parit Raja, Johor, Malaysia e-mail: [email protected] E. A. Noman Faculty of Applied Sciences and Technology (FAST), Universiti Tun Hussein Onn Malaysia, Pagoh, Johor, Malaysia A. H. Abdullah Micro-pollution Research Centre (MPRC), Department of Water and Environmental Engineering, Faculty of Civil and Environmental Engineering, Universiti Tun Hussein Onn Malaysia, Batu, Pahat, Malaysia © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_26

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reduce the pathogen level to less than the health risk which lies in the elimination of these pathogens or minimize their ability to regrow in the environment. Keywords

Pathogenic bacteria · Reuse · Sludge · Agricultural purpose

Introduction One of the most important and challenging steps in sewage treatment systems is the disposal process of the generated sludge due to high contents of organic matter, heavy metals, as well as pathogenic entities such as bacteria, viruses, protozoa, and helminths (Al-Gheethi et al. 2013). The stricter regulations adopted by several countries on the disposal of sewage into landfill, high costs of chemical fertilizers, and their toxicity on human health have increased interest in the reuse of sludge in the crop production. The sludge consists of organic matter and is rich with nutrients which play an important role in increasing soil fertilization. In contrast, the sludge is rich with high diversity of bacteria and heavy metals that might have adverse effects on crop production and human health. Therefore, before the reuse of sludge as fertilizers, it must be treated to remove the remaining putrescible material as well as enteric bacteria to prevent the danger of spreading enteric diseases. The management of the ever-increasing volume of the sludge has been one of the prime environmental issues in several counties. However, developing countries have yet to adopt a practical, economic, and acceptable approach in managing and disposing sludge. The amount of sludge resulted from sewage treatment plants (STPs) have been increasing at a rapid pace in recent years due to the high increase in population equivalent. The human wastes have drawn serious attention from the society (Cheremisinoff 1994). In Malaysia, 2.97 billion m3 of sewage sludge is produced by Indah Water Konsortium (IWK) annually; this sludge volume is expected to rise to 7 million cm3 by the year 2020 (Azman and Shaari 2013). However, there is no information on the reuse of sludge in crop production as the Government of Malaysia has yet to have a policy on the reuse of sludge for agriculture purposes (Azman and Shaari 2013). In the European Union (EU), the total quantity of produced sludge increased from 7.2 million tons of dry matter in 1998 to 9.8 million tons in 2009, about 37.0% of these sludges was used for agriculture, and the worldwide production of sludge is estimated in limits of 20 billion tons annually (Evans 2012). Many disposal methods are applied for the sludge including ocean dumping, incineration, spreading on agricultural land, soil incineration, land spreading in forestry, or landfilling (Supakata and Chunkao 2011). The most common practice of sludge in developing countries is disposal to landfill. However, in EU, these practices have reduced from 40% in 1994 to 17% in 2005. In 2009 only 1% of sludge was disposed into the landfill. These reductions were due to the presence of strict standards relating to landfill of biodegradable waste (Holt et al. 1995). Land application of sludge has several benefits for plants which include supplying nutrients (nitrogen, phosphorous, and trace elements) to crops. The application of sludge as fertilizers might improve the soil physical properties because sludge typically

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possesses excellent soil amendment properties and increases soil organic matter content (Hamouda and Abu-Shaaban 2014). Alongside, biosolids contain constituents that may pose a risk to soil, water, plants, animals, or public health. The concerns lie in the heavy metals, soluble salts, and other trace constituents which pose serious concerns and can damage soils, plants, animals, and humans as well as the potential for disease transfer by high pathogen diversity in the sludge. There are many sources of biosolids such as those generated from animals or poultry wastewater. However, this review focused only on biosolids resulted from human sewage. In this review, the health risk of human associated with reuse biosolids in the agriculture are discussed based on the pathogenic bacteria. This work will focus on the adverse effects of pathogenic bacteria on humans and plants and the possibility to accumulate these pollutants in land fertilized with the biosolids.

Biosolid-Borne Bacteria Microorganisms in raw sewage are transmitted to biosolids due to the stabilization and sedimentation process of sewage treatment system, in which bacterial cells are adsorbed on the solid materials and then precipitate with these materials (Strauch 1991, 1998). US EPA (2003) revealed that the pathogens in the domestic sewage are primarily associated with insoluble solids. These organisms become bound to solids following wastewater treatment and are transferred to biosolids. Hence, the biosolids have higher quantities of pathogens than wastewater (Al-Gheethi et al. 2014). However, it is important to indicate that not all bacteria in biosolids are considered harmful, since many different organisms live within the sewage itself, assisting in the decomposition of organic pollutants. These organisms also helped to convert the raw sewage into biosolids to become incorporated into a Class A or B biosolids (Gerba and Smith 2005). In this section, the adverse effect of pathogenic bacteria in sludge for plants and human will be discussed based on the potential to survive in the environment and the pathogenicity for human and plants. Biosolids contain different diversity and density of pathogenic bacteria depending on the treatment process efficiency used in sewage treatment plants (STPs). Further, raw and untreated biosolids contain significantly higher diversity and density of pathogens than secondary and treated or dried biosolids (Al-Gheethi et al. 2014). The quality of biosolids depends on the health of the population contributing to STPs, and as the health of the population varies from one time to another, the characteristics of biosolids also vary (Harrison et al. 1999). Therefore, occurrences of pathogenic organisms in biosolids in infected population communities even in the small community are more frequent than in the large communities (US EPA 2002); the quality of biosolids also varied geographically and seasonally at the same STPs. The nature of biosolids as human wastes always causes a hygienic risk in storage, collection, processing, handling, and utilization. Straub et al. (1993) reported that the raw sludge might release airborne pathogens more than the treated ones. Therefore, strict regulations for hygienic principles must be applied during the storage, transport, and distribution of these wastes (Martens et al. 1998). The basic hygienic risk is the occurrence of pathogens in biosolids which represents the starting point

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for understanding the epidemiological reflections and necessary precautions (Strauch 1991). There are three types of health risks related to the presence of pathogenic bacteria in biosolids which have to be considered during the reuse of biosolids in agriculture. These risks included occupational health risks, risks concerning product safety, and environmental risks (Strauch 1998). According to US EPA (2007), biosolids generated from STPs are classified into Class A and B based on microbiological characteristics. In Class A, fecal coliform (FC) bacteria should be less than 1000 cell/g of dry wt, and Salmonella spp. are less than 3 cell/25 g of dry wt. In Class B, FC are more than 1 million/g of dry wt, while no criteria are required for Salmonella spp. Therefore, the utilization of biosolids in the agriculture depends on the class of biosolids. In class A, biosolids can be utilized as fertilizers of soils used for crop production. In contrast, the soil amended with class B should be exposed to sunlight for at least 1 year before utilization for agriculture in which the pathogenic bacteria in biosolids would be reduced to be less than the risk level (US EPA 2003). Pathogenic bacteria in biosolids reused for agriculture can be divided into two groups: plant pathogenic bacteria and human pathogenic bacteria. Plant pathogenic bacteria are those released from the vegetable wastewater. These pathogens grow at ambient environmental temperature; therefore, the sewage treatment processes performed at mesophilic temperature might be enough to reduce their concentrations. In the term of human pathogenic bacteria, a wide range of pathogen and opportunistic pathogens have been reported in biosolids. The most significant pathogenic bacteria in biosolids include Campylobacter jejuni, enteropathogenic E. coli, Leptospira spp., Salmonella spp., Shigella spp., Vibrio cholerae, Listeria monocytogenes, Staphylococcus aureus, and Yersinia spp. (Straub et al. 1993; US EPA 1994; Viswanathan and Kaur 2001). Acinetobacter spp., Alcaligenes spp., Flavobacterium spp., Pseudomonas spp., and Zoogloea spp. have been also reported as predominant in biosolids (Kappesser et al. 1989; Wen et al. 2009). The pathogenic bacteria in biosolids include Aeromonas hydrophila, Acinetobacter cloaca, Bacillus anthracis, E. coli, Campylobacter spp., Clostridium botulinum, C. perfringens, diphtheroid spp., Edwardsiella spp., Enterobacter spp., Flavobacterium spp., Hafnia spp., Klebsiella spp., Leptospira spp., Listeria monocytogenes, Mycobacterium tuberculosis, M. bovis, Proteus spp., Providencia spp., Pseudomonas aeruginosa, Salmonella paratyphi, S. typhimurium, Sarcina spp., Serratia spp., Shigella spp., Staphylococcus spp., Vibrio cholerae, and Yersinia enterocolitica (Straub et al. 1993; Viswanathan and Kaur 2001; Al-Gheethi et al. 2014). Ashbolt et al. (2001) found Pseudomonas, Streptococcus, Flavobacterium, and Aeromonas species (called opportunistic pathogens) in biosolids. Markosyan et al. (2002) indicated that among several species of opportunistic bacteria in biosolids, Klebsiella pneumonia, K. oxytoca, Enterobacter, Hafnia, Serratia, Proteus, and Providencia are the most prevalent. Ibekwe and Grieve (2003) revealed that the dominant bacteria in biosolids are related to Bacillus, Clostridium, Mycoplasma, Eubacterium, and Proteobacteria which are originally retrieved from the gastrointestinal tracts of mammals. The bacterial diversity in the biosolids varied from one region to another, depending on the treatment process, for instance, Urdaci et al. (1991) isolated

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2–39% of bacteria from biosolids at Spain, and these bacteria were identified within 23 species of Vibrio. Autheunisse and Koene (1987) indicated that Acinetobacter spp., Pseudomonas spp., Aeromonas spp., and Flavobacterium spp. are more common in the biosolids generated from aerobic treatment process. The predominantly foodborne in biosolids include Arcobacter spp., Bacillus cereus, C. coli, C. jejuni, C. perfringens, Campylobacter fetus, Clostridium botulinum, E. coli O111:NM, E. coli O104:H21, Escherichia coli O157:H7, Listeria monocytogenes, S. enteritidis, S. typhimurium, Salmonella typhi, Shigella sp., Staphylococcus aureus, V. parahaemolyticus, V. vulnificus, Vibrio cholerae O1, and Y. enterocolitica (Dumontet et al. 2001; Stampi et al. 1999). In general, the pathogenic bacteria in biosolids can be divided into major and minor concerns based on their significance to human health as stated by Dudley et al. (1980) and Kowal (1983). Salmonella sp., Shigella sp., Vibrio cholerae, E. coli (pathogenic strains), Campylobacter jejuni, Leptospira sp., and Yersinia enterocolitica are classified within the major concern, while Aeromonas sp., Clostridium perfringens, Pseudomonas aeruginosa, and Staphylococcus are classified within the minor concern category. In some literature, the pathogenic bacteria in biosolids are classified as primary and opportunistic pathogens. The most important pathogenic bacteria are those transmitted by the fecal-oral route, as they have already acclimated to infect and grow at the human temperature (US EPA 2003). Enteric pathogenic bacteria within biosolids represent a highly diversified group due to the presence of high contents of nutrients. For instance, E. coli strains which is one of the main enteric bacteria in biosolids are composed of Shiga toxin-producing E. coli (STEC) and enterohemorrhagic E. coli (EHEC) associated with potable and recreational water (Kaper et al. 2004). However, non-enteric pathogenic bacteria which included Legionella spp., Mycobacterium spp., Leptospira spp., and Pseudomonas spp. have also been detected in biosolids (Toze 1997). The most attention in terms of risk to human health is Salmonella and E. coli O157:H7. Both bacteria have been isolated by many authors in the literature (Dudley et al. 1980; Droffner and Brinton 1995; Hoeller et al. 1999; Sahlström et al. 2004). According to WHO (1981), Salmonella spp. are the most relevant species in biosolids. However, the risk of this bacterium to humans is associated with the food animal which acts as carriers of salmonellosis as a result of grazing in pastures or fed crops grown on biosolid-amended fields. In Denmark, biosolids are considered Salmonella-positive when WWTP received sewage sludge from more than 4000 people (Larsen 1998). Burtscher and Wuertz (2003) indicated that Salmonella spp. were detected by 48% of the 46 biosolid samples collected during intermediate stages of biosolid treatment process. Lucero-Ramirez (2000) reported that Salmonella are pathogenic bacteria of major concern in biosolid management, primarily when biosolids are considered for land disposal. Dumontet et al. (2001) revealed that Salmonella spp. are the most widespread bacterial pathogens of significant global public health concern that are likely to cause an important biosolid contamination. S. senftenberg is one of the Salmonella spp. which have been associated with consumption of raw foods that have been fertilized or in contact with poultry wastes (Taormina et al. 1999). According to US EPA (2003), Salmonella spp. are bacteria of great concern as well as good representatives of reduction of other bacterial

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pathogens due to their abundant with higher densities in biosolids and ability to survive for the reduction process of pathogens in biosolids. S. senftenberg has been suggested as it is relatively heat resistant but rarely affects humans. S. stanley was the only pathogen isolated after the thermophilic anaerobic digestion (Sahlström et al. 2004). Among 2000 serotypes of Salmonella spp., two serotypes of Salmonella, S. typhi and S. paratyphi (A, B, C), are most dangerous to people. The high pathogenicity of Salmonella spp. is due to their ability to infect nearly all living vectors from insects to mammals (Strauch 1991), besides the potential to resist a wide range of antibiotics. There are four species of Shigella spp. that have been isolated from biosolids. However, S. dysenteriae is among the most common causes/sources of diarrhea in humans. Yersinia spp. have isolated from anaerobically biosolids (Dudley et al. 1980). This bacterium excretes in feces and transmits by direct contact or food poisoning. Moreover, the waterborne outbreaks are due to fecal contamination of water. Aeromonas caviae is one of the dominant species in biosolids; this bacterium is transmitted into human via water contaminated with biosolids, and thus it may be a potential indicator of biosolid pollution (Ramteke et al. 1993). Aeromonas hydrophila has also been noted in biosolids (Poffe and Beek 1991). The importance of biosolids in the dissemination of L. monocytogenes in the environment has been detected. The presence of L. monocytogenes has been reported in biosolids (Sahlström et al. 2004) and on plants (Weis and Seeliger 1975), including vegetables (Beuchat 1996). Besides, L. monocytogenes has been isolated from contaminated surface water (Watkins and Sleath 1981). De Luca et al. (1998) demonstrated that biosolids contain L. monocytogenes. The presence of L. monocytogenes in biosolids was confirmed by De Luca et al. (1998) who tested five different types of biosolids (primary raw, activated, thickened, digested, and dewatered) in an Italian sewage treatment plant and revealed that L. monocytogenes was detected in all biosolid types. Hence, L. monocytogenes should be considered as a potential health risk in soil amended with biosolids. De Luca et al. (1998) also demonstrated the presence of L. monocytogenes in biosolids. However, the occurrence of L. monocytogenes was correlated with the season as it was more abundant in spring and autumn. The presence of L. monocytogenes in biosolids is due to their ability to survive during the biological oxidation. Campylobacter spp. have been isolated from biosolids; among the different species, C. jejuni and C. coli are the most common (Sahlstrom et al. 2004; Bagge et al. 2005). Stampi et al. (1999) indicated that Campylobacter spp. are available only in untreated biosolids and are absent from the secondary biosolids. This is because Campylobacter spp. are quite sensitive to mesophilic anaerobic digestion (Stelzer and Jacob 1991). However, Kearney et al. (1993) revealed that C. jejuni could still be detected even after 112 days of the storage period of biosolids treated by mesophilic anaerobic digestion. Campylobacter are often connected with outbreaks caused by fecal contamination of water (Gallay et al. 2006). There are few reports which indicate the presence of Mycobacterium spp. in biosolids. However, Dudley et al. (1980) and Pickup et al. (2006) isolated

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unspecified Mycobacterium spp. from biosolids. Dailloux et al. (1999) isolated M. tuberculosis from biosolids of swage coming from hospitals. M. paratuberculosis has been found in biosolids in the United Kingdom (Pickup et al. 2005). Clostridium spp. are present in anaerobically digested sewage sludge (Dudley et al. 1980). Using digested residue as a fertilizer could contaminate the silage, because neither anaerobic digestion nor ensiling inactivates C. tyrobutyricum (Johansson et al. 2005). Several factors affecting the potential of pathogenic bacteria from biosolids to cause infection to human include the ability of the bacteria to survive in the biosolids and environment, infective dose, and pathogenicity. With regard to survival in the biosolids and environment, Strauss (2002) revealed that Salmonella spp. may survive in biosolids over 100 days in lower moisture (10–15  C) and for 30 days in higher moisture (20–30  C), while fecal coliforms survived for over 30 and 8 days, respectively. Winfield and Groisman (2003) reported that Salmonella spp. can survive for weeks in aquatic environments and for months in soils and sediments. There are two variables needed to quantify the risk from pathogens: (1) exposure to a sufficient quantity of pathogens from inhalation or ingestion and (2) the dose of pathogens must be in a sufficient quantity to overwhelm the immune system’s ability to contain the pathogen (Epstein 2002).

Risk Assessment of Pathogenic Bacteria in Soil Fertilized with Biosolids The risk assessment of pathogenic bacteria in soil fertilized with biosolids aimed to evaluate the presence of hazard in soil and to characterize the risks associated with the hazard (Grohmann et al. 2003). In the 1970s, the authors have indicated that there are no incidences of disease from land application of biosolids detected, but pathogenic bacteria associated with biosolids present a potential problem (Pahren et al. 1979). Those conclusions were based on the fact that the biosolid-borne bacteria in land application systems are poor competitors outside the host. However, these reasons are theoretical reasons, the absence of critical techniques for determination of pathogenic bacteria in the environment is the main reason to suppose the possibility and not confirmations to presence a health risk associated with pathogenic bacteria in soil fertilized with biosolids, since the researchers theorize that some of pathogenic bacteria may enter a viable but non-culturable (VBNC) state under the stress environment, these organism are not accurately measured by standard culturing methods. Moreover, the explanation for why only some of pathogenic bacteria have the potential to survive in the environment while others cannot needs more studies. Risk assessment requires a clear methodological analysis (Huertasa et al. 2008). Therefore, many of the technologies for the determination of pathogenic bacteria in the environment have developed during the last decade. These technologies contributed significantly to risk assessment associated with pathogenic bacteria in soil.

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Besides the survival of pathogenic bacteria in the soil is affected by temperature, moisture, sunlight, the availability of organic matter, soil pH, soil particles, and the presence of toxic substances, competitive organisms influence bacteria survival in soils (Cools et al. 2001; Zhang et al. 2013; Ma et al. 2014). These factors significantly influence the survival of bacteria. Naganandhini et al. (2015) investigated persistence of Shiga-like toxin-producing E. coli (STEC) strains (O157-TNAU) and nonpathogenic strain (MTCC433) in different soils of India. The study revealed that the red laterite and tropical latosol enhanced E. coli O157-TNAU and MTCC 433 survival more than wetland and black cotton soils. In coco peat, E. coli O157 survived longer than E. coli MTCC 433. The analysis of data using double Weibull model and the modeling parameters that were correlated with soil physicochemical and biological properties using principal component analysis (PCA) revealed that pH, microbial biomass carbon, dehydrogenase activity, and available N and P contents of the soil improved E. coli strains in those soils and coco peat.

Conclusion Many of the pathogenic bacteria are associated with the biosolids. Therefore, high health risk is associated with the application of these biosolids in the agriculture. More advanced technologies are required to minimize the risk associated with the biosolids by reducing the concentration of the pathogens to less than detection limits.

Cross-References ▶ Development In-House: A Trap Method for Pretreatment of Fat, Oil, and Grease in Kitchen Wastewater ▶ Removal of Pharmaceutically Active Compounds from Contaminated Water and Wastewater Using Biochar as Low-Cost Adsorbents, An Overview ▶ Treatment of Domestic Gray Water by Multicomponent Filters

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Monitoring and Risk Analysis of PAHs in the Environment

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Karishma Hussain, Raza R. Hoque, Srinivasan Balachandran, Subhash Medhi, Mohammad Ghaznavi Idris, Mirzanur Rahman, and Farhaz Liaquat Hussain

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Polycyclic Aromatic Hydrocarbon . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Molecular Structure and Physicochemical Properties of PAHs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Exposure to PAHs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Toxicity and Carcinogenicity of PAHs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sources of PAHs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Distribution in the Environment (Air, Soil, Street Dust, Water, and Sediment) . . . . . . . . . . . . . . . PAHs in the Atmosphere . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . PAHs in Soil . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . PAHs in Street Dust . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . PAHs in Water and Sediment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . PAHs Analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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K. Hussain (*) Department of Bioengineering and Technology, Gauhati University, Guwahati, Assam, India e-mail: [email protected] R. R. Hoque Department of Environmental Science, Tezpur University, Tezpur, Assam, India e-mail: [email protected] S. Balachandran Department of Environmental Studies, Visva-BharatiSantiniketan, Santiniketan, West Bengal, India e-mail: [email protected] S. Medhi · M. G. Idris Department of Bioengineering and Technology (GUIST), Gauhati University, Guwahati, India e-mail: [email protected]; [email protected] M. Rahman Department of Information Technology (GUIST), Gauhati University, Guwahati, India e-mail: [email protected] F. L. Hussain Research Scholar, Department of Chemistry, Dibrugarh University, Dibrugarh, Assam, India e-mail: fl[email protected] © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_29

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Ecotoxic Effects of PAHs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Toxicity Analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Ecosystem Risk . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

Polycyclic aromatic hydrocarbons (PAHs) are a unique class of organic pollutants containing two or more fused aromatic rings, which are toxic and potential carcinogens. They are extensively studied compounds, and their occurrence has been reported from various places over the world which indicates their ubiquitous nature in our environment. Anthropogenic sources of PAHs are more dominant than their natural source which include sources like combustion engines, residential heating, industrial activities, and biomass burning. USEPA has already listed 16 PAHs [naphthalene, acenaphthylene, acenaphthene, fluorine, phenanthrene, anthracene, fluoranthene, pyrene, benzo(a)anthracene, chrysene, benzo(b) fluoranthene, benzo(k)fluoranthene, benzo(a)pyrene, dibenzo(a,h)anthracene, benzo(g,h,i)perylene, and indeno(1,2,3-c,d)pyrene] as most priority ones to be analyzed in various environmental matrices. More so, benzo(a)pyrene is termed as index or gold standard of the whole group of PAHs due to its high carcinogenic potency. Once in the atmosphere, depending on their physical and chemical properties, PAHs get distributed between gas, particle, and droplet phase. Aerial movement is one of the major pathways for environmental distribution and transboundary deposition of PAHs. Eventually, PAHs settle down in soils and street dust and enter into aquatic environment. Soil and street dust act as direct sink of atmospheric PAHs near to traffic and other combustion sources. From these environmental compartments, rainwater and storm water easily wash away PAHs to nearby aquatic bodies. Due to hydrophobic nature, PAHs in aquatic environment are preferably partitioned and accumulate into the particulate phase of sediment. PAHs, thus, occur in multicompartmental system of the environment and paved the way for multiple routes of exposure to this class of carcinogen. Therefore, extensive studies have been carried out for PAHs in different environmental matrices over the world, and many places are revealed with very high exposure levels of PAHs. Environmental PAHs have harmful effects on different types of organisms of the ecosystem. PAHs attract considerable attention among researchers due to continuous rise in death toll of human cancer worldwide. Toxic equivalency factors (TEFs) were often employed to assess carcinogenic potential of individual PAHs. Here, the maximum TEF of one is assigned to BaP, and other individual PAHs are relative to BaP as BaP equivalents (BaPq). To characterize risk of PAHs to surrounding organisms and ecosystems in aquatic environment, ecosystem risk was often employed by researchers by using risk quotient (RQ) of individual PAHs. Risk quotients (RQ) value indicated the levels of risk posed by certain PAHs. Studies are revealed with high exposure risk in different environmental matrices in several places around the world. However, only a few recommendations or guidelines exist worldwide for concentration of PAHs.

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Keywords

Polycyclic aromatic hydrocarbon · Toxicity · Carcinogenicity · Ecotoxic Effects · Ecosystem risk

Introduction Nowadays growth of large cities are associated with problems such as segregation, neighborhood degradation, increased road traffic, socioeconomic deprivation, and inequities in health, well-being, and health-care accessibility, which have become central environmental as well as political issues in most countries (Irene et al. 2003). The urban environment is strongly man-made and intermittently depreciated in comparison to the characteristics of the natural environment where the respective city developed (Loghin and Murtoreanu n.d.). Such typical urban developments at the cost of environmental degradation lead to emergence of many environmental issues including air, water, and soil pollution (Yang et al. 2010). The continuous addition of motorization results in a sharp increase in the concentration of pollutants in the urban air which subsequently adds to soil, street dust, water, and sediment (Shuang et al. 2011). It is estimated that vehicular emission attributes about 40–80% of the air quality crisis (Ghose et al. 2004). The direct consequence of all these developments is degradation of environmental quality. This in turn affects not only the individual but also the whole community in the big cities (Loghin and Murtoreanu n.d.). Studies also reveal a close relationship of human health effect with the actual level of exposure to pollutants (Wei and Chapman 2001). Extensive research is already being performed in many such areas including inorganic and organic pollutants in the environment with major developments in case of organic pollutants in the last few decades (Gaga 2004). Since industrialization, many organic chemicals that are naturally occurring (e.g., petroleum) and synthetic are released into the environment from various anthropogenic activities. The presence of more than 600 organic compounds have already been reported in various matrices of the environment that belong to the classes of petroleum hydrocarbons, polycyclic aromatic hydrocarbons, ketones, aldehydes, volatile organohalogen compounds, monocarboxylic acids, pesticides, alcohols, dicarboxylic amines, fatty acids, saccharides, and amino acids (Polkowska et al. 2000). Petroleum or refined petroleum products are the largest source of contamination in terrestrial, marine, and groundwater (Douglas and Uhler 1993).

Polycyclic Aromatic Hydrocarbon The environmental contamination mainly by polycyclic aromatic hydrocarbons (PAHs) is a major threat to human health (Skupinska et al. 2004). PAHs are among the most notorious semivolatile organic pollutants and are considered as hazardous air pollutants (HAP) in the group of the non-halogenated organic

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compounds along with benzene, phenols, aldehydes, etc. (IARC 1983).PAHs are ubiquitous in the environment, either as natural component (i.e., as products of humus conversion by microorganism) or as pollutants (i.e., in dust emitted by carbo- and petrochemical industry, in cigarette smoke, as a product of incomplete combustion of organic materials, in particular during waste utilization and house heating) (Bojakowska and Sokołowska 2001). They are the class of compounds with two or more fused benzoid rings made up of only carbon and hydrogen (Velasco et al. 2004) in linear, angular, or clustered arrangements (Lundstedt 2003) (Fig. 1). Based on the molecular structure, PAHs are commonly classified into two categories, namely, low-molecular-weight (LMW) PAHs with four or fewer aromatic rings and high-molecular-weight (HMW) PAHs with five or more rings. The United States Environmental Protection Agency

naphthalene

Phenanthrene

acenaphthylene

Fluorene

Acenaphthene

Fluoranthene

Anthracene

Pyrene

Benzo(a)anthracene

Chrysene

Benzo(b)fluoranthene

Benzo(k)fluoranthene

Benzo(a)pyrene

Dibenzo(a,h)anthracene

Benzo(g,h,i)perylene

Fig. 1 Molecular structures of PAHs

Indeno(1,2,3-cd)pyrene

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(USEPA) has promulgated 16 unsubstituted PAHs as the most priority ones to be analyzed in different environmental matrices (US EPA 1977). Among them, benzo (a)anthracene (BaA), benzo(a)pyrene (BaP), benzo(b)fluoranthene (BbF), benzo(k) fluoranthene (BkF), chrysene (Chry), dibenzo(a)anthracene (DBA), and indeno (1,2,3-c,d)pyrene (IP) have been classified as possible or probable human carcinogens (US EPA 2002), and BaP is often chosen as a surrogate of the whole group of PAH because of its recognition as being carcinogens (IARC 1987). Their fate in the environment is controlled by their physicochemical properties, especially nonpolarity and hydrophobicity triggering their persistence nature in different environmental matrices (Srogi 2007). They are prone to bioaccumulation and biomagnification coupled with long-range transport (Liu et al. 2017). Due to the potential nature of carcinogenicity, mutability, and toxicity, PAHs have been listed as contaminants that required monitoring (Liu et al. 2017; Balcioglu et al. 2014; Wu et al. 2008). PAHs concentration levels have been set by the EU (Commission E. 2013) (7), USEPA (US Environmental Protection Agency Regional Screening Levels n.d.), the Netherlands (Ministry of Housing Spatial Planning and Environment 1994), and China (Ministry of Housing and Urban-Rural Development 2009). Depending on volatility and molecular weight, PAHs can adsorb on soot surface and can remain in the gas phase as well (Sánchez et al. 2013). Such particles then appear in different environmental matrices such as ambient air, soil, street dust, water, sediment, etc. and can be inhaled by human beings or consumed with food, leading to major health problems, such as tumors, birth defects, and a variety of pulmonary diseases (Kislov et al. 2013).

Molecular Structure and Physicochemical Properties of PAHs In 1976, over 100 different PAHs found in the atmosphere were recognized, and in the year 1981, in cigarette smoke, more than 200 PAHs were diagnosed (Lee et al. 1976, 1981). Many PAHs contain the same number of rings, but their differences in configuration lead to differences in the compound’s characteristics (Skupinska et al. 2004). Molecular structure of 16 PAHs is shown in Fig. 1, and physical properties of the 16 PAHs are shown in Table 1. The typical features of PAHs are high melting and boiling points (therefore they are solid), low vapor pressure, and very low aqueous solubility (Masih et al. 2012). Vapour pressure and aqueous solubility of PAHs tend to decrease with increasing molecular weight, on the contrary, PAHs are resistance to oxidation and their reduction increases with increasing molecular weight (US EPA 2002). Thus, aqueous solubility of PAHs reduces for each additional ring (Masih et al. 2010). PAHs dissolve effectively in organic solvent and are lipophilic in nature which is measured by water-octanol partition coefficients (Kow). All PAHs are solid having high melting and boiling points. Although PAHs are chemically inactive, they easily bond to particulate matter. PAHs become highly thermo- and photosensitive when adsorbed on dust surface and degraded at high temperature (50  C) as well as on exposure to light especially ultraviolet and visible light (Zakrzewski 1995). Photooxidation is one of the principal pathways of PAH decomposition and

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Table 1 Physical properties of 16 PAHs (Source: WHO 1998a) Compound Naphthalene Acenaphthylene Acenaphthene Fluorene Phenanthrene Anthracene Fluoranthene Pyrene Benzo(a) anthracene Chrysene Benzo(b) fluoranthene Benzo(k) fluoranthene Benzo(a)pyrene Dibenzo(a,h) anthracene Benzo(g,h,i) perylene Indeno(1,2,3cd)pyrene

Molecular weight 128.18 152.20 154.20 166.23 178.24 178.24 202.26 202.26 228.30

Melting point  C 80.2 92–93 90–96 116–118 96–101 216–219 107–111 150–156 157–167

Boiling point  C 218 265–280 278–279 293–295 339–340 340 375–393 360–404 435

Vapor pressure kPa 1.1  102 3.9  103 2.1  103 8.7  105 2.3  105 36  106 6.5  107 3.1  106 1.5  108

Solubility in water (mg/L) 3.93 3.93 1.93 1.68–1.98 1.2 0.076 0.2–2.6 0.077 0.01

228.30 252.32

252–256 167–168

441–448 481

5.7  1010 6.7  108

0.0028 0.0012

252.32

198–217

480–471

2.1  108

0.00076

252.32 278.35

177–179 266–270

493–496 524

7.3  1010 1.3  1011

0.0023 0.0005

276.34

275–278

525

1.3  1011

0.00026

276.34

162–163

530

Ca.1011

0.062

their secondary metabolite production in the atmosphere (Skupinska et al. 2004). In the environment, substituted PAHs with functional groups such as –OH, -NO2, =O, and –CH3 are also identified (Skupinska et al. 2004).

Exposure to PAHs Human exposure to PAHs occurs through three possible ways which include respiratory tract, gastrointestinal tract, and skin contact. For a nonsmoking person, up to 70% of PAH exposure can be associated with diet. The main sources of PAHs in diet include cereals, oils, and vegetables. Cooked food (in particular food prepared over open flame) is recognized as the major contributor of PAH intake in humans, for example, in barbecued meat, the level of PAH can be as high as 10–20 μg/kg (Phillips 1999). Water is another very important source of PAHs. The highest acceptable concentration of BaP in water is o.7 μg/L as recommended by WHO. It is estimated that the mean intake of PAH with water is 1% of total acceptable level (WHO 1998a). A primary factor of risk due to PAHs is smoking habit. It is estimated that having smoked one cigarette can cause intake of 20–40 ng of BaP (Phillips 1996; O’Neill 1997). Considering the population density, the human exposure risk to atmospheric

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PAHs is maximum in the cities (Caricchia et al. 1999). Although there is a shift in the usage of coal toward other fossil fuel (oil or natural gas) for domestic heating resulting into the drastic reduction of urban particulate pollution in cities, the densification of the urban net coupled with increasing importance of traffic has contributed to reinforce urban particulate pollution. Furthermore, particles generated from cars are much smaller than coal particles and are mostly confined in the breathable size fraction (Manoli et al. 2002). There is much information on the HMW PAHs, but studies are comparatively very less for the vapor-phase LMW PAH components. Although these LMW PAH compounds are considered to have weaker carcinogenic/mutagenic properties, they are the most abundant in the urban atmosphere and react with other pollutants to form more toxic derivatives (secondary pollutant) (Park et al. 2002). Thus, it can be inferred that human exposure to mixtures of PAHs, rather than to individual substances, is important. Industrial workplaces such as coke plants; coal tar and pitch producing and manufacturing industries; aluminum plants; iron and steel foundries; and creosote, rubber, mineral oil, soot, and carbon black producing or manufacturing companies are also associated with high exposure to PAHs. Such unusually high exposure to PAHs was revealed after a significant increase in the incidence of certain cancer diseases and was reported among the workers. Similarly, chimney sweeps, roadmen (pavement tarring), and roofers (roof tarring) are under increased risk, included among the highly exposed occupational groups (Jacob and Seidel 2002).

Toxicity and Carcinogenicity of PAHs In the eighteenth century, higher rate of occurrence of skin cancer was reported among roofers who were exposed to soot. Again in the year 1947, lung cancer was found to be associated with working condition of gas industry workers and that of coal tar (Kennaway 1995). It was then revealed that the induction of cancer was caused by PAHs available in coal tar and soot (Kjaerheim 1999). In 1983, the International Agency for Research on Cancer recognized 30 PAHs as carcinogenic to people. PAHs attract considerable attention among researchers with the continuous rise in death toll caused by cancer which account for about 13% of all human deaths worldwide in 2007 (Jemal et al. 2011). Carcinogenic and mutagenic properties of a few PAH members have already been well established by researchers (Grimmer 1983; Ramdahl and Bjorseth 1985; Dias 1987; U.S. EPA (Environmental Protection Agency) 1993; Clement Associates 1990; Department of Health and Human Services, Public Health Service, National Toxicology Program 2011; U.S. Environmental Protection Agency n.d.; ATSDR (Agency for Toxic Substances and Disease Registry) 1995; IARC (International Agency for Research on Cancer) 1996; IPCS/WHO (International Programme on Chemical Safety/World Health Organization) 1998; Boström et al. 2002; Larsen and Larsen 1998; Bartoszek 2002; Baek et al. 1991a; Harvey 1998; Howsam and Jones 1998; NRC (National Research Council) 1983) in experimental animals and humans. It is revealed from studies on experimental animals that PAHs may trigger various health effects, such

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as cytotoxicity, immunotoxicity, genotoxicity, carcinogenicity, reproductive toxicity, etc. (IARC 1983). Carcinogenic classifications of selected PAHs put forward by various agencies are given in Table 2. All PAHs do not show the same toxicity depending upon the structure of a particle and substituted groups of PAHs. Many PAHs belong to the group of carcinogens such as unsubstituted PAHs, the PAHs containing nitrate, methyl, and carboxylic group (Skupinska et al. 2004). Benzo(a)pyrene (BaP) is termed as index or “gold standard” of the whole group of PAHs due to its high carcinogenic potency (U.S. EPA (Environmental Protection Agency) 1993, n.d.). Carcinogenicity of BaP through inhalation was tested only in hamsters (Thyssen et al. 1981; WHO 2000). Risk analysis inferred that about 9 per 100,000 individual spending a lifetime in ambient air exposed to average level of BaP of 1 ng/m3 may die from respiratory tract cancer (Larsen and Larsen 1998; WHO 2000). Car exhausts (petrol, diesel), domestic coal-stove emissions, and tobacco smoke releasing mainly four- to seven-ring PAHs are found to exhibit nearly all the carcinogenic potential (Pott and Heinrich 1990). Elevated levels of DNA adduct, mutations, reproductive effects, and cancers of lung, respiratory tract, and urinary bladder are found to be associated with exposure of PAHs (Bosetti et al. 2007; Gaspari et al. 2003). According to Pankow et al. (Pankow et al. 1993), carcinogenic risk of inhaled PAHs depends upon whether it enters into lung in gas form or particulate; in case of particulate, PAH carcinogenicity prevails for longer period of time. Mutagenic activities of 67 PAHs were studied in human lymphoblastoid by Durant (Durant 1996) in which human lymphoblastoids were grown in the presence of different concentrations of the studied PAHs. According to the study, the mutagenic activities of PAHs are found to be as follows: dibenzo(a,l)pyrene > benzo(a)pyrene > indeno(1,2,3-c,d)pyrene > dibenzo(a,h)anthracene > benxo(b) fluoranthene > benzo(g,h,i)perylene > antraquinone >9-nitroanthracene > ben(e) pyrene phenanthrene and pyrene. In case of benzo(a)pyrene and dibenzo(a,l) pyrene, it was revealed that the change in the number of rings results into difference in toxicity of those PAHs. It was also observed that PAH’s biological activities were not only controlled by the number of rings but also the shape, the dimension of particles, and the presence of functional groups. In the studies, it was observed that some PAHs are even stronger carcinogens than BaP. Nevertheless, BaP is still used as an indicator of whole group of PAHs (Phillips 1999). It is critical to establish the levels of PAHs that are safe for humans. The highest acceptable levels of PAH concentrations in various countries are presented in Table 3 (Kjaerheim 1999; Khesina 1994; Slooff et al. 1989; WHO 1998b; European economic community 1980; Disposition of German Federal Department for worker safety 1989; Disposition of Swedish National Board of Occupational Safety and Health 1994; American Conference of Governmental Industrial Hygienists 1995). In Poland since 2001, the maximum acceptable level in a workplace has been set to be 2 μg/m3 (Corpus of Polish Law (Dziennik Ustaw) 2001). According to the Ministry of Environment in Great Britain Ascertained in the Air Quality norm, it is not possible to estimate the absolutely safe level of exposure to carcinogens (in particular PAH) (UK Department for Environment, Food, and Rural Affairs 1999). In reality, there is no such idea as safe concentration for carcinogenic and

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Table 2 Carcinogenic classifications of PAHs by various agencies Agency US Department of Health and Human Services (HHS)

International Agency for Research on Cancer (IARC)

US Environmental Protection Agency (EPA)

PAH compounds Benzo(a)anthracene Benzo(b) fluoranthene Benzo(a)pyrene Dibenzo(a,h) anthracene Indeno(1,2,3-cd) pyrene Benzo(a)anthracene Benzo(a)pyrene Benzo(a) fluoranthene Benzo(k) fluoranthene Indeno(1,2,3-cd) pyrene Anthracene Benzo(g,h,i) perylene Benzo(e)pyrene Chrysene Fluoranthene Fluorene Phenanthrene Pyrene Benzo(a)anthracene Benzo(a)pyrene Benzo(b) fluoranthene Benzo(k) fluoranthene Chrysene Dibenzo(a,h) anthracene Indeno(1,2,3-cd) pyrene Acenaphthylene Anthracene Benzo(g,h,i) perylene Fluoranthene Fluorene Phenanthrene and Pyrene

Carcinogenic classification Known animal carcinogens

Probably carcinogenic to humans Possibly carcinogenic to humans

Not classifiable as to their carcinogenicity to humans

Probable human carcinogens Not classifiable as to their carcinogenicity to humans

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Table 3 Maximal permissible concentrations (MPC) of PAHs in selected countries Country year Italy 1999 Former USSR 1985 EEC 1980

WHO 1995 EEC 1980

Occurrence Ambient air Ambient air Ambient water

Drinking water Drinking water

Compound Benzo(a)pyrene

MPC 1 ng/m3 (Kjaerheim 1999)

Benzo(a)pyrene

1 ng/m3 (Khesina 1994)

Sum of fluoranthene, Benzo(b) fluoranthene, Benzo(k) fluoranthene, Benzo(a)pyrene, Benzo(g,h,i) perylene, Indeno(1,2,3-c, d)pyrene Benzo(a)pyrene

1.2 μg/l (Slooff et al. 1989)

Germany 1989

Oven area

Sum of fluoranthene, Benzo(b) fluoranthene, Benzo(k) fluoranthene, Benzo(a)pyrene, Benzo(g,h,i) perylene, Indeno(1,2,3-c, d)pyrene Benzo(a)pyrene

Sweden 1993

Workplaces

Benzo(a)pyrene

USA 1993

Workplaces

Pyrene

0.7 μg/l (WHO 1998b) 0.2 μg/l (European economic community 1980)

2 μg/m3 (Disposition of German Federal Department for worker safety 1989) 2 μg/m3 (Disposition of Swedish National Board of Occupational Safety and Health 1994) 0.1 mg/m3 (American Conference of Governmental Industrial Hygienists 1995)

mutagenic substances. For such substance, even a very small amount can increase the risk of neoplastic diseases by accumulating continuously for years after entering the living organism (Skupinska et al. 2004).

Sources of PAHs Generation of PAHs is associated with incomplete combustion of materials containing carbon and hydrogen such as coal fuel, crude oil, wood, gas, and organic materials as well as combustion of polypropylene and polystyrene, communal and

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industrial waste, and used tires (Cherng 1996; Lewitas et al. 1997). The process of combustion is considered to be incomplete when the temperature of combustion is low and atmosphere is anaerobic. Free radicals initiate the chemical synthesis of hydrocarbon in the flame. Typically, free radicals of high molecular weight lead to the production of large particles of PAHs although methane can also give on to generation of PAHs particles. The aromatic compounds and diolefins can be precursors of PAHs (Skupinska et al. 2004). Although some PAHs in the environment are released from natural sources of combustion such as forest fires and volcanoes along with some minor biogenic sources, emissions from anthropogenic activities predominate (Baek et al. 1991a; Harvey 1998; Finalyson-Pitts and Pitts 1986). Anthropogenic sources of PAHs can be broadly classified into stationary and mobile sources. Stationary sources of PAHs include residential heating, industrial activities (e.g., aluminum production and coke manufacture), incineration, and power generation. Diesel and gasoline engine vehicular exhaust are categorized as mobile sources of PAHs (Gaga 2004). In general, stationary fuel sources are considered to be responsible for over 97% of PAH emissions (Pikes 1992). However, the profile is different in case of an urban area with high traffic density (Gaga 2004). According to Shen et al. (2013), global total atmospheric emission of USEPA-listed 16 priority PAHs was 504 Gg in 2007 with maximum contribution of about 60.5% from biomass fuels combustion, mainly firewood and crop (Samburova et al. 2017). Some of the examples of natural sources of PAHs are forest and brush fires, volcanoes, bacterial and algal synthesis, petroleum seeps, erosion of sedimentary rocks containing petroleum hydrocarbons, and decomposition of vegetative liter fall (Zhang and Tao 2009). Anthropogenic sources of PAHs can also be categorized as petrogenic and pyrogenic. Pyrogenic PAHs resulted from incomplete combustion of any organic material, while the petrogenic PAHs are found in oil and some of its product (Feng et al. 2009; Lang et al. 1962, 1964). In general, the pyrogenic PAHs are comprised of larger ring than the petrogenic PAHs (Pampanin and Sydnes 2013). Pyrogenic PAHs are formed by the process of pyrolysis (transformation of organic substances under high temperatures and low oxygen or no oxygen conditions) such as destructive distillation of coal into coke and coal tar or the thermal cracking of petroleum residuals into lighter hydrocarbons that occur intentionally. On the other hand, other unintentionally occur processes include the incomplete combustion of motor fuels in cars and trucks, the incomplete combustion of wood in forest fires and fireplaces, and the incomplete combustion of fuel oils in heating systems. The pyrogenic processes occur in the temperature ranging from about 350  C to more than 1200  C. Higher concentrations of pyrogenic PAHs are typically found in urban areas mainly close to major sources of PAHs. Moreover, PAHs can also be generated at lower temperatures. PAHs are produced during crude oil maturation, and similar processes are termed as petrogenic. Such petrogenic PAHs are released during transportation, storage, and use of crude oil and crude oil product such as oceanic and freshwater oil spills, underground and aboveground storage tank leaks, and the accumulation of vast numbers of small releases of gasoline, motor oil, and related substances associated with transportation (Tolosa and Bayona 1996; WHO (World Health Organization) 2003; Masih and Taneja 2006; Seo et al. 2007).

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There are numerous and well-known sources of PAHs in the environment depending upon the prevailing activities, atmospheric conditions, and topography of the area. Analysis of real environment samples helps in the identification of the PAHs. Differentiation between groups of PAH sources (coal-based, wood-based, or oil-based sources) is performed by chemical fingerprinting with several techniques. This can be performed by studying the behavior of different chemical indicators that exist in the environmental samples. This source identification technique can also be used to identify and assign nonpoint sources of PAHs to the environment irrespective of their land-use pattern (Wang et al. 2011a). Temperature of formation impart one fundamental information in identifying PAH sources due to the fact that higher temperature of formation tend to generate PAHs with fewer alkylated chains than that under lower temperature (Hussein et al. 2016). PAH dispersion is regulated by thermodynamic properties during low-temperature processes and by kinetic characteristics during high-temperature processes, such as pyrolysis of organic matter (Alberty and Reif 1988). As for example, phenanthrene is thermochemically more stable than anthracene; molar fraction of phenanthrene produced is much higher than that of anthracene at low temperature. On the other hand, high-temperature combustion such as incomplete burning of organic matter is characterized by low value of phenanthrene/anthracene ratio (Baumard et al. 1998).

Distribution in the Environment (Air, Soil, Street Dust, Water, and Sediment) PAHs are well-dispersed environmental pollutant arising from anthropogenic sources (Baek et al. 1991a; Nollet 2007; Petry et al. 1994, 1996). Aerial movement is one of the crucial pathways for environmental distribution and transboundary deposition of PAHs (Birgül et al. 2011). Due to their long-range transport, remote areas far from emission sources also get exposed to PAHs (Montelay-Massei et al. 2007a; World Health Organization 2002). Eventually, PAHs settle down in soils and street dust and enter into aquatic ecosystem (Christensen and Arora 2007; Tolosa et al. 2004). Soil and street dust behave as direct sink of atmospheric PAHs near to combustion sources such as traffic. Rainwater and storm water washed away PAHs from these environmental compartments to nearby aquatic bodies. Due to hydrophobic nature, PAHs in aquatic environment are preferably partitioned and accumulate into the particulate phase of sediment (Gaga 2004). PAHs, thus, occur in multicompartmental system of the environment (Holoubek et al. 2000; Finlayson-Pitts and Pitts 2000) and paved the way for multiple routes of exposure to this class of carcinogen. Figure 2 illustrates distribution of PAHs in various environmental matrices.

PAHs in the Atmosphere It is fully established that atmospheric transfer is the principal pathways for the global distribution of PAHs (Birgül et al. 2011). Once in the atmosphere, depending

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Fig. 2 PAHs distribution via air, terrestrial and aquatic environment

on their physical and chemical properties (vapor pressure, Henry’s law constant, and solubility), PAHs get distributed between gas, particle, and droplet phase (Junge 1977; Bidleman 1988; Larsen III and Baker 2003; Gocht et al. 2007). Most of the studies revealed that the two- and three-ring PAHs are mainly found in the vapor phase, while four- to six-ring PAHs occur in the particle phase at ambient temperature, (Zhang et al. 2009; Teixeira et al. 2013; Delgado-Saborit et al. 2013; Alam et al. 2013). PAH’s wash out mechanism from atmosphere occurs through two major pathways: oxidative and photolytic reactions and atmospheric fallouts (Garban et al. 2002; Manoli et al. 2000). Atmospheric bulk (dry+wet) deposition is another mechanism associated with the removal of vapor as well as particle-bound PAHs in the air according to Bidleman et al. (Bidleman 1988). Particle-phase concentration and meteorological factors play a very crucial role in deposition of PAHs (Liu et al. n.d.). Extensive studies have been carried out to explore relationship between meteorological parameters and PAH concentrations. Remarkable relationship of PAHs with temperature and relative humidity were documented by Pankow et al. (1993) and Chetwittayachan et al. (2002). It was also observed that temperature variation has more impact on gas-phase LMW PAH dispersion than particulatephase HMW PAHs (Sofuoglu et al. 2001). In most of the studies, PAH concentrations were observed with minimal value during summertime or monsoon than that of winter, due to the fact that there is increase in the mean inversion height, decrease in the number of inversion days during winter, and lack of a major PAH source and residential fuel combustion for heating (Baek et al. 1991b; Hussain et al. 2016a). Another observation also revealed

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that maximum concentration of particle-associated PAH was found during winter, while vapor-phase PAHs were maximum during summer (Baek et al. 1991b). Considering the risk of human exposure to atmospheric PAHs, it was revealed from studies that exposure risk to atmospheric PAHs is highest in cities due to increasing density of population, vehicular traffic, and insufficient dispersion of the atmospheric pollutants (Rockens et al. 2000). The prescribed or mandatory concentration level of BaP is 1 or 10 ng/m3 air in various countries (e.g., Italy or Germany) (Jacob and Seidel 2002).

Gas to Particle Distribution of PAHs in the Atmosphere PAHs exist in both vapor and particulate phase [mainly the fine fraction (5 rings are exposed to relatively high-temperature condensation and are mostly adsorbed to atmospheric particles (Van Jaarsveld et al. 1997; Wania and Mackay 1996). These PAHs are often subjected to rapid deposition and retention close to the source and are therefore listed as “single-hop” persistent organic pollutants (POPs) due to their low mobility (Wania and Mackay 1996; Yang et al. 1991; Agarwal 2009). On the contrary, the LMW PAHs with two to three rings undergo low-temperature condensation and found mostly in the gas phase (Subramanyam et al. 1994; Van Jaarsveld et al. 1997; Wania and Mackay 1996; Valerio et al. 1984; Wild and Jones 1995). This category of PAHs is termed as “multi-hop” chemicals due to their high or moderately high mobility (Wania and Mackay 1996). As a result, LMW PAHs diffuse in the air worldwide and preferentially cumulate in polar latitudes (Van Jaarsveld et al. 1997; Wania and Mackay 1996). Semivolatile (fourring) PAHs can exist both in particle and gas phase and are mostly induced by environmental factors (Subramanyam et al. 1994). They get preferably settled and enriched in midlatitudes (Wania and Mackay 1996). Thus, individual PAH depending upon their molecular weight and structure functions differentially under different meteorological and climatic conditions. Meteorological parameters (temperature and relative humidity) play a crucial role in diffusion of LMW PAHs (Pankow et al. 1993; Zielinska et al. 2004). Vapor pressure influencing

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volatilization has been found to be governed by temperature, humidity, and wind speed. With the increase of temperature, PAH volatilization from air particles, soil, vegetation, and water also increases (Sofuoglu et al. 2001; Ausma et al. 2001). There is also a report of rise in absorption of semivolatile organic compounds as relative humidity approaches 100% (Thibodeaux et al. 1991). Such increment at very high relative humidity is associated with partitioning into the liquid or nearly liquid water when the gas phase is nearly saturated with water (Thibodeaux et al. 1991). Significant correlation between vapor pressure and molecular weight with a correlation coefficient of 0.9017 was also reported (Schwarzenbach et al. 1993). As for example, PAHs with lower vapor pressures such as BaP will prefer to be bind to particles, while PAHs with higher vapor pressures such as naphthalene will tend to be sorbent with the vapor phase. Thus, for atmospheric samples, the relative dispersion of PAHs in the two phases will be distinct. According to the atmospheric study in Portland, Oregon, conducted by “Electric Power Research Institute (EPRI)” (EPRI (Electric Power Research Institute) 2000), the sum of PAH concentration for the vapor phase (741 ng/m3) was much higher than that of the particulate phase (12 ng/m3). Moreover, it was also revealed in the study that the majority of LMW/higher vapor pressure PAHs were identified in the vapor phase, whereas the HMW/lower vapor pressure PAHs were detected with much lower concentrations (EPRI (Electric Power Research Institute) 2000). The amounts of dust in the air and PAH concentrations in the particulate phase were also found to exhibit significant correlation (Kuo et al. 2013). As a result of which, the levels of gas-phase PAHs are dominant in summer or typically in tropical regions, whereas particulate-phase PAHs are found to be prevalent during winter or typically in Arctic regions (Lai et al. 2011; Mohanraj et al. 2012). Furthermore, types of suspended particulates such as soot, dust, fly ash, pyrogenic metal oxides, pollens, etc. also influence PAH adsorption (Zhang and Tao 2009).

Air Water Gas Exchange of PAHs Air-water transfer processes for PAHs include volatilization and absorption of gases, dry deposition with particles, and wet deposition by rain or snow, i.e., particle and gas “washout,” spray transfer, and bubble scavenging (Gustafson and Dickhut 1997). Air-water transfer is regulated by Henry’s law, according to which PAH concentration in water is proportional to the partial pressure of PAHs in air: pa ¼ H a X w where pa is the partial pressure in air, Ha is Henry’s constant (both in pressure units), and Xw is the mole fraction in water. The lower is the Henry’s constant, it is more preferable that PAHs will partition from air to water (Fisher 2001).

Atmospheric Pollutant Removal Processes Anthropogenic activities are basically liable for release of chemical pollutants into the environment. If all the pollutants persist in various environmental compartments as it is, their ambient levels should equal to the aggregate amounts

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discharged. However, typically this is not the case, as each environmental matrix in nature has some active naturally occurring removal or “self-purification” processes (Schlesinger 1979). Gas- and particle-associated PAHs in air are subjected to removal mechanism such as oxidative and photolytic reactions and ultimately from the atmosphere by two mechanisms, namely, wet and dry deposition (Garban et al. 2002). Removal of pollutants from the atmosphere in the absence of precipitations onto various surfaces of earth is termed as dry deposition. Whereas, the scavenging processes by which pollutants are encapsulated by hydrometeors and consequently dropped to the Earth’s surface is referred as wet deposition (Seinfeld and Pandis 1998). Thus, the released airborne PAHs are moved by the prevailing meteorology before being washed from the atmosphere by different scavenging processes. Wet deposition (rain, sleet, snow, hail) and dry deposition are considered as the major mechanisms of removal of PAHs from the atmosphere. Studies revealed that the wet removal pathway of gaseous PAHs was better understood than particulate PAHs (Ligocki et al. 1985a). The scavenging mechanism, such as in-cloud or below-cloud, collection efficiency of falling precipitation, solubility, and size particles have been well investigated in literature (McVeety 1986; Sharma and Patil 1994). Nonreactive gaseous organic compounds are washed out by rain between the gas and aqueous phases according to the Henry’s law equilibrium (Ligocki et al. 1985a, b). According to Pankow et al. (1984), the total degree of scavenging of a given compound when there is no exchange of material between the particulate and dissolved phases in the rain can be expressed as follows: W ¼ W g ð1  ϕÞ þ W p ϕ where W is the overall scavenging ratio: W ¼ ½rain, total =½air, total Wg is the gas scavenging ratio: W g ¼ ½rain, dissolved =½air, gas Wp is the particle scavenging ratio: W p ¼ ½rain, particulate =½air, particulate and ϕ is the fraction of the atmospheric concentration which is associated with particles. For compounds which (1) are scavenged to some degree from the atmospheric particulate phase and (2) remain on particulate material inside the raindrop, W will differ from the equilibrium Wp value. The second state become necessary as material which is moved to the dissolved phase will undergo rapid re-equilibration with the atmosphere.

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Fig. 3 Sources, transport, and deposition of PAHs via the atmosphere

Dry deposition is a function of prevailing atmospheric state and the concentration of PAHs in surface level. PAHs adsorbed particle size greater than 20 μm are vulnerable to settle near the source due to their high-settling velocity. However, this PAH removal mechanism occurs only to minor percentage as PAH is predominantly adsorbed on particle diameter of 1 during both post-monsoon and pre-monsoon seasons. Value RQ(NCs) of PAHs was also >800 during both the studied season. This signified high ecosystem risk of Bharalu River water during both the season (Hussain 2014; Hussain et al. 2014). Moreover investigation was also performed in Bharalu River sediment during the same season, and it was observed that RQ(MPCs) of PAHs was >1 during both post-monsoon and pre-monsoon season. RQ(NCs) of PAHs during post-monsoon was >800 while during pre-monsoon was 2.3

Ni ≤5.5 5.6– 9.3 9.4– 13.1 >13.1

Cu ≤7.1 7.2– 11.6 11.7– 16 >16.0

Zn ≤31.7 31.8– 45.1 45.2– 58.6 >58.6

Fe ≤180 181– 309 310– 438 >438

V ≤0.32 0.33– 0.49 0.50– 0.65 >0.65

3 (high), and class 4 (very high), as shown by Table 1. With respect to PAHs, the results are usually compared with those of other studies that monitored these pollutants in similar areas. Table 2 presents studies with Lolium sp.

Monitoring of Atmospheric Pollutants Particulate Matter Sampling MP can be sampled using dichotomous samplers, which are devices capable of segregating the PM into coarse and fine fractions (Fig. 8). The dichotomous samplers are generally equipped with atmospheric pressure and temperature sensors, to allow it to control the volumetric inflow rate, as well as a computer, for control of the system and storage of data. The air sample enters a fractionating inlet, usually at a flow rate of 1 m3 h1, i.e., 16.7 L min1. Due to the axial symmetry shape of the inlet, the sampling efficiency is independent of wind direction. A coarse screen

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Table 2 Studies with Lolium sp. Site Cubatão region, Brazil

Species Lolium multiflorum italieum cv. Lema Italian rye grass (Lolium multiflorum “Lema”)

Evaluated pollutants Fluor, sulfur

Lolium perenne

Fluoride

Cubatão, Brazil

Lolium multiflorum‘Lema

São Paulo, Brazil

L. multiflorum Lam. sp. italicum Beck cv. Lema Lolium multiflorum sp. italicum cv. Lema. Lolium multiflorum

Naphthalene, acenaphthene, fluorene, phenanthrene, anthracene, fluoranthene, pyrene, benzo[a]anthracene, chrysene, benzo[b] fluoranthene, benzo[k] fluoranthene and benzo[a] pyrene. Fructose

Edinburgh, Sheffield, Copenhagen, Düsseldorf, Nancy, Hohenheim, Klagenfurt, Lyon, Verona, Barcelona, Valencia, Europe Galicia, Spain

Sinos River basin, Southern Brazil

Tropical areas of Southeast Brazil, South America Southern of Brazil, South America

Lolium multiflorum sp. italicum cv. Lema.

Sulfur, Fe, Cu, Zn, Pb, Cd, Cr, Ni, Sb, As, and V

Al, Ba, Cd, Cu, Pb, Cr, Fe, Mn, Ni, and Zn

N, S, Mg, Fe, Mn, Cu and Zn Al, Ba, Cd, Cu, Pb, Cr, Fe, Mn, Ni, and Zn

Reference Klumpp et al. (1994) Klumpp (2004)

ReyAsensio and Carballeira (2007) Rinaldi et al. (2012)

Sandrin et al. (2013) Alves et al. (2015)

Bulbovas et al. (2015) Illi et al. (2017)

prevents the drag of large particles and insects. Thus, the particles smaller than 10 μm pass through an inlet tube, which serves to straighten the flow, and then, into the virtual impactor. The virtual impactor is the basis for the operation of a dichotomous sampler. The sample flow entering the virtual impactor is split into two separate flow systems. Particles are accelerated through a nozzle and enter the fractionation zone. Due to the geometry of the acceleration nozzle, only 10% of the total flow, which enters the impactor (100 L h1), goes through the collection nozzle. Particles larger than 2.5 μm (PM2.5–10) are drawn directly into the collection nozzle, due to their greater

45

Integrated Evaluation of Quantitative Factors Related to the. . .

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PM10 inlet

Virtual impactor

PM2.5

PM2.5-10 Mass Flow Controller

Flow 900 L h−1

Flow 100 L h−1

Filter holder

Pump

Exhaust Temperature measurement

Fig. 8 General flow scheme of a dichotomous sampler (Source: Adapted from MCZ 2017)

inertia. These particles “impact” into the low flow region of the collection tube and then deposited onto a membrane filter. The remaining 90% of the flow (900 L h1), which contains the smaller particles (PM2.5), passes around the nozzle and through a tube, and the particles are collected in a separate membrane filter (MCZ 2017).

Research Results: Quantitative Factors A study conducted by Alves et al. (2015) in the lower part of the Sinos River basin (SRB), Southern Brazil, between October 2013 and March 2014, assessed the air quality through the determination of the concentrations of MP2,5-10, MP2.5 and the metallic elements Al, Ba, Cd, Cr, Cu, Fe, Mn, Ni, Pb, Zn, and Hg in the leaf area of the Lolium multiflorum. The mean concentrations of MP2.5–10 and MP2.5 were 9.1 μg m3 and 4.7 μg m3. The concentrations of metallic elements, especially Pb, Cr, and Zn, resulted in Class 4 (very high pollution levels), according to the classification proposed by Klumpp (2004). Atmospheric air genotoxicity studies conducted in urban areas in Caraá, Taquara, Campo Bom, São Leopoldo, Esteio, and Canoas districts located in southern of Brazil and in riparian forest environments in Caraá, Taquara, and Campo Bom in the upper, middle, and lower stretches of the SRB evidenced that the flower buds of T. pallida var. purpurea exposed to urban areas presented MCN frequencies ranging from 2.37 to 5.44; the lowest frequency in the municipality of Caraá, located in the

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upper stretch of the SRB, and the highest in Canoas, in the lower stretch. In the environments of riparian forest, the MCN frequencies ranged from 1.92 to 3.38, with the highest average in the municipality of Taquara, located in the middle stretch of the SRB. The lower MCN frequencies recorded for Caraá may be related to the good state of conservation of the vegetation and the low urbanization in the upper stretch of the SRB. High urbanization and industrialization, as well as the intense vehicular traffic, contribute to the increase of potentially genotoxic atmospheric pollutants. High MCN frequencies in Tradescantia pallida var. purpurea have also been reported in other studies in the SRB. Costa and Droste (2012) observed MCN frequencies ranging from 1.03 to 8.13 in plants exposed in a rural area of the municipality of Novo Hamburgo and in an urban area in the municipality of Est^ancia Velha. Blume et al. (2014) observed MCN frequencies ranging from 4.77 to 8.28 at an urban area of the municipality of Sapucaia do Sul. Cassanego et al. (2015) recorded MCN frequencies ranging from 2.13 to 7.23 in urban areas and from 1.50 to 4.80 in riparian forest environments, located in the upper, middle, and lower stretches of the SRB. These MCN frequencies were caused by pollutants present in the environments monitored. However, it is not possible to relate the results of the MCN frequency with any specific compound, due to the complex mixture of pollutants present in the atmosphere emitted from vehicular traffic and industry. Heavy vehicles, like buses and trucks, are responsible for most of the sulfur and nitrogen oxide emissions (Santos et al. 2015), while light vehicles and mixed-use vehicles, gasoline and alcohol-powered, are the main producers of carbon monoxide and hydrocarbons (Teixeira et al. 2008). Air quality monitoring studies carried out in the municipalities of Canoas, Esteio, and Sapucaia do Sul, in the lower stretch of the SRB revealed the presence of CO, SO2, NOx, PAHs, and PM, which were attributed to vehicle emissions and to industrial processes (Migliavacca et al. 2012; Teixeira et al. 2012). In metropolitan regions with a high incidence of traffic congestion, about 90% of CO emissions come from the burning of fossil fuels by motor vehicles (Teixeira et al. 2008). In addition, studies showed that stressing meteorological conditions might influence the response of the Tradescantia pallida var. purpurea (Savóia et al. 2009; Pereira et al. 2013; Spósito et al. 2015), since the low relative air humidity and high temperatures seem to stimulate the opening of the stomata, thus increasing the absorption and transportation of genotoxic substances to the target cells (Klumpp 2004).

Conclusion An air quality monitoring program that integrates biomonitoring and conventional monitoring techniques of key parameters, such as MP, metallic elements, and PAHs, and that integrates the data obtained from these techniques, is able to provide sufficiently robust conclusions regarding the air quality of an area of interest, such as, for instance, a large urban center. The results obtained from the monitoring of several points, located in strategic sites, allow the definition of priority areas, in

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terms of atmospheric pollution, for which, specific actions can be taken, aiming to reduce the environmental impacts caused by air pollution. The bioassays presented in this chapter can provide information on the areas of risk and also allow a record of the effects of complex mixtures of pollutants on living organisms. The survey of meteorological data, vehicular traffic, and the profile of the industries are also fundamental to verify a possible relation with the response provided by the bioindicators. Considering that the bioindication methods are important tools for the assessment of particular conditions of each environment, the implementation of bioassays as a technique for the management and control of atmospheric pollution is suggested to the public agencies, especially in regions influenced by pollution sources from different nature, with different profiles. Therefore, to carry out an air quality assessment, it is necessary to use methods capable of integrating variables, which are representative of the environmental conditions of the evaluated region. Air monitoring data may assist public management by providing reliable bases and methodologies for environmental control, supporting the adoption of more restrictive environmental policies.

Cross-References ▶ Air Pollution ▶ Air Pollution Prevention Technologies ▶ Monitoring and Risk Analysis of PAHS in the Environment ▶ Status of Particulate Matter Pollution in India: A Review Acknowledgments The authors thank Larissa Meincke for the edition of Figures 6 and 7.

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Characterizing the Cell Surface Properties of Hydrocarbon-Degrading Bacterial Strains, a Case Study

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Hamid M. Pouran, Steve A. Banwart, and Maria Romero-Gonzalez

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Experimental Section and Sphingomonas spp., Sph2 Case Study . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sph2 Growth Condition, and Preparation for Infrared Spectroscopy of Its Cell Surface . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Potentiometric Titration of Sph2 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Modelling Surface Protonation of Sph2 Based on the Titration Data . . . . . . . . . . . . . . . . . . . . . Zeta Potential Measurements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sph2 Biofilm Formation on Hematite . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Results and Discussions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Infrared Spectroscopy of Sph2 Cell Wall . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sph2 Surface Potential in Different Ionic Strengths . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sph2 Potentiometric Titration and the Data Optimization Using ProtoFit . . . . . . . . . . . . . . . . Sph2 Biofilm Formation on the Hematite Surface Under Different Ionic Strengths . . . . . . Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cross-References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Appendix . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Preparing the Hematite Surface for Cell Adhesion Studies (Synthesis, Coating, and Characterization) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Bacterial Strains, Growth Conditions, and Sample Preparation . . . . . . . . . . . . . . . . . . . . . . . . . . . Biofilm Formation Studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Glycosphingolipids . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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H. M. Pouran (*) Faculty of Science and Engineering, University of Wolverhampton, Wolverhampton, UK e-mail: [email protected] S. A. Banwart School of Earth and Environment, University of Leeds, Leeds, UK M. Romero-Gonzalez Department of Geography, University of Sheffield, Sheffield, UK © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_131

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Abstract

This chapter describes some of the most common methods used to characterize the cell surface properties of the bacterial cells. As a case study, the focus of this chapter is on Sphingomonas spp., Sph2, which is a Gram negative and hydrophilic bacterial strain. The species used in this research was isolated from groundwater at a phenol-contaminated site. This hydrocarbon-degrading strain that can participate in bioremediation of polluted environments belongs to Sphingomonadaceae family. This group of bacteria is unique among Gramnegative cells because of having glycosphingolipids (GSL) instead of the lipopolysaccharide (LPS) layer in their cell wall. To characterize this strain, its surface properties were examined using potentiometric titration, modelling surface protonation sites using ProtoFit, zeta potential measurements, and attenuated total reflection Fourier-transform infrared (ATR-FTIR) spectroscopy. There is no published detailed study about cell wall characteristics of Sph2 yet, and this research reports such information for the first time. In addition, to investigate effects of the solution ionic strength on Sph2 adhesion behavior on metal oxides, its biofilm formation on hematite, as the model mineral, was evaluated in three different ionic strengths; 200 mM, 100 mM, and 20 mM. The ATR-FTIR analysis showed that despite the unique cell wall chemistry of Sph2 among the Gram-negative strains, its surface functional groups are similar to other bacterial species. Hydroxyl, carboxyl, phosphoryl, and amide groups were detected in Sph2 infrared spectra. The potentiometric titration results showed that Sph2 PZC is approximately 4.3. Optimizing the titration data based on ProtoFit non-electrostatic model (NEM) provided compatible results to the infrared spectroscopy analysis and four pKa values were identified; 3.9  0.3, 5.9  0.2, 8.9  0.0, and 10.2  0.1, which could be assigned to carboxyl, phosphate, amine, and hydroxyl groups, respectively. Zeta potential measurements

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demonstrated that changing the ionic strength from 200 mM to 20 mM shifts the zeta potential by 20 mV. Direct observation showed that this alteration in the ionic strength coincides with a tenfold increase in the number of Sph2 attached cells to the hematite surface. This could be attributed to both electrostatic interactions between the cell and surface, and conformational changes of Sph2 surface biopolymers. In addition to reporting Sph2 cell wall characterization results for the first time, this study highlights importance of ionic strength in the cell adhesion to the mineral surfaces, which directly influence biofilm formation, bioremediation, and bacterial transport in aqueous systems. Keywords

Sphingomonas spp. · Attenuated total reflection Fourier-transform infrared spectroscopy (ATR-FTIR) · Potentiometric titration · Biofilm formation · Cell surface · Cell wall chemistry · Hematite · Electrostatic interactions · Ionic strength · Biopolymers · Solution chemistry · Attached growth · Interface interactions · ProtoFit

Introduction Bioremediation or clean-up of hydrocarbon contaminated sites relying on the indigenous microorganisms (e.g., bacteria) is a cost effective and environmental-friendly approach. In this method, the indigenous bacteria play the dominant role and convert the contaminants to harmless/less harmful products (Gutman et al. 2014; Pouran et al. 2008, 2014; Ojeda et al. 2008). Bacterial cells, despite their small size, are dynamic (e.g., using different sources of carbon and energy) and capable of adapting to different environmental conditions (O’Toole et al. 2000; O’Toole and Wong 2016). Their lifestyle includes both planktonic and attached forms. Cell attachment, which is followed by biofilm formation, is the prevailing bacterial form (Pouran et al. 2017). It is well established that the bacterial life style and its physiological properties directly influence the biodegradation rate (Andrews et al. 2010). To design better bioremediation processes, and predict attenuation of the contaminants in the affected sites, deeper knowledge of the bacterial cells behaviors in natural environments, particularly biofilm formation, is essential. Sphingomonas is a metabolically versatile bacterial strain and capable of participating in bioremediation of contaminated environments (Chen et al. 2008; Gabriel et al. 2005a, b). For example, we know that Sphingomonas xenophaga Bayram, isolated from the activated sludge of a municipal wastewater treatment plant, can degrade nonylphenol mixtures by utilizing these as a sole carbon and energy source (Gabriel et al. 2005a). In our previous publications, we evaluated biofilm formation of different bacterial strains, isolated from contaminated soils and aquifers, on iron oxide surfaces. Effects of these minerals’ hydrophobicity and pH-dependent surface charge on the attached growth of Rhodococcus spp. (RC92 & RC291), Pseudomonas spp. (Pse1 & Pse2), and Sphingomonas spp. (Sph1 & Sph2) were studied in this research (Pouran et al. 2009,

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2014, 2017; Andrews et al. 2010; Geoghegan et al. 2008). Among these genetically diverse strains, Sphingomonas spp., Sph2, a hydrophilic strain with negative surface charge in the experimental conditions (pH 6.5), showed attachment behavior anomalies contrary to the expected electrostatic attractions between the minerals and the cell surface (Pouran et al. 2017). Sph2 attached most onto hydrophobic, electrostatically neutral polystyrene compared to hydrophilic and positively charged hematite surface after 96 h of incubation (pH 6.5), please see the appendix for more information and details of the experimental conditions. Sphingomonas spp., Sph2 is a Gram negative bacterial species, which was isolated from groundwater at a phenol-contaminated site in the West Midlands (England) (Geoghegan et al. 2008; Andrews et al. 2009). In this research, we further characterize its cell surface properties and try to elucidate the underlying reasons for its attachment anomalies on the previously studied metal oxide surfaces (Pouran et al. 2014, 2017). Bacterial cells are expected to have relatively low point of zero charge (PZC), a specific pH value at which the surface charge is neutral, and therefore have negative surface charge in the pH range of natural environments (Ojeda et al. 2008; Pouran et al. 2014, 2017; Andrews et al. 2010). Nevertheless, the dynamic nature of the bacterial species allow them to adapt with different environmental conditions and capable them of changing their surface chemistry (Claessens et al. 2004, 2006). Presence of different biopolymers and nonuniform surface charge distribution means that the bacterial surface charge is a very specific characteristic that depends on the physiological and environmental conditions (Fein et al. 2005; Hong and Brown 2008). For example, we know that the ionic strength influences surface potential and affects the electrokinetic and surface properties of the bacterial cells (Chen and Walker 2007). Here we aim to identify Sphingomonas spp., Sph2 cell wall functional groups for the first time, determine its surface electrostatic nature, and explain its observed attachment anomalies (tendency to form biofilm on the electrostatically neutral and hydrophobic polystyrene compared to the positively charged hydrophilic metal oxide surfaces) that we have seen in the previous studies (Pouran et al. 2014, 2017). For these purposes we used Infrared spectroscopy, zeta potential measurements, potentiometric titrations, and modelling techniques. In addition, the effects of different ionic strengths on biofilm formation of Sph2 on the hematite surface were studied.

Experimental Section and Sphingomonas spp., Sph2 Case Study In this section, we describe some of the most common techniques used to study the cell surface of bacterial strains. We provide details of the performed experiments for understanding Sphingomonas spp., Sph2 cell wall properties. The rest of this chapter is focused on this bacterial strain as a case study. The chemicals used in this research were certified ACS reagents, chemicals that meet or exceed the latest ACS specifications. Ultra-high quality water (UHQ, conductivity 18.2 MH/cm at 25  C) was used throughout the experiments. By no means the methods mentioned in this

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chapter are exhaustive. They are some of the most frequently used and basic tests that scientists perform to analyze the bacterial cell surface properties. A range of other techniques like Atomic Force Microscopy and Scanning Electron Microscopy are available that could help to elucidate more features of these microorganisms like their surface biopolymers.

Sph2 Growth Condition, and Preparation for Infrared Spectroscopy of Its Cell Surface Sph2 was grown in AB10 medium (a defined mineral medium with known exact chemical composition) (Tolker-Nielsen et al. 2000) for 96 h, please see the appendix for more information and details of the experimental conditions (Table 4). Immediately after the incubation, the bacterial cells were harvested by centrifugation and washed in 10 ml of sterile 0.9% NaCl solution. To analyze Sph2 cell surface by infrared spectroscopy, the harvested cells were re-suspended in 0.1 M NaCl and the suspension pH was adjusted to 4, 6.5, and 9.0. These three pH values were chosen to assay cell functional group changes in acidic, close to neutral (and the pH of the pervious experiments (Pouran et al. 2009, 2014, 2017; Geoghegan et al. 2008)), and alkaline environments. To adjust the pH values to acidic and alkaline, pH buffers were added to suspension of Sph2 cells in 0.1 M NaCl. For pH 6.5 instead of a pH buffer AB10 medium, which maintains the pH at this level, was added. After 6 h, Sph2 cells were re-harvested from the pH adjusted samples (pH values of the suspensions were also measured at the end of the exposures). The harvested Sph2 cells were mounted on the ATR-FTIR, attenuated total reflection-Fourier transform infrared, to record their spectra at room temperature. Prior to scanning the samples, as the harvested cells were still wet, the excess water was dried using pressurized nitrogen gas. The measurements were performed using an ATR-FTIR system comprising of a Germanium crystal (Ge) attached to a Perkin Elmer Spectrum One Fourier Transformation Infrared Spectrophotometer. For all samples, 200 infrared scans with the resolution of 4 cm1 wavenumber were collected.

Potentiometric Titration of Sph2 Potentiometric titration experiments were performed using an automated potentiometric titrator (Metrohm, 718 STAT, Titrino). The titration vessel consisted of a Pyrex cell of 100 ml capacity with a rubber lid equipped with holes for electrode, micro burette, thermometer, and N2 flushing plastic tube. Concentrations of 6.0 g/l of freeze-dried Sph2 were titrated in three different concentrations of background electrolytes (0.01, 0.1, and 0.3 M NaCl). Throughout the titrations, acid and base were added by computer-controlled microburette with a dispensing volume of 0.01 ml. The titrator was adjusted to add successive acid or base when the absolute value of the potential drift was equal or less than 5 mV/min. The cell suspensions were purged by N2 gas for approximately 2 h before titration and the titration was

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performed under N2 atmosphere. In these experiments, continuous stirring was provided by magnetic stirrer and the suspension temperature was kept at 25  C during the titration period. A similar titration procedure was followed for the blank solution. The collected titration data for the cell suspensions were plotted in terms of moles of deprotonated sites per mass of bacteria and was calculated by spread sheet (Fein et al. 2005). During this process, the surface charge of Sph2 in the aqueous media was calculated using the following equations (Eq. 1) (Ojeda et al. 2008; Fein et al. 2005; Stumm and Morgan 1996; Mustafa et al. 2004). Equation 1. Calculation of Sph2 consumed/released protons during adding acid and/or base in an aqueous medium δ0 ¼

ðCa  Cb þ ½OH   ½H þ Þ m

(1)

In this equation, δ0 is molar equivalents of surface charge (mmol/g), Ca, and Cb are concentrations (mol/l) of added acid and base to the cell suspension, respectively, m is mass of the freeze-dried bacterial cell in suspension (g/l), and [H+] and [OH] are molar concentrations of H+ and OH determined from measured solution pH. In this research, the potentiometric titration curves of the bacterial strain were plotted as a function of consumed proton concentrations versus suspension pH. PZC was obtained as the common intersection point of titration curves for three different concentrations of background electrolytes (0.01, 0.1, and 0.3 M NaCl).

Modelling Surface Protonation of Sph2 Based on the Titration Data To calculate protonation constants from the titration data, ProtoFit 2.1 (a computer program for analyzing potentiometric titration) was used to optimize the data. ProtoFit has been used before to optimize surface protonation models from acidbase titration data and to characterize the chemistry of bacterial cell walls (Ojeda et al. 2008; Turner and Fein 2006; ProtoFit n.d.; Yue et al. 2015; Zhao et al. 2015). To model the acidity constants, based on the experimental data, nonelectrostatic and constant capacitance models [NEM and CCM] were used. In a nonelectrostatic model (NEM), electrostatic interactions are assumed to be negligible, in other words no electrostatic effects affect adsorption, while the constant capacitance model (CCM) assumes that the charge distribution on the bacterial surface is homogenous. These two models have been used before to model surface protonation of different bacterial species (Ojeda et al. 2008; Turner and Fein 2006). These calculations consider a functional group acidic when it releases a proton to form a negatively charged surface species. A functional group is basic when it adsorbs a proton to form a positively charged surface species as seen in Eqs. 2 and 3 (Ojeda et al. 2008; Dittrich and Sibler 2005).

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Equation 2. Stoichiometric reaction for an acidic surface functional group (>S = Surface) > SOH $> SO þ H þ

(2)

Equation 3. Stoichiometric reaction for a basic surface functional group (>S = Surface) þ > SOH þ 2 $> SOH þ H

(3)

Zeta Potential Measurements Zeta potential is defined as the electrical potential difference at the shear plane or the interface between the bulk aqueous medium and static fluid layer attached to a bacterial cell (Martinez et al. 2008), please see the appendix for more information and details of the experimental conditions (Fig. 9). Sph2 zeta potential was obtained in 1 mM KCl solution, adjusted to different pH values. A zeta potential analyzer (Zeta Plus, Brookhaven Instruments, Huntsville, NY) was used for this purpose. Zeta Plus is operating based on light scattering (laser beam) to determine electrophoretic mobility of charged colloidal suspensions; in these studies, the mean value of 10 readings was reported. In the previous publications (Pouran et al. 2014, 2017; Geoghegan et al. 2008), Sph2 attachments on the mineral surfaces were studied when the cells were grown and incubated in AB10 growth medium, which has a high ionic strength, please see the appendix. To compare the zeta potential obtained in 1 mM KCl solution with an ionic strength compatible to the experimental conditions, Sph2 zeta potential in AB10 medium (ionic strength = 196.08 mM) was measured. As seen in the appendix, NaCl has the highest concentration in this defined growth medium. In fact, NaCl high concentration in AB10 contributes to more than 91% of this medium ionic strength (179 mM of 196.08 mM). Increasing ionic strength reduces the length of the diffusive double layer formed at the surface of a charged colloidal particle (including bacterial cells) and hampers the electrostatic interactions with other charged surfaces (Pouran et al. 2014; Chen and Walker 2007). To better understand how changes in the growth medium ionic strength affects Sph2 surface potential, the zeta potential was also measured in AB10 media with altered ionic strengths. As mentioned earlier, in AB10 more than 91% of the ionic strength is because of the sodium chloride concentration. To reduce the original ionic strength from 196.08 mM to new levels, NaCl initial concentration in AB10 was reduced to form two lower ionic strengths: 98.80 mM and 19.06 mM, and then Sph2 surface potential in these altered media were measured. For more details please see the appendix.

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Sph2 Biofilm Formation on Hematite As shown in the previous studies (Pouran et al. 2009, 2014, 2017; Andrews et al. 2010; Geoghegan et al. 2008), this hydrophilic bacterial strain with expected negative surface charge (in pH of the experiment, 6.5) shows negligible attachment on the positively charged hydrophilic iron oxides surfaces after 96 h incubation in AB10 medium. To better understand how electrostatic interactions affect Sph2 attached growth on the metal oxide mineral surfaces, hematite was chosen as the model metal oxide mineral. The hematite used in this experiment has PZC 7.5, which makes it positive in the experimental condition (pH 6.5) (Pouran et al. 2014, 2017). After 96 h incubation in AB10 growth medium with original and altered ionic strengths (as explained in the previous section), 98.80 mM and 19.06 mM, the biofilm formed by Sph2 on the hematite surface was evaluated through direct imaging, please see the appendix for more information and details of the experimental conditions. The experimental protocols including preparing the samples, staining the bacterial cells for the direct imaging, and the direct imaging procedure have been explained in details in the previous publications (Pouran et al. 2014, 2017).

Results and Discussions Infrared Spectroscopy of Sph2 Cell Wall Figure 1 shows the infrared spectra of Sph2 cells at three different pH values: 4, 6.5, and 9. As seen the observed spectra are compatible with the expected bacterial cell wall characteristics and consistent with the most common and universal functional groups of the bacterial surface: hydroxyl, carboxyl, phosphoryl, and amine groups (Claessens et al. 2004, 2006; Ojeda et al. 2008; Turner and Fein 2006; Dittrich and Sibler 2005; Dittrich and Luttge 2008; Fein 2006; Jiang et al. 2004; Vijayaraghavan and Yun 2008). Changing the pH value from acidic to basic does not lead to unexpected changes in Sph2 spectra. Deprotonation of the carboxylic group at higher pH values, present at

1720–1740 cm1, leads to lower intensity of this peak. These changes can be noticed when it is compared with the peak at 1542 cm1. The contribution of C═O of esters to this peak ( 1720–1740 cm1) prevents its complete disappearance due to deprotonation of the carboxylic group (Ojeda et al. 2008; Dittrich and Sibler 2005; Dittrich and Luttge 2008; Pouran et al. 2013). The suggested band assignments of Sph2 infrared spectra, based on the previous publications (Ojeda et al. 2008; Dittrich and Sibler 2005; Dittrich and Luttge 2008; Jiang et al. 2004; Pouran et al. 2013; Yee et al. 2004), studying bacterial species cell surface functional groups, are shown in Table 1.

Sph2 Surface Potential in Different Ionic Strengths Figure 2a demonstrates Sph2 surface potential values when suspended in 1 mM solution of KCl. This figure clearly shows that this bacterial strain has negative

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Fig. 1 The infrared spectra of Sph2 strain. After 96 h incubation in AB10 medium, bacterial cells were washed and re-suspended in 0.1 M NaCl with different adjusted pH values

surface potential at approximately neutral pH values, which is consistent with the expected, relatively low, point of zero charge of bacterial cells (Pouran et al. 2017; Geoghegan et al. 2008; Claessens et al. 2004, 2006; Leone et al. 2007). Figure 2b shows the dramatic impact of the solution ionic strength (IS) on Sph2 zeta potential. Increasing the solution ionic strength from 1 mM KCl to approximately 200 mM AB10 led to nearly 20 mV difference in the measured zeta potential of Sph2 cell suspensions in the above-mentioned media. As seen the cell surface charge was close to zero in AB10 medium (denoted IS1 in this figure). In Fig. 2a the most significant change in the surface potential can be seen, when the solution ionic strength was reduced from 98.08 mM (IS2) to 19.06 mM (IS3). This agrees with the previously published studies that suggest in the ionic strengths above 100 mM role of the

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Table 1 Absorption bands of the Sphingomonas spp., Sph2 functional groups Wavenumber (cm 1) ~1740–1725 ~1645 ~1540 ~1458 ~1402 ~1385 ~1315 ~1255 ~900–1200 ~1089 ~976

Functional group assignment Stretching C═O of ester functional groups primarily from membrane lipids and fatty acids, stretching C═O of carboxylic acids Stretching C═O of amides associated with proteins Bending of N─H and C─N stretching in amides, asymmetric stretching C═O of carboxylate Bending CH2/CH3 of proteins Bending CH2/CH3 of proteins and stretching C─O of carboxylic groups Stretching COO; bending CH2/CH3 Vibration of CH and CH2 Bending C─O of COOH, P═O stretching of phosphoryl groups PO2 and P(OH)2 stretching in phosphates, C─C, C─O─C and C─OH vibrations for polysaccharides P═O stretching of phophorylated proteins, phosphate storage products, or phophodiester Vibration of phosphoryl group

electrostatic interactions between the bacterial cells and their ambient environments is weakened because of the hampered diffusive double layer (Pouran et al. 2014; Chen and Walker 2007).

Sph2 Potentiometric Titration and the Data Optimization Using ProtoFit The potentiometric titration results show that Sph2 has PZC of approximately 4.3. As Fig. 3 shows, the titration curves at different ionic strengths exhibit a common intersection point approximately at pH 4.3. This value is compatible with the obtained zeta potential for Sph2 through the direct zeta potential measurements (Fig. 2a). In the data optimization process, the constant capacitance model (CCM), which is based on the concept of homogenous charge distribution on the bacterial surface (Ojeda et al. 2008; Turner and Fein 2006; Boily et al. 2001; Boily and Felmy 2008; Leone et al. 2007), was not able to provide a good fit to the potentiometric titration data compared to the non-electrostatic model (NEM). ProtoFit can be used to optimize surface protonation models with up to four discrete surface sites (deprotonation constants) (Turner and Fein 2006). For Sph2 potentiometric data optimization based on NEM, two options were chosen; NEM with three and four surface sites. Table 2, summarizes results of the data optimization by ProtoFit. As seen in this table, both three- and four-site models show good fits, but four-site model has a smaller weighted sum of squares value, SS*, which is an indication of better fit. As these calculations indicate the ProtoFit calculated PZC based on the four-site non-electrostatic model is compatible with the intersection point of the titration curves (Fig. 3). Figures 4 and 5 respectively demonstrate

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pH

a

0 3

4

5

6

7

8

9

Zeta Potential (mV)

–5

–10

–15

–20

–25

–30

Changes in surface potential as a result of changing ionic strength

b 0.00

Surface potential (mV)

IS1 (196.08 mM)

IS2 (98.08 mM)

IS3 (19.06 mM)

1 mM KCI

–5.00

–10.00

–15.00

–20.00

–25.00

Fig. 2 (a) Zeta potential of Sph2 strain suspended in 1 mM of KCl at different pH values. (b) Variations of the zeta potential of Sph2 strain in AB10 medium at different ionic strengths

calculated surface charge of Sph2 versus pH and compare fitting of the optimized data to the raw potentiometric titration values based on the four-site (four deprotonation constants) non-electrostatic model (NEM). As these figures suggest the estimated values are compatible with the raw titration data. Table 3 lists calculated acidity constants for Sphingomonas spp., Sph2. As seen it also shows examples of using NEM to calculate the acidity constants for other bacterial species (Fein et al. 2005; Ojeda et al. 2008; Yee et al. 2004). In this table, C indicates the concentration in 104 mole/g. In Gram-negative bacterial cells, the outermost layer consists of lipopolysaccharide. Although, the

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Fig. 3 Potentiometric titration data for 6 g/l (dry weight) of Sphingomonas spp., Sph2 in three different background electrolytes Table 2 Deprotonation constants and point of zero charge (PZC) as calculated by ProtoFit using NEM No of sites 3 4

pK1 3.9  0.2 4.1  0.3

pK2 5.9  0.4 5.9  0.2

pK3 9.4  0.2 8.9  0.0

pK4 – 10.2  0.1

pZC 4.37 4.55

SS* 4.18*102 2.90*102

SS*value is weighted sum of squares and reflects the goodness of fit

Gram-positive bacteria do not have this layer, they show similar surface charge pattern and PZC values, which may suggest that lipopolysaccharide is not a major contributor to the bacterial cell wall charge (Pouran et al. 2017; Claessens et al. 2006; Ojeda et al. 2008). Concentrations of the surface sites for Sph2 are within the range of reported surface functional groups’ concentrations. Total concentrations as low as 0.78  104 mol/g and as high as 16.6 104 mol/g have been reported in the previous studies (Ojeda et al. 2008; Dittrich and Sibler 2005). Although, the Sphingomonas species are Gram-negative but are unique among other bacterial species in this group. Instead of the lipopolysaccharide (LPS) layer, the Sphingomonaceae family have glycosphingolipids (GSL)

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Fig. 4 Calculated surface charge by Protofit modeling tool for Sph2. The four-site NEM model is applied

(Pouran et al. 2014, 2017). Glyscosphingolipids are a subgroup of glycolipids (lipids that are linked to a carbohydrate chain) and contain an sphingosine (amino alcohol) moiety (Gutman et al. 2014; Kawahara et al. 2001; Varki et al. 2008). For further information about glycosphingolipids, please see the appendix. Calculated Sph2 pKa values for the four-site model are 3.9  0.3, 5.9  0.2, 8.9  0.0, and 10.2  0.1, respectively. These pKa values could be tentatively assigned to carboxyl (2  pKa  6), phosphoryl (5.6  pKa  7.2), amine (8.6  pKa  9), and hydroxyl (8  pKa  12) groups (Ojeda et al. 2008; Dittrich and Sibler 2005; Jiang et al. 2004). For the four-site NEM: carboxyl, phosphoryl, and hydroxyl groups were assumed to be acidic functional groups, while amine considered to function as a basic group on the Sph2 cell surface. The definitions of acid and basic functional groups in this research have been mentioned in the previous sections (please see Eqs. 2 and 3). The proposed protonation/deprotonation reactions of Sph2 cell wall functional groups are shown in the following equations (Eqs. 4–7). Equation 4. Deprotonation of the carboxyl group (acidic functional group). >R- is the side chain þ > R  CO2 H $> R  CO 2 þH

(4)

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Fig. 5 Fitting of four-site nonelectrostatic model (NEM) to raw titration data. The raw titration graph corresponds to the second step of titration when NaOH is added to initially acid-added suspension

Equation 5. Deprotonation of the phosphoryl group (acidic functional group). >R- is the side chain > R  PO3 H 2 $> R  PO3 H  þ H þ

(5)

Equation 6. Deprotonation of the hydroxyl group (acidic functional group). >R- is the side chain > R  OH $> R  O þ H þ

(6)

Equation 7. Deprotonation of the amine group (basic functional group). >R- is the side chain þ > R  NH þ 3 $> R  NH 2 þ H

(7)

Species Sphingomnas spp., Sph2 gramnegative Bacilus subtilis gram-positive Calothorix sp. gram-negative Aquabacterium commune gramnegative

pK2 5.9  0.2 4.7 6.6  0.2 5.7  0.9

pK1 3.9  0.3

3.3 4.7  0.4 3.5  0.8

6.8 9.1  0.3 9.1  1.6

pK3 8.9  0.0 8.9 N/A N/A

pK4 10.2  0.1 2.77 14.6 14.2

C total 4.70 0.75 3.3  0.3 4.6  1.5

C1 1.79  0.3

0.96 4.1  0.3 1.9  0.6

C2 0.56  0.0

0.31 7.16  1.0 7.3  3.1

C3 1.19  0.3

0.75 N/A N/A

C4 1.16  0.2

Table 3 Comparison of deprotonation constants and surface site concentrations between Sphingomonas spp., Sph2 and some of other strains from different studies. NEM is used for all of these strains. Concentrations are shown in 104 mol/g

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Sph2 Biofilm Formation on the Hematite Surface Under Different Ionic Strengths Sph2 zeta potential measurements in the solutions with different ionic strengths (Fig. 2) underline the role of the growth medium’s ionic strength in the electrostatic behavior of this bacterial species. In other words, changing the solution ionic strength and consequently Sph2 surface potential should lead to a different attachment behavior. Figure 6 shows biofilm formation of Sph2 on the hematite surface after 96 h incubation in AB10 medium with the original and altered ionic strengths. This figure also displays the quantified number of the attached cells. As can be seen, when the ionic strength was reduced by a factor of ten, the number of the cells attached to the hematite increased approximately tenfold. The difference in Sph2 surface potential under these two conditions is approximately 20 mV, which reinforces the electrostatic attraction between positively charged hematite and a negatively charged Sph2 surface. This figure indicates that the electrostatic interactions between the hematite and Sph2 surface plays an important role in the number of attached cells. The electrostatic behavior of the bacterial surfaces is considerably more complex than minerals. The presence of different types of biopolymers at the cell surface, which can undergo conformational changes, can also influence the bacterial cell adhesion to the mineral surfaces (Andrews et al. 2010; Geoghegan et al. 2008). Changing the solution’s ionic strength induces such conformational changes to the bacterial surface polyelectrolytes. Bacterial polyelectrolytic polymeric layers that extend out the cell wall can function like anchors and aid the adhesion when the cell approaches a surface (Hong and Brown 2008; Emerson et al. 2006). These cell

Fig. 6 Attachment of Sph2 to the hematite surface under different ionic strengths (a: IS = 196.08 mM, b: IS = 98.08 mM, c: IS = 19.06 mM). Quantified number of cells attached to hematite after 96 h of incubation under different ionic strength (IS) conditions

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surface biopolymers show conformational changes due to changing the solution chemistry, specifically ionic strength and pH (Chen and Walker 2007; Redman et al. 2004). The presence of protons (low pH values) and/or high ionic strengths can change cell surface polymers from stretched to coiled forms as well as altering the rigidity of these macromolecules (Geoghegan et al. 2008; Abu-Lail and Camesano 2003). In an opposite way, increasing pH and decreasing the ionic strength leads to increasing ionization of the biopolymers and electrostatic interaction (repulsion) between the side chains. These non-screened electrostatic interactions cause lack of conformational stability of the surface polymers and random shape polymer structures (Fedorov et al. 2009), please see the appendix for more information and details of the experimental conditions. In the case of hematite-Sph2 interface, increasing the ionic strength shortens the distance, in which effective electrostatic attraction between these two surfaces occurs. In addition, due to the expected relatively rigid and compressed coiled conformation of Sph2 surface polymers in the high ionic strength conditions, these biopolymers become less capable of interacting with the hematite surface within a fixed distance from the surface. By reducing the ionic strength from approximately 200 mM to 20 mM, the cell surface polymers can be extended out and have random stretched shapes, which in addition to the stronger electrostatic attraction between the hematite and Sph2 likely aid Sph2 attachment to the hematite surface. Hydrogen bonding, although not as important as electrostatic interactions and biopolymers conformational changes, may also contribute to the adhesion behavior of Sph2 on the hematite surfaces. All sphingolipids have a hydrophobic backbone due to ceramide. Ceramide can be defined as the lipid moiety of glycosphingolipids; it consists of sphingosine (aminoalcohol with unsaturated hydrocarbon chain) and fatty acid (Lackie 2007). They are distinguished by the presence of a planar system including the amide to the hydroxyl groups disposed at the water/lipid interface. Sphingolipids act as both donors and acceptors to form hydrogen bonds, and that is because of the presence of amide, carboxyl, and hydroxyl groups (Prinetti et al. 2009a). In the pH of the experiment (pH 6.5), the hematite surface is positively charged and protonated. This suggests that though the hydrogen bonds still may form between Sph2 surface polymers and few deprotonated sites on the hematite, its contribution to the biofilm formation is considerably less important than the electrostatic interactions or conformational changes of the biopolymers in the lower ionic strengths.

Conclusions Surface properties of a Gram-negative bacterium Sphingomonas spp., Sph2 (isolated from groundwater at a phenol-contaminated site) and effects of the growth medium ionic strength on its adhesion to hematite were investigated. The ATR-FTIR analysis showed that despite the unique cell wall chemistry of Sph2 (as a member of Sphingomonaceae family and having glycosphingolipids instead of

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lipopolysaccharide) among the Gram-negative strains, its cell wall infrared spectra are compatible to other bacterial cells. Hydroxyl, carboxyl, phosphoryl, and amide groups were detected in Sph2 infrared spectra. The potentiometric titrations were performed between pH 3.5 to 10 in three different background electrolytes. The titration data indicated that Sph2 point of zero charge is 4.3. ProtoFit modeling showed that four-site non-electrostatic model (NEM) provides a good fit to the raw titration data. The pKa values for these four sites are 3.9  0.3, 5.9  0.2, 8.9  0.0, and 10.2  0.1, which could be assigned to carboxyl, phosphate, amine, and hydroxyl groups, respectively. Based on the four-site NEM, calculated PZC for Sph2 is 4.55, which agrees with the titration data. Measured zeta potential for Sph2 in 1 mM KCl was approximately 20 mV. However, this value considerably changed in AB10 medium and moved towards zero. Alterations of AB10 ionic strength (by changing NaCl concentrations in this defined mineral medium) from 20 mM to 100 mM shifted Sph2 surface potential from 20 mV to 5 mV. This shift was considerably less, 3 mV, when the ionic strength was increased from

100 mM to 200 mM. Changing the solution ionic strength and consequently surface potential led to dramatic increase in the number of the attached cells to the hematite surface. Electrostatic interactions, conformational changes of the cell surface biopolymers, and to considerably less extent hydrogen bonding can all contribute to the adhesion patterns of Sph2 to the metal oxides. Nevertheless, distinguishing the degree of contribution of each of them requires further investigation.

Cross-References ▶ Application of Novel Microbial Consortia for Environmental Site Remediation and Hazardous Waste Management Toward Low- and High-Density Polyethylene and Prioritizing the Cost-Effective, Eco-friendly, and Sustainable Biotechnological Intervention ▶ Bacterial Cell-mineral Interface, Its Impacts on Biofilm Formation and Bioremediation ▶ Bioremediation of Hormones from Waste Water ▶ Bioremediation of Mined Waste Land ▶ Biostimulation and Bioaugmentation: An alternative Strategy for Bioremediation of Ground Water Contaminated Mixed Landfill Leachate and Sea Water in Low Income ASEAN Countries ▶ Characterizing the Cell Surface Properties of Hydrocarbon-Degrading Bacterial Strains, a Case Study ▶ Development of an Environmentally Sustainable Approach for Safe Disposal of Arsenic-Rich Sludge

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▶ Investigation of the Chemical Content of Two Specific Streams in Municipal Waste: The Case of Hazardous Household Waste and Dental Waste ▶ Management of Radioactive Wastes ▶ Micro-remediation of Metals: A New Frontier in Bioremediation ▶ Soil Pollution and Remediation

Appendix Preparing the Hematite Surface for Cell Adhesion Studies (Synthesis, Coating, and Characterization) Hematite was prepared by heating an acidic solution of FeCl3. A STOE STADI P X-ray powder diffractometer and a Perkin Elmer Spectrum Spotlight FTIR imaging system for Fourier-transform infrared spectroscopy (FTIR) were used to analyze the synthetized materials. For XRD analysis, copper K alfa was the radiation source; a range of 10–70 degrees and a step size of 0.02 degrees were the test parameters. In FTIR experiments, the spectrum resolution was 4 cm1, covering the range of 4000–400 cm1 wave numbers, and 150 scans were collected for each sample. To determine the point of zero charge (PZC) of the synthetic metal oxide, potentiometric titration was done. An automated potentiometric titrator (Metrohm, 718 STAT, Titrino) was used. During titrations, acid (HCl, 0.1 M) and base (NaOH, 0.1 M) were added by a computer-controlled microburette with a dispensing volume of 0.01 ml. The titrator was adjusted to add successive acid or base when the absolute value of the potential drift was equal to or less than 5 mV/min. The sample suspensions were purged with N2 gas to remove carbon dioxide from the system for approximately 2 h before titration, which was performed in an N2 atmosphere. In these tests, a magnetic stirrer provided continuous stirring and the suspension temperature was kept at 25  C during the titration period. Surface hydrophobic/ hydrophilic properties of the synthetic minerals were obtained by measuring the water-drop contact angle in air. Contact angles were obtained using the sessile drop method and a KRÜSS DSA 100 drop-shape analysis system. An aliquote of 3 μl of UHQ water was added to the mineral surfaces at room temperature. The contact angle between the surface and a tangent drawn on the drop surface, passing through the triple point of atmosphere-liquid-solid, was measured. Iron oxides’ hydrophilic nature stems from their surface hydroxyl groups. In general, surfaces with a waterdrop contact angle of less than 90 degrees are hydrophilic; nevertheless, for the surfaces studied, the expected water-drop contact angles were considerably less. The coating process involved the direct deposition of mineral particles from an aqueous suspension by evaporation, which has been explained in detail in previous publications (Pouran et al. 2014, 2017). After this step, the coated polystyrene surfaces were assessed using optical microscopy (Zeiss, Axiovision), direct imaging, and contact-angle measurements to determine their hydrophobicity (as described above). The ATR-FTIR, attenuated total reflection-Fourier transform infrared, technique using a Specac Silver Gate Essential Single Reflection ATR System and XPS,

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and X-ray photoelectron spectroscopy (KRATOS-Axis 165) were also used to compare the chemical properties of altered surfaces with those of reference polystyrene and mineral surfaces (Pouran et al. 2014).

Bacterial Strains, Growth Conditions, and Sample Preparation Six bacterial strains were isolated for bacterial-adhesion and attached-growth studies. Rhodococcus spp., RC92 and RC291, both Gram-positive, were isolated from soil samples from a polluted gasworks site in northeast England. The bacteria Pseudomonas spp. (Pse1 and Pse2) and Sphingomonas spp. (Sph1 and Sph2) were isolated from groundwater at a phenol-contaminated site in the West Midlands (England). The strains Pse1, Pse2, Sph1, and Sph2 are Gram-negative. They have been classified using comparative 16S rRNA sequencing (Geoghegan et al. 2008). All strains were maintained on a solid R2A medium (Oxoid). The bacterial strains were grown in an AB10 medium, which is a defined medium with known exact chemical composition. The carbon source was 2 mM of glucose, and the incubation time was 96 h at 20  C on a shaker at 150 rpm. After incubation, cells were harvested by centrifugation in an early stationary phase and washed in 10 ml of sterile 0.9% NaCl solution. Samples of washed and resuspended strains (in 0.9% NaCl), with an optical density (OD) of 0.01 at λ = 600 nm, were resuspended in the AB10 medium. The aim of this study was to perform experiments, including bacterial cell growth and attachment, in a well-controlled environment.

Biofilm Formation Studies Six strains, four different surfaces, two carbon sources, and one experimental control (AB10 medium with no carbon source) were analyzed in triplicate to assay biofilm formation for a total of 216 samples. In these experiments, reference polystyrene plates were prepacked and radiation-sterilized. The mineral-coated polystyrene plates were sterilized by immersion in a 70% ethanol medium for 1 h prior to incubation and dried under aseptic conditions in a laminar flow cabinet. Noninvasive, in situ direct imaging using Syto9 stain (green fluorescent nucleic acid stain, supplied by Invitrogen) was used as the primary technique to assay biofilm. The reference polystyrene and metal-oxide coated polystyrene well-plates, each with 12 wells and a nominal culture area of 3.82 cm2 for each well, were used as substrata for biofilm formation studies. Samples of bacteria suspension were prepared at an optical density (OD) of 0.01 at λ = 600 nm using AB10 medium, pH 6.5. Then, 2 ml of prepared medium was added to each micro-well. The 12 well-plates were incubated for 96 h at 20  C (Fig. 7); then, 200 μl of each of the bacterial samples, from their planktonic phase, was transferred to a 96-micro well-plate and the OD was measured at λ = 630 nm to determine planktonic phase growth. To assess the planktonic phase of individual environmental isolates, the

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Fig. 7 Schematic representation of incubating polystyrene and mineral-coated 12-well plates and directly imaging the strains attached to the studied surfaces. (a) Depicts confined lateral movements of the water-dipping objective due to the well’s sides. As seen, a circle of diameter 11 mm located at the centre of each well’s base was imaged for the studied substrata. (b) Shows direct imaging of the aluminium hydroxide-coated plates. (c) Illustrates the function of Z-height focusing, Z stacking, used in evaluating biofilm formation. This method was used for dense biofilms to better assess the numbers of cells attached to polystyrene and mineral surfaces

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Fig. 8 Total number of bacterial cells attached to mineral-coated polystyrene and polystyrene surfaces after 96 h of incubation in AB10 medium with a glucose carbon source

measured optical density (OD) at λ = 630 nm was calibrated against the number of colony-forming units (CFU) for each strain. This calibration was used to compare growth in the planktonic phase for each individual strain. The rest of the planktonic phase was discarded and each well was gently washed three times by adding 5 ml of 0.9% sterile NaCl solution that was slowly added to the well wall and bottom intersection, using a pipette tip, to remove cells in the planktonic phase and ensure that only bacterial cells which had attached to the surface were present. Each well of the reference polystyrene and coated plates was stained by adding 0.5 ml of Syto 9, which was diluted 500. The thickness of the added stain layer that formed on the bottom of the well was approximately 1.25 mm (the surface area of each well was 3.82 cm2). The stained wells were directly imaged in situ using a 100 magnification Zeiss Achroplan water-dipping objective (Fig. 7). For imaging, a Zeiss AxioVision epifluorescence microscope with automated Z-height focusing (Z-stacking) was used for extended depth and field imaging. With this technique, a series of images are acquired at different focus positions, which allows imaging through a thick section or of a rough surface (Fig. 7). Images were captured with an Axiocam black & white camera using a 450–490 nm narrow-band pass filter. For each sample, 15 images were captured and then analyzed using AxioVision 4.6 and Image J software. From these digital

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Fig. 9 Schematic representation of distribution of electric charges and position of zeta potential around a negatively charged bacterial cell

images, direct cell counts were obtained and reported as cells/cm2 (since each experiment was conducted in triplicate, each data point represents an average of 45 data points). The microscope water-dipping objective had restricted lateral motion, due to the well’s sides, which confined the imaging area. Images to study bacterial cell attachment on the substrate, at the bottom of each well, were taken from a circular accessible surface with a diameter of 11 mm located at the center of the wells. As mentioned earlier, microscope Z-stacking provided the option of acquiring images at different focus positions. This technique was used to determine biofilm depth when the cells had formed dense biofilms (Figs. 8 and 9) (Table 4). Glycosphingolipids These chemical structures are amphiphilic molecules and generally possess similarities with physicochemical and functional properties of the LPS (Gutman et al. 2014; Kawahara et al. 2001). Glycosphingolipids have a versatile chemical

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Table 4 AB10 medium and its ionic strength

No 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16

Chemical (NH4)2SO4 Na2HPO4 KH2PO4 NaCl CaCl2 MgCl2 C6H12O6glucose FeCl3 CaSO4 FeSO4.7H2O MnSO4.H2O CuSO4 ZnSo4.7H2O CoSO4.7H2O NaMoO4. H2O H3BO3

Element 1 mM Concentration 3.02E + 00 6.74E + 00 2.20E + 00 1.79E + 02 1.00E-02 1.00E-01 2.00E-01 1.00E-03 1.47E-06 7.20E-07 1.18E-07 1.25E-07 6.96E-08 3.56E-08 4.98E-08

Valence 1 1 1 1 2 2 0

Element 2 mM Concentration 1.51E + 00 3.37E + 00 2.20E + 00 1.79E + 02 2.00E-02 2.00E-02 0.00E + 00

Valence 2 2 1 1 1 1 0

Ionic Strength 4.53E + 00 1.01E + 01 2.20E + 00 1.79E + 02 3.00E-02 2.10E-01 0.00E + 00

3 2 2 2 2 2 2 1

3.00E-03 1.47E-06 7.20E-07 1.18E-07 1.25E-07 6.96E-08 3.56E-08 4.98E-08

1 2 2 2 2 2 2 1

6.00E-03 5.88E-06 2.88E-06 4.73E-07 5.02E-07 2.78E-07 1.42E-07 4.98E-08

1

8.09E-08

8.09E-08 1 8.09E-08 Solution ionic strength in mM = 1.96E + 02

AB10 original medium (IS1 = 196.08 mM, [NaCl] = 179.00), AB10-IS2 (IS2 = 98.08 mM, [NaCl] = 81.00), AB10-IS3 (IS3 = 19.06 mM, [NaCl] = 2.52)

structure and can be found in the cell membranes of different organisms. Figure 10 shows the chemical structure of a glycosphingolipid; Gangiloside (GT1b) (Varki et al. 2008). Sph2 are unique among all Gram-negative bacteria as they have glycosphingolipid (GSL) instead of lipopolysaccharide (LPS) layer. Glycosphingolipid has a functional role in cell-cell recognition and signalling (Hakomori and Igarashi 1995); GSL molecular structure exhibits more than 200 variations in carbohydrate structure, which can be combined with at least ten common molecular species of ceramide (a composition of sphingosine and a fatty acid). This combination can create over 2000 possibilities of GSLs molecular species (Hakomori and Igarashi 1995). The presence of GSLs also can influence membrane proteins architecture as several classes of membrane associated proteins display a strong preference for the association of lipid-rich membrane domains (Prinetti et al. 2009). Some research shows that by changing KCl concentration from 0 to 100 mM, Pseudomonas putida KT2442 cell wall biopolymer length changes from 440 nm to 160 nm (pH = 8) (Abu-Lail and Camesano 2003). Biopolymers can undergo a saltinduced conformational change from a soft, random structure in low ionic strength to

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Fig. 10 Chemical structure of a glycosphingolipid molecule, Gangiloside (GT1b)

Fig. 11 Schematic representation of conformational changes of biopolymers due to changes in ionic strength. Low IS favors attachment, while high IS hinders it

a rigid structure in a highly ionic medium (Fig. 11). The effects of sodium chloride concentrations on Poly-L-glutamate (PGA) as a multifunctional biopolymer has been mentioned in the previous studies as a small amount of NaCl switches the preferred conformation of this polymer. Adding a concentration of 0.3 M NaCl to DI-water is sufficient to keep the dominant conformation of Poly-L-glutamate a complex α-helix rather than an extended structure, on a nanosecond time scale (Fedorov et al. 2009).

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References Abu-Lail NI, Camesano TA (2003) Role of ionic strength on the relationship of biopolymer conformation, DLVO contributions, and steric interactions to bioadhesion of pseudomonas putida KT2442. Biomacromolecules 4:1000–1012 Andrews JS, Pouran HM, Scholes J, Rolfe SA, Banwart SA (2009) Multi-factorial analysis of surface interactions in single species environmental bacteria and model surfaces. Geochim Cosmochim Acta 73:A44–A44 Andrews JS, Rolfe SA, Huang WE, Scholes JD, Banwart SA (2010) Biofilm formation in environmental bacteria is influenced by different macromolecules depending on genus and species. Environ Microbiol 12:2496–2507 Boily JF, Felmy AR (2008) On the protonation of oxo- and hydroxo-groups of the goethite (alphaFeOOH) surface: a FTIR spectroscopic investigation of surface O-H stretching vibrations. Geochim Cosmochim Acta 72:3338–3357 Boily JF, Lutzenkirchen J, Balmes O, Beattie J, Sjoberg S (2001) Modeling proton binding at the goethite (alpha-FeOOH)-water interface. Colloids Surf A Physicochem Eng Asp 179:11–27 Chen GX, Walker SL (2007) Role of solution chemistry and ion valence on the adhesion kinetics of groundwater and marine bacteria. Langmuir 23:7162–7169 Chen J, Wong MH, Wong YS, Tam NFY (2008) Multi-factors on biodegradation kinetics of polycyclic aromatic hydrocarbons (PAHs) by Sphingomonas sp. a bacterial strain isolated from mangrove sediment. Mar Pollut Bull 57(6–12):695–702 Claessens J, Behrends T, Van Cappellen P (2004) What do acid-base titrations of live bacteria tell us? A preliminary assessment. Aquat Sci 66:19–26 Claessens J, van Lith Y, Laverman AM, van Lith Y, Laverman AM, Van Cappellen P (2006) Acidbase activity of live bacteria: implications for quantifying cell wall charge. Geochim Cosmochim Acta 70:267–276 Dittrich M, Luttge A (2008) Microorganisms, mineral surfaces, and aquatic environments: learning from the past for future progress. Geobiology 6:201–213 Dittrich M, Sibler S (2005) Cell surface groups of two picocyanobacteria strains studied by zeta potential investigations, potentiometric titration, and infrared spectroscopy. J Colloid Interface Sci 286:487–495 Emerson RJ, Bergstrom TS, Liu YT, Soto ER, Brown CA, McGimpsey WG, Camesano TA (2006) Microscale correlation between surface chemistry, texture, and the adhesive strength of Staphylococcus epidermidis. Langmuir 22:11311–11321 Fedorov MV, Goodman JM, Schumm S (2009) The effect of sodium chloride on poly-L-glutamate conformation. Chem Commun, 896–898 Fein JB (2006) Thermodynamic modeling of metal adsorption onto bacterial cell walls: current challenges. Adv Agron:179–202 Fein JB, Boily J-F, Yee N, Gorman-Lewis D, Turner BF (2005) Potentiometric titrations of Bacillus subtilis cells to low pH and a comparison of modeling approaches. Geochim Cosmochim Acta 69(5):1123–1132 Gabriel FLP, Giger W, Guenther K, Kohler H-PE (2005a) Differential degradation of nonylphenol isomers by Sphingomonas xenophaga Bayram. Appl Environ Microbiol 71(3):1123–1129 Gabriel FLP, Heidlberger A, Rentsch D, Giger W, Guenther K, Kohler H-PE (2005b) A novel metabolic pathway for degradation of 4-nonylphenol environmental contaminants by Sphingomonas xenophaga Bayram: ipso-hydroxylation and intramolecular rearrangement. J Biol Chem 280(16):15526–15533 Geoghegan M, Andrews JS, Biggs CA, Eboigbodin KE, Elliott DR, Rolfe S, Scholes J, Ojeda JJ, Romero-Gonzalez ME, Edyvean RGJ et al (2008) The polymer physics and chemistry of microbial cell attachment and adhesion. Faraday Discuss 139:85–103 Gutman J, Kaufman Y, Kawahara K, Walker SL, Freger V, Herzberg M (2014) Interactions of glycosphingolipids and lipopolysaccharides with silica and polyamide surfaces: adsorption and viscoelastic properties. Biomacromolecules 15(6):2128–2137

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Hakomori S, Igarashi Y (1995) Functional-role of glycosphingolipids in cell recognition and signaling. J Biochem 118:1091–1103 Hong Y, Brown DG (2008) Electrostatic behavior of the charge-regulated bacterial cell surface. Langmuir 24:5003–5009 Jiang W, Saxena A, Song B, Ward BB, Beveridge TJ, Myneni SCB (2004) Elucidation of functional groups on gram-positive and gram-negative bacterial surfaces using infrared spectroscopy. Langmuir 20(14):11433–11442 Kawahara K, Lindner B, Isshiki Y, Jakob K, Knirel YA, Zähringer U, Zahringer U (2001) Structural analysis of a new glycosphingolipid from the lipopolysaccharide-lacking bacterium Sphingomonas adhaesiva. Carbohydr Res 333(1):87–93 Lackie JM (2007) The dictionary of cell and molecular biology, Academic Press, USA Leone L, Ferri D, Manfredi C, Persson P, Shchukarev A, Sjoberg S, Loring J (2007) Modeling the acid-base properties of bacterial surfaces: a combined spectroscopic and potentiometric study of the gram-positive bacterium Bacillus subtilis. Environ Sci Technol 41:6465–6471 Martinez RE, Pokrovsky OS, Schott J, Oelkers EH (2008) Surface charge and zeta-potential of metabolically active and dead cyanobacteria. J Colloid Interface Sci 323:317–325 Mustafa S, Tasleem S, Naeem A (2004) Surface charge properties of Fe2O3 in aqeous and alcoholic mixed solvents. J Colloid Interface Sci 275:523–529 O’Toole GA, Wong GC (2016) Sensational biofilms: surface sensing in bacteria. Curr Opin Microbiol 30:139–146 O’Toole G, Kaplan HB, Kolter R (2000) Biofilm formation as microbial development. Annu Rev Microbiol 54:49–79 Ojeda JJ, Romero-Gonzalez ME, Pouran HM, Banwart SA (2008) In situ monitoring of the biofilm formation of Pseudomonas putida on hematite using flow-cell ATR-FTIR spectroscopy to investigate the formation of inner-sphere bonds between the bacteria and the mineral. Mineral Mag 72(1):101–106. https://doi.org/10.1180/minmag.2008.072.1.101 Pouran HM, Fotovat A, Haghnia G, Halajnia A, Chamsaz M (2008) A case study: chromium concentration and its species in a calcareous soil affected by leather industries effluents. World Appl Sci J 5(4):484–489 Pouran HM, Andrews JS, Romero-Gonzalez M, Banwart SA (2009) Effects of surface charge and hydrophobicity of synthetic metal oxides on attached growth of environmental bacterial isolates. Geochim Cosmochim Acta 73:A1048–A1048 Pouran HM, Llabjani V, Martin FL, Zhang H (2013) Evaluation of ATR-FTIR spectroscopy with multivariate analysis to study the binding mechanisms of ZnO nanoparticles or Zn to Chelex100 or Metsorb. Environ Sci Technol 47(19):11115–11121 Pouran HM, Banwart SA, Romero-Gonzalez M (2014) Coating a polystyrene well-plate surface with synthetic hematite, goethite and aluminium hydroxide for cell mineral adhesion studies in a controlled environment. Appl Geochem 42(1986):60–68 Pouran HM, Banwart SA, Romero-Gonzalez M (2017) Effects of synthetic iron and aluminium oxide surface charge and hydrophobicity on the formation of bacterial biofilm. Environ Sci Process Impacts 19(4):622–634 Prinetti A, Loberto N, Chigorno V, Sonnino S (2009) Glycosphingolipid behaviour in complex membranes. Biochim Biophys Acta 1788(1):184–193 ProtoFit: A program for optimizing surface protonation models (home page). http://protofit. sourceforge.net/. Accessed 2 July 2017 Redman JA, Walker SL, Elimelech M (2004) Bacterial adhesion and transport in porous media: role of the secondary energy minimum. Environ Sci Technol 38:1777–1785 Stumm W, Morgan JJ (1996) Aquatic chemistry, chemical equilibria and rates in natural waters. Wiley, New York Tolker-Nielsen T, Brinch UC, Ragas PC, Andersen JB, Jacobsen CS, Molin S (2000) Development and dynamics of Pseudomonas sp. biofilms. J Bacteriol 182(22):6482–6489 Turner BF, Fein JB (2006) Protofit: a program for determining surface protonation constants from titration data. Comput Geosci 32(9):1344–1356

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Preparation, Characterization, and Heavy Metal Ion Adsorption Property of APTESModified Kaolin: Comparative Study with Original Clay

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Contents Definitions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Kaolin and Modified Kaolin . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Kaolin in the Environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Structural Features . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . The Main Important Characteristics of Kaolin . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Functionalization of Kaolin with 3-Aminopropyltriethoxysilane (APTES) . . . . . . . . . . . . . . . Heavy Metals Removal by Adsorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cobalt . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Nickel . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Copper . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Zinc . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

Surface modification of clay minerals has become increasingly important for improving the practical applications of clays such as fillers and adsorbents. In this study, we developed an effective adsorbent for retention of Co(II), Ni(II), Cu(II), and Zn (II) by modifying kaolin clay with an amino-terminated organosilicon (3-aminopropyltriethoxysilane, APTES). The lamellar filler (original clay) used is the kaolin (K08) from Bechar–Algeria region. Characterization of modified clay material and original clay was carried out by different methods XRD, FTIR, TGA, and SEM to establish the link between syntheses, structures, and properties. The adsorption of heavy metal ions onto APTES-modified kaolin showed greater efficiency than original kaolin. These results indicate that the

B. Meroufel · M. A. Zenasni (*) Faculty of Technology, University Abou Bekr Belkaïd of Tlemcen, Tlemcen, Algeria © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_132

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APTES-modified kaolin may be used as very effective adsorbent for removal of heavy metals from aqueous media. Keywords

Clay · Kaolin · Modified kaolin · Heavy metal · Adsorption Abbreviations

APTES K08 KS

3-Aminopropyltriethoxysilane Kaolin of Bechar Modified kaolin by 3-aminopropyltriethoxysilane

Definitions Clay

Kaolin Heavy Metal

Although broadly distributed and well known, clays are difficult to define precisely. The term is applied to finely grained natural materials which are plastic when wet and hard and brittle when dried, especially after firing. They are usually complex mixtures of various minerals, with the main component being a platy aluminosilicate. A clay where the principal clay mineral is kaolinite, but significant amounts of other minerals can be present. Cobalt, nickel, copper, and zinc are some of the metals (called “heavy” because of their high relative atomic mass) which persist in nature and can cause damage or death in animals, humans, and plants even at very low concentrations (1 or 2 micrograms in some cases). Used in industrial processes, they are carried by air and water when discharged in the environment. Since heavy metals have a propensity to accumulate in selective body organs (such as the brain and liver), their prescribed average safety levels in food or water are often misleadingly high.

Introduction Over the years, the percolation of heavy metals into the water bodies and ecosystem remains as one of the most elusive and pervasive environmental threat to the global occupants. Heavy metal ions are classified as priority pollutants based on their toxicity and mobility in natural water streams. Nevertheless, the heavy metal ions are stable and persistent to environment changes since they cannot either be degraded or destroyed (Demirbas 2008). The increment of industrialization has

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aggravated the situation due to the mass loading of highly concentrated metal ion containment effluent into the waterways. The feasibility and reliability of lignocellulosic biomass, natural clay minerals, and biological-based materials used as low-cost adsorbent have been evaluated by many researchers. These materials include sugarcane bagasse, risk husk, tea leaves, bamboo dust, maize cob, tree sawdust (Mohamad Ibrahim et al. 2013), zeolite (Mohamad Ibrahim et al. 2010; Castaldi et al. 2008), bentonite (Karapinar and Donat 2009), montmorillonite (Ijagbemi et al. 2009), kaolin (Meroufel et al. 2013), Cephalosporium aphidicola (Tunali et al. 2006), Pinus sylvestris Ucun et al. 2003), Saccharomyces cerevisiae (Huang et al. 1990), and so forth. Kaolin has received considerable recognition as an adsorbent because of its high adsorption capacity. It is generally referred to as clay that is mainly composed of kaolinite and a lower amount of minerals such as quartz and mica. It is the most abundant mineral in soils and sediments and interacts with other soil elements to contribute to the mechanical stability of the soil column (Chen et al. 2000). The adsorption properties of kaolinite are likely determined by its surface structure and the edges (Schoonheydt and Johnston 2006). The edges possess a variable charge that can be correlated to the reaction between ionizable surface groups along the edges and the clay mineral surface and the ions present in aqueous solution. The modification process for this clay is based on the strategy commonly used for kaolin functionalization. It consists of a direct condensation reaction between 3-aminopropyltriethoxysilane (APTES) and the hydroxyl groups of the clay minerals. In this work, we modified kaolin with organosilicon (3-aminopropyltriethoxysilane, APTES) to investigate kinetics and the mechanism of adsorption of Co(II), Ni(II), Cu(II), and Zn (II) onto modified kaolin (KS). Langmuir and Freundlich equations were used to determine the isotherm which gives the best correlation with experimental data. Pseudo-second-order model was used to evaluate the adsorption kinetics for both adsorbents.

Kaolin and Modified Kaolin Kaolin in the Environment Clays are hydrous aluminosilicates broadly defined as those minerals that make up the colloid fraction (< 2 μm) of soils, sediments, rocks, and water (Pinnavaia 1983) and may be composed of mixtures of fine-grained clay minerals and clay-sized crystals of other minerals such as quartz, carbonate, and metal oxides. Usually the term clay is used for materials that become plastic when mixed with a small amount of water. Clays play an important role in the environment by acting as a natural scavenger of pollutants by taking up cations and anions either through ion exchange

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or adsorption or both. Thus, clays invariably contain exchangeable cations and anions held to the surface. The prominent cations and anions found on clay surface are Ca2+, Mg2+, H+, K+, NH4+, and Na+ and SO42, Cl, PO43, and NO3. These ions can be exchanged with other ions relatively easily without affecting the clay mineral structure. Large specific surface area, chemical and mechanical stability, layered structure, high cation exchange capacity (CEC), etc. have made the clays excellent adsorbent materials. Both Brönsted and Lewis types of acidity in clays (Tanabe 1981) have boosted the adsorption capacity of clay minerals to a great extent. The Brönsted acidity arises from H+ ions on the surface, formed by dissociation of water molecules of hydrated exchangeable metal cations on the surface:  nþ M ðH 2 O Þx

!

 ðn1Þþ MðOHÞðH2 OÞx1 þ Hþ

(1)

The Brönsted acidity may also arise if there is a net negative charge on the surface due to the substitution of Si4+ by Al3+ in some of the tetrahedral positions and the resultant charge is balanced by H3O+ cations. The Lewis acidity arises from exposed trivalent cations, mostly Al3+ at the edges, or Al3+ arising from rupture of Si–O–Al bonds, or through dehydroxylation of some Brönsted acid sites. The edges and the faces of clay particles can adsorb anions, cations, and nonionic and polar contaminants from natural water. The contaminants accumulate on clay surface leading to their immobilization through the processes of ion exchange, coordination, or ion–dipole interactions. Sometimes the pollutants can be held through H-bonding, van der Waals interactions, or hydrophobic bonding arising from either strong or weak interactions. The strength of the interactions is determined by various structural and other features of the clay mineral. van Olphen (van Olphen 1977) has cited several types of active sites in clays, viz.: (i) Brönsted acid or proton donor sites, created by interactions of adsorbed or interlayer water molecules (ii) Lewis acid or electron acceptor sites occurring due to dehydroxylation (iii) Oxidizing sites, due to the presence of some cations (e.g., Fe3+) in octahedral positions or due to adsorbed oxygen on surfaces (iv) Reducing sites produced due to the presence of some cations (e.g., Fe2+) (v) Surface hydroxyl groups, mostly found in the edges, bound to Si, Al, or other octahedral cations The present study determined the mineralogy and thermal properties of kaolin from Tabelbala (Algeria). For many centuries kaolinic clays have been exploited in the Tabelbala (Algeria) as a cement or waterproof cover on the roofs. They are associated with lower Paleozoic sandstones. In the context of the valorization of Algeria’s natural resources, a clay collected at Tabelbala in the Bechar region (southwestern Algeria) and its clay fraction (particles with a diameter of less than 2 μm) have undergone a series of mineralogical, chemical, and physicochemical analyses.

47

Preparation, Characterization, and Heavy Metal Ion Adsorption. . .

1171

Tabelbala is a department of the province of Bechar, located about 145 km southwest of Béni Abbès and 400 km south of Bechar (Fig. 1). Tabelbala consists of several small villages, the largest of which are Sidi Zekri to the east and Ksar Cheria and Makhlouf to the west. There is a deposit of kaolin in Makhlouf, a prodigious wealth (Fig. 2).

Structural Features The structural features of the clays are well established. Kaolinite has a 1:1 layer structure, first suggested by Pauling (Pauling 1930), with the basic unit consisting of

Fig. 1 Location of Tabelbala in far southwest of Bechar (Algeria). Location of Makhlouf quarry

Fig. 2 Deposit of kaolin in Makhlouf–Tabelbala–Bechar

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Fig. 3 Structure of kaolinite

OH

OH

OH OH

=O

= Shared O

OH OH

= Si

= Al

a tetrahedral sheet of SiO4 and an octahedral sheet with Al+3 as the octahedral cation. Both the sheets combine to form a common layer such that the tips of the silica tetrahedra point toward the octahedral layer. The tetrahedral layer is inverted over the octahedral layer with the apical “O” atoms being shared by the two layers (Fig. 3). In the layer common to the octahedral and tetrahedral sheets, two-thirds of the O-atoms are shared between Si and Al atoms. The remaining one-third of the sites in this layer consists of hydroxyl groups coordinated to the octahedral Al atoms alone. Two-thirds of the possible positions in the octahedral sheet are filled with Al, and the remaining one-third of the sites are vacant. The Al atoms are placed in such a manner that any two Al atoms are separated by two hydroxyl groups, one above and one below, making a hexagonal distribution in a single plane in the center of the octahedral sheet. The hydroxyl groups are placed directly against the centers of oxygen hexagons of the basal plane of the tetrahedral layer (Theng 1979). Kaolinite, (Si4)IV(Al4)VIO10(OH)8, has the theoretical composition of SiO2 46.54%, Al2O3 39.50%, and H2O 13.96% expressed in terms of the oxides. The formula indicates that there is no substitution of Si4+ with Al3+ in the tetrahedral layer and no substitution of Al3+ with other ions (e.g., Mg2+, Zn2+, Fe2+, Ca2+, Na+, or K+) in the octahedral layer. Thus, the net layer charge of kaolinite is [4 (+4)] + [4 (+3)] + [10 (2)] + [8 (1)] = 0, but in nature, kaolinite has a small net negative change arising from broken edges on the clay crystals. This negative charge, although small, is responsible for the surface not being completely inert. Some workers have also reported substitution of octahedral Al3+ with Fe2+ and/or Ti4+ in kaolinite (Deer et al. 1985).

The Main Important Characteristics of Kaolin The granulometric analysis indicates that the particles with a diameter of Cu > Ni > Co. Kaolin used in this work represents a cheap and viable option for the removal of low-concentration metal ions from polluted water. Acknowledgments The authors gratefully acknowledge Dr. Yves Pillet (Faculty of Sciences and Technology, group PGCM, University of Lorraine, Nancy, France) because of his contribution to our study and are thankful for Joint Service Electronic Microscopy and Microanalysis at the University Henri Poincare of Nancy for MEB-EDS analysis.

References Babel S, Kurniawan TA (2003) Low-cost adsorbents for heavy metals uptake from contaminated water: a review. J Hazard Mater B97:219 Castaldi P, Santona L, Enzo S, Melis P (2008) Sorption processes and XRD analysis of a natural zeolite exchanged with Pb(2+), Cd(2+) and Zn(2+) cations. J Hazard Mater 156(1–3):428–434 Chen J, Anandarajah A, Inyang H (2000) Pore fluid properties and compressibility of kaolinite. J Geotech Geoenviron Eng 126:798 Cristóbal AGS, Castelló R, Luengo MAM, Vizcayno C (2010) Zeolites prepared from calcined and mechanically modified Kaolins: a comparative study. Appl Clay Sci 49(3):239–246 Deer WA, Howie RA, Zussman J (1985) An introduction to the rock-forming minerals. ELBS Longman, Essex, pp 260–263 Demirbas A (2008) Heavy metal adsorption onto agro-based waste materials: a review. J Hazard Mater 157(2–3):220–229 Ekosse G (2000) The Makoro kaolin deposit, southeastern Botswana: its genesis and possible industrial applications. Appl Clay Sci 16:301–320 Freundlich HMF (1906) Over the adsorption in solution. J Phys Chem 57:385–470 Galan E, Aparicio P, Miras A, Michailidis K, Tsirambides A (1996) Technical properties of compounded kaolin sample from Griva (Macedonia, Greece). Appl Clay Sci 10(6):477–490 Huang CP, Huang CP, Morehart AL (1990) The removal of Cu (II) from dilute aqueous solutions by Saccharomyces cerevisiae. Water Res 24(4):433–439

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Ijagbemi CO, Baek MH, Kim DS (2009) Montmorillonite surface properties and sorption characteristics for heavy metal removal from aqueous solutions. J Hazard Mater 166(1):538–546 Jenne E (2007) Adsorption of metals by geomedia variables, mechanisms, and model applications. Elsevier, Washington, DC Karapinar N, Donat R (2009) Adsorption behaviour of Cu2+ and Cd2+ onto natural bentonite. Desalination 249(1):123–129 Langmuir I (1918) The adsorption of gases on plane surfaces of glass, mica and platinum. J Am Chem Soc 40:1361–1403 Le Pluart L (2002) Nanocomposites epoxy/amine/montmorillonite: role of interactions on formation, morphology at different scale levels and mechanical properties of networks. Doctoral thesis of the National Institute of Applied Sciences of Lyon Martinez-Ramirez (2007) Alkali activation of metakaolins: parameters affecting mechanical, structural and microstructural properties. J Mater Sci 42(9):2934–2943 Meroufel B, Benali O, Benyahia M, Benmoussa Y, Zenasni MA (2013) Adsorptive removal of anionic dye from aqueous solutions by Algerian kaolin: characteristics, isotherm, kinetic and thermodynamic studies. J Mater Environ Sci 4(3):482–491 Mohamad Ibrahim MN,Wan Ngah WS, Norliyana MS,Wan Daud WR, Rafatullah M, Sulaiman O, Hashim R (2010) A novel agricultural waste adsorbent for the removal of lead (II) ions from aqueous solutions. Hazard. J Mater 182(1–3):377–385 Mohanty K, Das D, Biswas MN (2006) Preparation and characterization of activated carbons from Sterculiaalata nutshell by chemical activation with zinc chloride to remove phenol from wastewater. Adsorption 12:119 Mureseanu M, Cioatera N, Trandafir I, Georgescu I, Fajula F, Galarneau A (2011) Selective Cu2+ adsorption and recovery from contaminated water using mesoporous hybrid silica bio-adsorbents. Microporous Mesoporous Mater 146:141–150 Pauling L (1930) The structure of chlorites. Proc Natl Acad Sci U S A 16:578 Pinnavaia TJ (1983) Intercalated clay catalysts. Science 220:365 Schoonheydt RA, Johnston CT (2006) Surface and interface chemistry of clay minerals. Develop. Clay Sci 1:87 Tanabe K (1981) Solid acid and base catalysis. In: Anderson JR, Boudart M (eds) Catalysis-science and technology. Springer, New York, p 231 Theng BKG (1979) Formation and properties of clay polymer complexes, vol 551. Elsevier, New York, pp 1–12 Tran HH, Roddick FA, O’Donnell JA (1999) Comparison of chromatography and desiccant silica gels for the adsorption of metal ions  I. Adsorption and kinetics. Water Res 33:2992 Tunali S, Akar T, Özcan AS, Kiran I, Özcan Sep A (2006) Equilibrium and kinetics of biosorption of lead(II) from aqueous solutions by Cephalosporium aphidicola. Purif Technol 47(3):105–112 Ucun H, Bayhan YK, Kaya Y, Cakici A, Algur OF (2003) Biosorption of lead(II) from aqueous solution by cone biomass of Pinus sylvestris. Desalination 154(3):233–238 Van Olphen H (1977) An introduction to clay colloid chemistry. Wiley Interscience, New York, p 187 Xue A, Zhou S, Zhao Y, Lu X, Han P (2011) Effective NH2-grafting on attapulgite surfaces for adsorption of reactive dyes. J Hazard Mater 194:7–14 Zafar MN, Nadeem R, Hanif MA (2006) Biosorption of nickel from protonated Rice bran. J Hazard Mater 143:478–485

Concept Note on Method Development for Speciation and Measurement of Arsenic (As) in Its Valence States (As (III) and As (V)) in Solid and Semisolid Organic Environmental Samples

48

A. M. M. Maruf Hossain

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Experimental Approaches . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sampling . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sample Pretreatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Speciation of As (III) and As (V) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Work Plan . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sampling . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Treatments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

1192 1193 1193 1193 1193 1195 1195 1196 1197

Abstract

Arsenic, a metalloid widely distributed in nature, is released into the environment from a wide range of sources. Its toxicity and carcinogenicity mainly result from its trivalent form rather than its pentavalent state. This makes the speciation and measurement of arsenic in its valence states extremely important for assessing potential health hazard. Arsenic speciation methods exist for natural water samples. However, for solid samples it is often the total arsenic content that is measured. This chapter lays out a concept note on method development for speciation and measurement of arsenic in its valence states (i.e., As (III) and As (V)) in solid and semisolid organic environmental samples. The concept is

A. M. M. M. Hossain (*) School of Global, Urban and Social Studies, College of Design and Social Context, RMIT University, Melbourne, VIC, Australia Center for Integrated Knowledge Invention, Laverton, VIC, Australia e-mail: [email protected] © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_159

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based on chemical extraction of As (III) and As (V) species through respective precipitation followed by their dissolution into solution phase, which could be measured through “high-performance liquid chromatography inductively coupled plasma mass spectrometry” (HPLC/ICP-MS). Experimental approaches and work plan on this concept have also been elaborated. Keywords

Arsenic · Arsenic (III) · Arsenic (V) · Speciation method · Environmental samples

Introduction Arsenic (As) is widely distributed in nature (in Earth’s crust at 2–5 mg/Kg (Onishi and Sandell 1955)), and is classified as metalloid, i.e., occurring both in solid and liquid states (Carson et al. 1986). It is released into the environment from a wide range of sources including industrial processes (such as use in hardening lead, in glass manufacturing, in electrical devices), during power generation from coalfired furnace, from use in agricultural and silvicultural products, as well as as feed additive for livestock. As the toxicity and carcinogenicity of arsenic mainly result from exposure to its trivalent form, i.e., As (III), rather than its pentavalent state, i.e., As (V) (Costa 2000), there is a necessity of measuring arsenic species in a given sample instead of simply determining total arsenic content in that sample. Liquid chromatographic methods exist for measuring arsenic species in natural waters, such as the HPLC/ICP-MS technique (ISO/IEC 2005; Komorowicz and Baralkiewicz 2014). However, solid and semisolid samples cannot be measured using similar method. For solid samples it is often the total arsenic content that is measured (primarily through acid digestion, such as the method described by FDA (2015)) in assessing the corresponding arsenic hazard. As As (III) is mostly responsible for arsenic toxicity, a concept note is developed in this article towards method development for speciation and measurement of arsenic in its valence states (i.e., As (III) and As (V)) in solid and semisolid organic environmental samples. This concept note is intended for organic environmental samples as it is not possible to measure arsenic contents in soils and sediments without destroying their lattice structure that involves digestion method, rendering the sample unsuitable for speciation and measurement of ions in different valence states. This is due to the alterations that occur in relative abundance of ions through oxidation reaction, resulting in As (III) turned into As (V). This concept note is based on avoiding acid digestion in order not to disturb the natural relative abundance of As (III) and As (V) in the samples, and instead through inducing definitive chemical reactions that can extract and isolate As (III) and As (V) for their determination.

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Concept Note on Method Development for Speciation and Measurement of. . .

1193

Experimental Approaches Sampling The in situ-collected samples ought to be kept in airtight condition in order to prevent sampling substance modification through oxidation, and be quickly transported to laboratory for analysis. The sample treatment and analysis are to be carried out in original condition (i.e., raw weight basis) while the dry weight coefficient is to be determined and adjusted from a separate portion of the sample.

Sample Pretreatment The solid samples are to be ground into small pieces in order to expose the maximum possible surface areas of the sample matrix, and be analyzed as soon as the grinding is complete in order to prevent alteration of the redox condition of sample matrix through oxidation from air.

Speciation of As (III) and As (V) Step 1 Arsenic present both as As (III) and As (V) compounds or As3+ and As5+ ions in organic solid or semisolid samples in absorbed, adsorbed, or free state form that can be precipitated as arsenic (III) sulfide and arsenic (V) sulfide through use of hydrosulfuric acid in cold acid (hydrochloric) solution. As in neutral or basic condition the hydrosulfuric acid would undergo no reaction with As (III) and As (V) compounds, hence no precipitate of arsenic (III) sulfide and arsenic (V) sulfide; it is necessary to maintain acidic condition in the medium for the reaction to occur. Although the color of both of the precipitates is yellow, due to similar potential reaction occurring with other compounds present in the sample matrix can yield precipitates of various colors, hence interfering with the color of arsenic (III) sulfide and arsenic (V) sulfide precipitates. As ðIIIÞ compound þ3H2 S þ 6HCl! ðSuch as 2Na3 AsO3 Þ

2½AsCl4  þ 3H2 S !

acid medium

2AsO2  þ 5H2 S !

acid medium

As2 S3 # þ8Cl ðYellow precipitateÞ

acid medium

ðSuch as 2H3 AsO4 Þ



As2 S3 # þ2HS ðYellow precipitateÞ

2HAsO2 þ 3H2 S ! As ðVÞ compound þ5H2 S

As2 S3 # þ6NaCl ðYellow precipitateÞ

þ 6H2 O

þ 6Hþ

þ 4H2 O

As2 S3 # þ4H2 O ðYellow precipitateÞ

cold, HCl medium

!

As2 S5 # þ8H2 O ðYellow precipitateÞ

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Step 2 The precipitate is to be filtered together with the sample residue, and dried with filter paper. If the mass is to be washed with water in order to remove any water-soluble impurities, a correction factor must be introduced to account for the solubility of arsenic (III) sulfide and arsenic (V) sulfide precipitates. For example, the solubility of arsenic (III) sulfide in water is 5  105 g/100 mL (i.e., 0.5 mg/L). The filtered solution phase is to be kept for checking any presence of arsenic through applying acid digestion method (FDA 2015). This accounts for arsenic (if any) released from the sample into the solution phase instead of forming precipitate. Step 3 Both of the arsenic (III) sulfide and arsenic (V) sulfide precipitates are readily soluble in solutions of alkali hydroxides, ammonium hydroxide, and ammonium sulfide. LiOH∕ NaOH∕ KOH in aqueous medium 2As2 S3 þ 4OH ! 

2As2 S3 þ 4NH4 OH!

AsO2  þ 3AsS2  þ 2H2 O ðDissolved in colorless solutionÞ

AsO2  þ 3AsS2  þ 2H2 O þ 4NH4 þ ðDissolved in colorless solutionÞ

As2 S3 þ 3S2 !2AsS3 3 LiOH in aqueous medium 4As2 S5 þ 24OH  ! 4As2 S5 þ 24NH4 OH!

3AsO4 3 þ 5AsS4 3 þ 12H2 O ðDissolved in colorless solutionÞ

3AsO4 3 þ 5AsS4 3 þ 12H2 O þ 24NH4 þ ðDissolved in colorless solutionÞ

As2 S5 þ 3S2 2 !2AsS4 3 þ 3S

In this step, the filtered mass containing arsenic (III) sulfide and arsenic (V) sulfide precipitates is to be mixed with ammonium hydroxide or lithium hydroxide (in aqueous medium), which would dissolve both of the precipitates into liquid phase. This should separate the As (III) and As (V) species from solid phase (i.e., filtered mass) and integrate them into the solution phase. Although the dissolution of arsenic (III) sulfide and arsenic (V) sulfide should produce a colorless solution, potential mixture of some other colors could also occur due to dissolution of some other precipitates present in the filtered mass. The solution phase is to be filtered and collected for subsequent step, while the remaining solid residue is to be kept for checking any yet uncollected arsenic content through applying acid digestion method (FDA 2015).

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Concept Note on Method Development for Speciation and Measurement of. . .

1195

Step 4 The solution collected from step # 3 would contain As (III) as AsO2 and AsS2, and As (V) as AsO43 and AsS43. These As (III) and As (V) species could be measured through high-performance liquid chromatography inductively coupled plasma mass spectrometry (HPLC/ICP-MS) as described by Komorowicz and Baralkiewicz (2014).

Work Plan The work plan on this concept needs to individually study all representative chemical reactions at sufficient details so that the efficiency and best working range for each step could be ascertained. The impacts of various factors for common sample types also need to be studied for the optimization of working procedures.

Sampling There should be two types of samples: theoretical and environmental samples.

Theoretical Samples The theoretical samples include four pure-grade arsenic compounds as described in Table 1. Table 2 studies the solubility of the theoretical arsenic sample compounds in acids, bases, and aqueous medium. Environmental Samples • Primary samples: Heavily arsenic-contaminated organic environmental samples • Secondary samples: Plant- and animal-sourced organic samples such as vegetables, fruits, crops, meat, dairy, and egg • Tertiary samples: Other organic solid and semisolid environmental samples There is no pretreatment required for the theoretical samples, while the pretreatment for the environmental samples should be the same as described in section “Sample Pretreatment.” Table 1 List of theoretical samples

Name Arsenic oxide Arsenic sulfide

[Oxidation state + 3] As3+, AsO2, AsO33 ions (colorless) As2O3 (white color) As2S3 (yellow color)

[Oxidation state + 5] As5+, AsO43, As2O74 (colorless) As2O5 (white color) As2S5 (yellow color)

1196 Table 2 Solubility observation table for theoretical arsenic sample compounds

A. M. M. M. Hossain

Oxidation state +3 +5 +3 +5

Compound As2O3 As2O5 As2S3 As2S5

Solubility in H2O (g/100 mL)

Solubility in acids

Solubility in bases

Treatments Theoretical Arsenic Speciation The following treatments are to be conducted for each of the theoretical arsenic sample compounds: 1. As in sample: H2S: HCl ratio 1 mole:3 moles:6 moles 1 mole:3 moles:8 moles 1 mole:3 moles:10 moles 1 mole:5 moles:6 moles 1 mole:5 moles:8 moles 1 mole:5 moles:10 moles 1 mole:7 moles:6 moles 1 mole:7 moles:8 moles 1 mole:7 moles:10 moles Result As (III) and As (V) species must be in the precipitate, which can be confirmed through analyzing the filtered solution phase for total arsenic. The highest precipitate weights should be forwarded to the subsequent step. 2. [As in precipitate: NH4OH] 1 mole:2 moles (conc.) 1 mole:4 moles (conc.) 1 mole:6 moles (conc.) 1 mole:8 moles (conc.)

and

[As in precipitate: LiOH]

1 mole:2 moles (aqueous) 1 mole:4 moles (aqueous) 1 mole:6 moles (aqueous) 1 mole:8 moles (aqueous)

Result As (III) and As (V) species must be in the solution phase, which can be confirmed from observing no precipitate, or analyzing the precipitate (if any) for total arsenic. The treatments resulting in complete solution or the solutions with the least precipitates—after confirming any arsenic presence in the precipitate—should be taken for As (III) and As (V) species measurement through HPLC/ICP-MS (Komorowicz and Baralkiewicz 2014).

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Concept Note on Method Development for Speciation and Measurement of. . .

1197

Empirical Total Arsenic The theoretical arsenic sample compounds are to be acid digested (FDA 2015) to measure empirical total arsenic, which is to be compared with the sum of the determined As (III) and As (V) species for verifying the efficiency of the speciation.

Arsenic Speciation in Environmental Samples The optimum ranges of reaction conditions observed for theoretical arsenic speciation are to be utilized to study the speciation efficiency for all three types of organic environmental samples. This includes measuring empirical total arsenic for each sample type in order to determine the extent of arsenic retrieved through the speciation method, and following through any further improvements necessary for improving the efficiency of the method including optimization for each sample type.

References Carson BL, Ellis HV, McCann JL (1986) Toxicology and biological monitoring of metals in humans: including feasibility and need. Lewis Publishers, Chalsea Costa M (2000) Trace elements: aluminum, arsenic, cadmium, and nickel. In: Lippmann M (ed) Environmental toxicants: human exposures and their health effects, 2nd edn. Wiley, Hoboken, pp 811–850 FDA (2015) Elemental analysis manual for food and related products, Section 4.7: Inductively coupled plasma-mass spectrometric determination of arsenic, cadmium, chromium, lead, mercury, and other elements in food using microwave assisted digestion. United States Food and Drug Administration. https://www.fda.gov/downloads/food/foodscienceresearch/laborator ymethods/ucm377005.pdf. Accessed 23 Sept 2017 ISO/IEC 17025:2005 (2005) General requirements for the competence of testing and calibration laboratories Komorowicz I, Baralkiewicz D (2014) Arsenic speciation in water by high-performance liquid chromatography/inductively coupled plasma mass spectrometry-method validation and uncertainty estimation. Rapid Commun Mass Spectrom 28:159–168 Onishi H, Sandell EB (1955) Geochemistry of arsenic. Geochim Cosmochim Acta 7:1–33

Determination of Select Heavy Metals in Air Samples from Aurangabad City

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Onkar Jogdand, N. N. Bandela, Geetanjali Kaushik, and Arvind Chel

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Study Location . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Materials and Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chemical Analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Meteorological Observations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Results and Discussion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cross-References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

Living organisms require trace amounts of certain heavy metals, including cobalt, copper, manganese, molybdenum, strontium, and zinc but excessive levels can be detrimental to the organism. Other heavy metals such as mercury, lead, and cadmium are known to have vital impact on organisms; however, their accumulation over time in the bodies of mammals can cause serious illness. The particulate heavy metals can have severe toxic and carcinogenic effect for humans when inhaled in higher concentration. Therefore monitoring of heavy metals present in particulate matter is an important environmental issue. In this work, the atmospheric concentrations of selected heavy metals including Lead (Pb), Cadmium (Cd), Nickel (Ni), Manganese (Mn), and Zinc (Zn) were measured for O. Jogdand · N. N. Bandela Department of Environmental Sciences, Dr. Babasaheb Ambedkar Marathwada University, Aurangabad, Maharashtra, India G. Kaushik (*) · A. Chel MGM’s Jawaharlal Nehru Engineering College, Mahatma Gandhi Mission, Aurangabad, Maharashtra, India e-mail: [email protected] © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_164

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different sampling sites in Aurangabad city of Maharashtra. High volume air samplers and glass fiber filters were used to collect the samples. The collected samples were digested using a mixture of analytical grade hydrogen fluoride and analyzed to evaluate the levels of heavy metals by Atomic Absorption Spectrophotometry. Five heavy metals (Pb, Ni, Mn, Cd, and Zn) were monitored. Locations namely Kranti Chowk, Railway Station, Waluj Industrial area, SB College, Gulmandi Chowkand Harshul T Point have revealed high concentrations of selected heavy metals. To determine the emission sources of these metals, it is recommended to undertake more detailed and comprehensive study. Keywords

Heavy metals · Air pollution · Atomic absorption spectrophotometer · Particulate matter · Filters

Introduction World over air pollution is a public health problem. In 2012 air pollution was declared as the largest environmental health risk with almost seven million deaths globally attributed to it (WHO 2014). Data from India’s major regulator the Central Pollution Control Board (CPCB) revealed that 77% of Indian urban clusters clearly exceeded the National Ambient Air Quality Standard (NAAQS) for respirable suspended particulate matter (PM10) in 2010 (CPCB 2014). It is quite alarming to note that the satellite measures of fine particulates created for the entire India reveal that our populations living both in urban and rural areas are exposed to hazardously high levels of particulates. Almost 670 million people comprising 54.5% of the population reside in regions that do not meet the Indian NAAQS for fine particulate matter (Greenstone et al. 2015; Dey et al. 2012). Numerous studies have revealed a consistent correlation for particulate matter concentration with health than any other air pollutant. Studies show a statistically significant correlation between mortality and ambient particulate matter concentration (Lee et al. 2006). Particulate Matter (PM) refers to tiny particles which remain suspended in air, in the form of either solid or liquid droplets which originate from various sources that pollute the ambient air. Particulate matter comprises of various organic and inorganic components; the major components include acids, ammonia, sodium chloride, black carbon, water, and mineral dust. These respirable particulates having aerodynamic diameter  10 μm (PM 10) are an important part of the atmosphere. These particles have a high probability of deposition deeper into the respiratory tract and are likely to trigger respiratory diseases such as asthma, bronchitis, cardiopulmonary infections (Apte et al. 2011; WHO 2014). These particles have also been implicated as carriers of toxic air pollutants including heavy metals and organic compounds (Satsangi et al. 2011). Particulate heavy metals can have severe toxic and carcinogenic effect for humans when inhaled in higher concentration (Panne et al. 2001). Living organisms require trace amounts of some heavy metals, including cobalt, copper, manganese,

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molybdenum, strontium, and zinc, but excessive levels can be detrimental to the organism. Other heavy metals such as mercury, lead, and cadmium are known to have vital or beneficial effect on organisms, and their accumulation over time in the bodies of mammals can cause serious illness (Srivastava and Majumder 2008; Hsan 2008). The main input for many elements in the atmosphere is related to particle emission processes. Regarding trace metals (Pd, Cd, Zn) anthropogenic sources play a more significant role than natural sources, such as continental dust, salt spray, and biogenetic particles (Silvia et al. 2004; Katja et al. 1998; Figen et al. 2000; Munir and Shaheen 2008). Atmospheric metal concentration shows spatial variation since they are dependent on the distances to the sources and of the relative importance of local sources. Seasonal variation has been observed in atmospheric metal concentration in temperate or cold climates, but their levels are also influenced by meteorological variables such as wind speed and direction. Short-term differences of atmospheric metal concentration have been observed in a day-to-day or even an hour-to-hour basis (Munir and Shaheen 2008). Trace quantities of heavy metals are found in fossil fuels, and they are released into the atmosphere following combustion processes, including power generation and emissions from vehicles. Industrial processes, including the manufacture of steel and iron, and other metallurgical and chloralkali industries are also significant sources of heavy metals. Different industries release different metals, for example, lead emissions, which were previously almost completely from road transport, are now dominated by processes in the iron and steel sector. The largest source of arsenic is the burning of wood which has been treated with copper-chrome arsenate. Noncombustion sources of heavy metals include demolition of buildings, corrosion and abrasion of sources such as road surfaces, tire and brake wear. It is important to note that in addition to these anthropogenic sources, heavy metals are also released into the atmosphere from natural sources including volcanoes, forest fires, sea-spray, and windblown soil particles. As they are chemical elements, heavy metals do not degrade. This means that any metals which are released to the environment have the potential to become resuspended in the atmosphere, for example, windblown particles of soil and road dust (Arora et al. 2017). Several heavy metals have been determined in air samples in different countries such as Brazil (Silvia et al. 2004), Denmark (Karl et al. 2002), Portugal (Vasconcelos and Tavares 1997), India (Khillare et al. 2004), Spain (Moreno-Grau et al. 1997, 2000; Mateu et al. 1999), Egypt (Abdel-Shafy et al. 1992), and Russia (Drobyshev and Emelina 2001), not to mention national and international continuous monitoring programs for regulatory purposes like the European Monitoring and Evaluation Program. Various studies have been reported in the literature regarding heavy metal concentrations in air within Indian cities (Pandit et al. 2013; Jha 2010; Singh et al. 2008; Dubey et al. 2012; Gupta and Karar 2006). The main aim of this research is to determine the levels of heavy metals including Pb, Ni, Mn, Cd, and Zn in air samples collected from 19 different sites in Aurangabad city, then compare the levels of heavy metals in the various sites, and deduce causes for the presence of the heavy metals in air within the study area. To the best of our knowledge this study would be the first to provide valuable preliminary data on ambient concentrations of heavy metals in Aurangabad city.

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Study Location Aurangabad city in Maharashtra is the site of headquarters of Marathwada region (Fig. 1). It has a rich historical background and is a popular tourist place for both domestic and international tourists. The city is situated at a latitude of 19 530 5900 North and longitude 75 200 East. Aurangabad’s area is about 138 km 2.The climate of Marathwada region is generally hot and dry. The average temperature for day ranges from 27.7  C to 38.0  C while it ranges from 26.9  C to 20.0  C during the night. Average annual rainfall in the city and adjoining areas is 725.8 mm while the relative humidity is extremely low in this region for a major part of the year and ranges between 35% and 50%, while it is the highest (85%) during monsoon. The total area under forest cover is about 557 km2 which comprises of only 7.6% area of the total land area in Aurangabad (Kaushik et al. 2016). Recent census data revealed that the population of the city is about 1,500,000 (AMC 2011).The city boasts of a total number of industrial units (small-, medium-, and large-scale) as about 1020, and almost 35,000 workers find their employment in these units (Bhosale et al. 2010). The rapid industrial growth of Aurangabad has resulted in urbanization of the city and has also increased air pollution. The number of vehicles has also increased making a significant contribution to the vehicular traffic. The sampling sites according to their different land use patterns, populations, and traffic densities have been selected for monitoring particulate matters. The measurements of PM10 have been carried out on the terrace of the building (above 3–5 m above the ground) at each site (Table 1).

Fig. 1 Location of Aurangabad city (Latitude 19 530 5900 North and Longitude 75 200 East) (Source: Environment Status Report Aurangabad (2013–2014))

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Table 1 Location of monitoring sites Location 1. Seven Hill 2. Doodh Dairy 3. Gulmandi 4. Kranti Chowk 5. (MIDC) Chikalthana 6. CIDCO – Bus stand 7. Mill corner 8. Railway Station 9. Harsul T Point 10. JNEC – MGM campus 11. SB College, Aurangpura 12. Airport, Chikalthana 13. TV Center 14. City Chowk 15. MIDC – Waluj 16. Beed bypass 17. Gajanan Mandir Chowk

Description Near to Flyover and commercial complex Site of office of government entity, MSEDCL

Traffic density High High

Land use pattern Mixed Commercial

Busy densely populated market area Near to Flyover and commercial complex

Medium High

Mixed Commercial

Open industrial area with only few units functioning City bus stand buzzing with commuters and bus drivers and cleaning staff City Police Commissioner Office City Railway Station with high rush of visitors and also the residents Outskirt of city interstate buses, vans pass through the area Educational institution with around 8000 students and 500 faculty members Unpaved roads

Medium

Industrial

High

Commercial

High High

Mixed Commercial

High

Mixed

Medium

Institution

High

Mixed

Airport is on city outskirts, large open area

Low

Airport

Government TV transmission center

Mixed

Densely populated residential area with market

Highmedium Medium

Over 1200 industries

High

Industrial

City outskirts where vehicles meant for other locations bypass the city. Heavy truck traffic Center of city, densely populated

High

Commercial

High

Mixed

Mixed

Materials and Methods PM10 monitoring was carried out in selected city locations during December, 2015–January, 2016. The samples were collected for 24 h. PM10 samples were collected on Whatman filter papers with the help of a Respirable Dust Sampler (Model – APM 460 DXNL, Envirotech, New Delhi). The high volume sampler was operated at a flow rate of 1.1 m3/min. Field blanks were also collected. Before beginning the sampling all the filter papers were preweighed with the help of a Metler analytical weighing balance. Then the filter papers were desiccated for the duration of 24 h. In order to avoid any sort of contamination, conditioned and

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preweighed filter papers were kept in a zip lock polybag for taking to the field for sampling. Prior to loading the filter papers on the sampler, the initial manometer and timer readings were noted. Subsequently the filter papers were loaded on the sampler and after ensuring that the sampler was properly screwed the sampler was started. At the end of sampling period, the loaded filter paper was removed with the help of forceps, wrapped in aluminum foil, and placed in a zip lock polybag. In the laboratory the filter paper was conditioned and was again weighed to determine the PM10 mass concentration (Satsangi et al. 2011).

Chemical Analysis A portion (100 cm2) of the filter sample was cut into small pieces with stainless steel scissors before being subjected to digestion. Papers were then digested. HF was added and the beaker was covered with a watch glass, then the mixture was heated to near dryness at the hot plate. The sample was allowed to cool to room temperature, then transferred into a 100-mL volumetric flask and diluted to volume with distilled water. The heavy metal concentrations were determined by flame atomic absorption spectrometry with air/acetylene burner. Each result is the average of three readings.

Meteorological Observations During the air quality monitoring the wind direction (WD) and speed (WS), temperature, relative humidity (RH), and the rainfall were also recorded. The average minimum and maximum temperatures were 14  C and 30  C. Average minimum and maximum RH were 26 and 51, while the prominent WD was East with WS ranging from 6 to 14 km/hr. During the period of sampling there was no rainfall (IMD 2016) (Table 2).

Results and Discussion The mean quantity of Ni in PM10 of different selected sites was in order of Cidco bus stand > Kranti Chowk > Seven Hill > Harshul T Point > SB College. Concentration of Ni was found highest at Cidco bus stand site probably due to high vehicular movements. In a similar manner Kranti Chowk, Seven hill, and Harshul T Point have a high Ni concentration owing to high vehicular traffic. Use of Ni for plating the external part of a motor vehicle and as a fuel additive may be a probable source of Ni in the atmosphere. At six other locations Nickel was found below the detection limit. The higher concentration at Harshul T Point is due to a combination of transportation and civil construction work impact in this area.

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Table 2 Heavy metal concentrations r.No. Sampling locations

Nickel (Ni) ppm

Cadmium Manganese Lead (Cd) ppm (Mn) ppm (Pb) ppm

Zinc (Zn) ppm

1

Gajanan Maharaj Chowk

n.d.

n.d.

0.07

0.45

2.19

2

Beed Bypass Road

0.30

n.d.

0.09

0.25

2.51

3

Waluj Industrial Area

n.d.

n.d.

0.19

0.93

2.37

4

City Chowk Police Station

0.30

n.d.

0.11

0.73

2.55

5

TV Center Chowk

0.21

n.d.

0.01

1.21

n.d.

6

Airport, Chikalthana

n.d.

n.d.

0.05

1.13

2.44

7

S. B. College, Aurangpura

0.75

0.01

0.10

0.61

2.28

8

Railway Station Chowk

n.d.

0.01

0.13

0.57

2.38

9

Seven Hill, Aurangabad

1.02

n.d.

0.15

n.d.

2.48

10

Cidco Bus Stand, Aurangabad

2.09

n.d.

n.d.

0.69

1.48

11

Gulmandi Chowk

0.21

0.01

0.07

1.42

2.42

12

Hotel Amarpreet Chowk

0.57

n.d.

0.15

0.85

2.33

13

Kranti Chowk

1.20

0.01

0.11

0.45

2.66

14

Mill Corner Police Headquarter

n.d.

n.d.

0.07

n.d.

1.35

15

MIDC, Chikalthana

n.d.

0.01

0.09

n.d.

2.35

16

JNEC – West Wing

0.66

n.d.

0.07

n.d.

2.16

17

MGM Hospital Nursing College

0.11

n.d.

0.07

n.d.

2.34

18

Harshul T Point

0.84

n.d.

0.13

n.d.

2.22

19

MGM – Jr. College Clover Dale School

0.57

n.d.

n.d.

n.d.

2.50

n.d. denotes not detected 3 2.5

Nickel (Ni) PPM Cadmium (Cd) PPM

2

Manganese (Mn) PPM 1.5

Lead (Pb) Zinc (Zn) PPM

1 0.5 0 1

2

3

4

5

6

7

8

9 10 11 12 13 14 15 16 17 18 19 20

Cd was detected at SB College, Railway Station, Gulmandi Chowk, Kranti Chowk, and MIDC Chikalthana. Vehicular emissions, automobile lubricants, wear and tear of tires, and construction activities are the possible sources of Cd in ambient

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air (Khillare and Sarkar 2012). Various products like batteries, toys, etc., contain Cd. It is also more common in household garbage. In a previous study Cd was found relatively enriched in Jharia street dust (Rout et al. 2012). The mean quantity of Mn in PM10 of different selected sites was in order of Waluj Industrial Area > Seven Hill > Railway Station >Harshul T Point. Mn is used in unleaded gasoline to enhance octane rating and lessen engine knocking (Roy et al. 2012). Vehicular emission may account for the higher concentration of Mn in the study area. Previous studies reported the earth crust/windblown soil as a contributor of Mn in PM10 (Shah et al. 2006; Dubey et al. 2012). Coal also contains Mn in trace amount so contribution of coal-burning activities and airborne coal dust cannot be neglected. High concentrations of Lead (Pb) were detected in Gulmandi Chowk, TV center Chowk, Airport, Waluj Industrial area, and Hotel Amarpreet Chowk. This is mainly due to the heavy vehicular transportation. These values at the first three locations are slightly higher than the Pb National Ambient Air Quality Standard (NAAQS) (1.0 μg/m3) prescribed by the Central Pollution Control Board (CPCB) of India (CPCB 2009). Lead is very toxic to human health as well as for the environment. Its relatively high toxicity should represent a concern for the city. High concentrations of Zn are observed at most of the locations monitored. Zn is used in lubricating oil and Zn is also released from vehicular activities such as tire wear (ATSDR 2005; TR MEF 2010).

Conclusion The main aim of this research is to determine the levels of heavy metals including Pb, Ni, Mn, Cd, and Zn in air samples collected from 19 different sites in Aurangabad city, then compare the levels of heavy metals in the various sites, and deduce causes for the presence of the heavy metals in air within the study area. Locations namely Kranti Chowk, Railway Station, Waluj Industrial area, SB College, Gulmandi Chowk, and Harshul T Point have revealed high concentrations of selected heavy metals mainly due to resuspension of road dust by vehicular turbulence. Higher levels of air pollution on account of growing number of automobiles, increase in the number of restaurants, roadside eateries, and power generators have also been contributing factors. Waluj is an industrial area and a cause of concern from declining air quality point of view. Waste burning and construction activities across the city might have served as other sources. To determine the sources of these metals, it is recommended to undertake more detailed and comprehensive study.

Cross-References ▶ Air Quality Status and Management in Tier II and III Indian Cities: A Case Study of Aurangabad City, Maharashtra ▶ Indoor Air Pollution Around Industrial Areas and Its Effect: A Case Study in Delhi City

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▶ Modern Air Pollution Prevention Strategies in the Urban Environment: A Case Study of Delhi City ▶ Status of Particulate Matter Pollution in India: A Review

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Khillare PS, Sarkar S (2012) Airborne inhalable metals in residential areas of Delhi, India. Atmos Pollut Res 3:46–54 Khillare PS, Balachandran S, Bharat RM (2004) Special and temporal variation of heavy metal in atmospheric aerosol of Delhi. Environ Monit Assess 90:1–21 Lee SC, Cheng Y, Ho KF, Cao JJ, Louie PKK, Chow JC (2006) PM1.0 and PM2.5 characteristics in the roadside environment of Hong Kong. Aerosol Sci Technol 40:157–165 Mateu J, Mirabo FB, Forteza R, Cerda V, Colom M, Oms M (1999) Heavy metals in the aerosols collected at two stations in Mallorca (Spain). Water Air Soil Pollut 112:349–363 Moreno-Grau S, Perez-Tornell A, Moreno-Grau J, Moreno-Clavel J (1997) Determination of lead in aerosol sample collected on glass fiber filters by an improved atomic absorption spectrometry method. Water Air Soil Pollut 96:145–153 Moreno-Grau S, Tornell A, Bayo J, Moreno J, Angosto JM, Moreno-Clavel J (2000) Particulate matter and heavy metals in the atmospheric aerosol from Cartagena, Spain. Atmos Environ 34:5161–5167 Munir HS, Shaheen N (2008) Annual and seasonal variations of trace metals in atmospheric suspended particulate matter in Islamabad. Pakistan. Water Air Soil Pollut 190:13–25 NAAQS (2009) Available at http://cpcb.nic.in/National_Ambient_Air_Quality_Standards.php. Accessed July 2017 Pandit RJ, Patel B, Kunjadia PD, Nagee A (2013) Isolation, characterization and molecular identification of heavy metal resistant bacteria from industrial effluents, Amala-khadiAnkleshwar, Gujarat. Int J Environ Sci 3(5):1689–1699 Panne U, Neuhauser RE, Theisen M, Fink H, Niessner R (2001) Analysis of heavy metal aerosols on filters by laser-induced plasma spectroscopy. Spectroacta 56:839–850 Rout TK, Masto RE, Ram LC (2012) Assessment of human health risks from heavy metals in outdoor dust samples in a coal mining area. Environ Geochem Health. https://doi.org/10.1007/ s10653-012-9499-2 Roy P, Sikdar PK, Singh G, Pal C (2012) Source apportionment of ambient PM10. A case study from a mining belt of Orissa. Atmósfera 25(3):311–324 Satsangi GP, Kulshrestha A, Taneja A, Rao PSP (2011) Measurements of PM10 and PM 2.5 aerosols in Agra, a semi-arid region of India. Indian J Radio Space Phys 40:203–210 Shah MH, Shaheen N, Jaffar M (2006) Characterization, source identification and apportionment of selected metals in TSP in an urban atmosphere. Environ Monit Assess 114:573–587 Silvia MS, Annibal DPNAD, Pereira N, Emmanoel VSF, Martha TA (2004) Short term and spatial variation of selected metals in the atmosphere of Niteroi City, Brazil. Microchem J 78:85–90 Singh R, Barman SC, Negi MPS, Bhargava SK (2008) Metals concentration associated with respirable particulate matter (PM10) in industrial area of eastern U.P. India. J Environ Biol 29(1):63–68 Srivastava NK, Majumder CB (2008) Novel biofiltration methods for the treatment of heavy metals from industrial wastewater. J Hazard Mater 151:1–8 T. R. Ministry of Environment and Forestry, Air Quality Assessment and Management 2010 Vasconcelos MTSD, Tavares HMF (1997) Atmospheric metal pollution (Cr, Cu, Fe, Mn, Ni, Pb and Zn) in Oporto City derived from results for low-volume aerosol samplers. Sci Total Environ 212:11–20 WHO (2014) Seven million premature deaths annually linked to air pollution, World Health Organization. Accessed on 15 Jan 2017. http://www.who.int/mediacentre/news/releases/2014/ air-pollution/en/

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Recent Advances in Membrane Extraction Techniques for Environmental Samples Analysis

Hadi Tabani, Saeed Nojavan, Kamal Khodaei, and Alireza Bazargan

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Hollow Fiber Liquid-Phase Microextraction (HF-LPME) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Electromembrane Extraction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Emulsion Liquid Membrane . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Preparation of Emulsion Liquid Membranes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Applications of the Emulsion Liquid Membrane Procedure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Polymer Inclusion Membrane . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Applications of Polymer Inclusion Membrane Procedure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

1210 1211 1213 1217 1222 1223 1227 1229 1233

Abstract

The quantification of analytes present in environmental samples at trace levels needs a preliminary stage of isolation and enrichment of compounds, and this issue is a hot challenge in separation science. Usually, a direct analysis of environmental samples to determine analytes is impossible. In recent years, attempts have been made to develop new methodologies for the analysis of environmental samples. Many methods have been developed for extracting H. Tabani (*) · K. Khodaei Department of Environmental Geology, Research Institute of Applied Sciences (ACECR), Shahid Beheshti University, Tehran, Iran e-mail: [email protected] S. Nojavan Department of Analytical Chemistry and Pollutants, Shahid Beheshti University, G. C., Evin, Tehran, Iran e-mail: [email protected] A. Bazargan (*) Department of Civil Engineering, K. N. Toosi University of Technology, Tehran, Iran e-mail: [email protected] © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_165

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different compounds from environmental samples. Among the emerging techniques, membrane-based extraction is an efficient alternative to classical pretreatment methodologies. Membrane techniques have emerged as an efficient method for the extraction of both ionizable and nonionizable analytes in complex matrixes. The main advantages of these techniques include high selectivity, the use of minimal organic solvents, and acceptable clean-up efficiency, with high enrichment factors. Based on the recently published literature data, this chapter provides an update of the advantages and applications of membrane extraction techniques for analysis of environmental samples. Keywords

Hollow Fiber Liquid-Phase Microextraction (HF-LPME) · Electromembrane · Polymer inclusion membrane · Emulsion liquid membrane

Introduction Sample preparation has always been a somewhat neglected part of sample analysis. The pretreatment allows for the migration of analytes from the primary matrix to the receiving matrix with the simultaneous removal of macromolecules and other matrix constituents that may interfere with the detection system. Frequently, liquid–liquid extraction (LLE) is used as a conventional sample preparation method for extraction of analytes from complex matrices (Carasek and Merib 2015; Płotka-Wasylka et al. 2016). The limitations of LLE techniques include long extraction times as well as the use of large amounts of hazardous extraction solvents which are not environmentally friendly. Also, often small enrichment factors and selectivity are achieved with LLE. An alternative and complementary way for addressing this issue is applying membrane technology. Membrane-based methods have brought about powerful options in separation science, having some distinct advantages when compared to other classical sample preparation methodologies. These methods have acceptable enrichment factors, good cleanup efficiency, and selectivity while using only little or no organic solvents (Jakubowska et al. 2005; Barri and Jonsson 2008). In membrane-based methods, generally, a membrane (polymeric or liquid phase) is placed between two liquid phases and acts as a selective barrier. One of these phases, the donor phase (DP), is containing the analyte in a complex matrix. The analytes are then extracted to the other side of the membrane which is called the acceptor phase (AP). Migration of analytes in the membrane process results from the differences in the rate of mass transfer of analytes across the membrane, and it depends on the type of the driving force (van Hout et al. 2003; Jonsson et al. 2003). The main driving forces are based on electric potential difference (ΔE), concentration difference (ΔC), and pressure difference (ΔP). Up to now, several kinds of membranes with different structures and transport mechanisms have been introduced for extraction of analytes from environmental samples. The distinction between membranes mainly originates from their diverse preparation methods or raw materials. In this chapter, a comprehensive literature study on the principles and applications of membrane extraction techniques based on

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liquid and polymeric membranes for analysis of analytes from environmental samples is discussed and outlined.

Hollow Fiber Liquid-Phase Microextraction (HF-LPME) Generally, two extraction modes, namely two-phase extraction and three-phase extraction, are applied for HF-LPME (Amdany et al. 2015; Pedersen-Bjergaard and Rasmussen 1999; Vora-adisak and Varanusupakul 2006; Fakhari et al. 2013a). Two-phase (aqueous-organic) LPME is used for extraction of analytes with high solubility in organic solvents. In this mode, neutral compounds are extracted from the DP into the membrane and subsequently into the organic solvent placed inside the hollow fiber as the AP (Fakhari et al. 2013a; Li and Hu 2007; Varanusupakul et al. 2007; Chiang and Huang 2007; Ma et al. 2011). The equipment used for the extraction is illustrated in Fig. 1. In this method, the AP is in direct contact (as immobilized in the membrane pores) with the aqueous sample: analyte molecules can diffuse directly into the AP. After the extraction, the AP containing extracted organic phase is injected into an analytical instrument for analysis. Conversely, in the three-phase (aqueous-organicaqueous) mode, the analytes migrate from the DP into an organic phase which is immobilized inside the pores of the membrane, and then, they are extracted into another liquid, the AP (Xia et al. 2007; Zhang et al. 2011; Li and Hu 2011; Payan et al. 2011). Both two- and three-phase modes of HF-LPME have been applied for extraction of analytes from environmental samples, and the present section focuses on recent applications of the HF-LPME procedure in this area.

Fig. 1 Schematic illustration of the setup for EME. (Reproduced with permission from Elsevier (Fakhari et al. 2013a))

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One of the first publications which proposed using HF-LPME for analysis was by Pedersen-Bjergaard and Rasmussen (Pedersen-Bjergaard and Rasmussen 1999). In that work, three-phase LPME was used for preconcentration of methamphetamine and subsequent analysis by capillary electrophoresis (CE). The study showed the excellent potential of this technique for sample preparation. In the case where the inner diameter of the hollow fiber does not tightly fit with the outer diameter of the microsyringe needle, the hollow fiber might be separated from the microsyringe needle during the extraction. In order to address this issue, Berhanu et al. developed a new design for the quantification of three pesticides (fenthion, chlorpyrifos, and diazinon) from environmental samples. In their method, after impregnating the hollow fiber with n-undecane as the organic extraction solvent, it was connected to a copper wire with appropriate outer diameter, in order to get it tightly attached to the hollow fiber. This kind of configuration was found to be a simple method with acceptable limit of detections (LODs) in the range of 15–80 ng L1 (Berhanu et al. 2008). In an interesting study, Raharjo et al. used nylon instead of polypropylene as a new support for determining quinalphos and methidation (Raharjo et al. 2009). In another study, Nyoni et al. (2011) introduced a membrane which was made from silicone rubber for quantification of ionizable organic compounds in environmental samples. In this method, no organic solvents were used; hence, this method can be classified as a green extraction technique. Also, the device was solid and it provided stable and reputable results during the extraction with a longer lifetime compared to conventional membranes. In addition, there have been some reports about using polyvinylidene difluoride (de Jager and Andrews 2001), polysulfone (Vora-adisak and Varanusupakul 2006), and cellulose (Zhang et al. 1996) as membranes for HF-LPME. Generally, when a pure solvent is used as the extraction solvent, analytes with log P values lower than 2.0 are not efficiently extracted. It has been reported that ion-pairing reagents such as tri-n-octylphosphine oxide (TOPO), tri-n- butyl phosphate (TBP), and di (2-ethylhexyl) phosphoric acid (DEHP) can alter the selectivity of the HF-LPME method. For example, the addition of DEHP to membrane improves the extraction of polar basic analytes with low log P values (log P < 1) and, on the contrary, decreases the extractability of compounds with high log P values (log P > 2). Trtić-Petrović proposed a method with mixing 10% of TOPO and 10% of TBP in n-hexyl ether to form the extraction solvent for the analysis of 16 pesticides from environmental samples (Trtić-Petrović et al. 2010). The results showed that extraction efficiency of polar pesticides improved with the addition of the mentioned reagents to the extraction solvent. Also, the HF-LPME technique has been used for the quantification of metals in environmental samples. Lopez-Garcia et al. determined trace concentrations of mercury by adding a solution of 1-(2-pyridylazo)-2-naphthol (PAN) to the DP (Lopez-Garcia et al. 2012). At first, a reversible complex formation between mercury ions and PAN occurred and then diffused across the membrane to the AP. Peng et al. reported application of ionic liquids as a new extraction solvent in HF-LPME. In this study, 1-octyl-3- methylimidazolium hexafluorophosphate was reinforced within the

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wall pores of the hollow fiber, and then chlorophenols were extracted from environmental samples (Peng et al. 2007). Further improvement in extraction efficiencies was achieved with the combination of solid-phase extraction (SPE) with LPME. Hu et al. (2009) introduced a novel technique by adding molecular imprinted polymer (MIP) to the membrane for the extraction of triazines from environmental samples. In this method, the mentioned compounds were extracted by toluene as the organic phase in the membrane and then adsorbed by the MIP. In another recent report, Yamini et al. (Moradi et al. 2012) proposed an HF-LPME method for quantification of halogenated amines from water samples. In this study, supramolecular solvents were used as extraction solvents in the membrane, and the results showed that these kinds of solvents can be good alternatives to conventional organic solvents. Recently, some scientific reports on the automation of LPME instruments have been developed in the literature (Maya et al. 2014; Vallecillos et al. 2012; Li et al. 2015). To conclude this subject, Table 1 summarizes recent applications of HF-LPME in environmental samples.

Electromembrane Extraction Although the problem of organic solvent instability has been resolved by the HF-LPME method, long extraction times are needed in order for the process to achieve equilibrium (Pedersen-Bjergaard and Rasmussen 2006). In order to increase flux of analytes across the membrane, transfer has also been carried out by application of an electric field in a technique called electromembrane extraction (EME) (Fig. 2) (Pedersen-Bjergaard and Rasmussen 2006; Fakhari et al. 2013b; Middelthon-Bruer et al. 2008; Eibak et al. 2010; Koruni et al. 2014). The main mass transfer mechanism for the migration of analytes across the membrane has been evaluated by the Nernst–Planck equation (Middelthon-Bruer et al. 2008): Di Ji ¼ h



ν 1þ lnχ



χ1 ðCih  Ci0 expðνÞÞ χ  expðνÞ

where h is the thickness of the membrane, Di denotes the diffusion coefficient for the analyte, Cih represents the analyte concentration at the membrane/DP interface, and Ci0 is the analyte concentration at the AP/membrane interface; ν is a dimensionless driving force defined as: ν¼

zi eΔ∅ KT

where ΔØ is the electrical potential across the membrane, zi is the charge of the analyte, e is the elementary charge, T is the absolute temperature, and K is Boltzmann’s constant.

Aqueous samples

Aqueous samples

Aqueous samples, food Lake water

Water samples

Aqueous samples

Environmental samples Natural water samples

Aqueous samples

Haloacetic acids

Salicylates

Organometallic

BTEX

Sulfonamides

Ni (II), Pb (II)

Hg (II)

Cr (III)

Haloethers

V (IV), V (V)

Matrix Seafood, environmental samples Natural waters

Target analytes Organomercury

Polypropylene hollow fiber Polypropylene hollow fiber Polypropylene hollow fiber Polypropylene hollow fiber Polypropylene hollow fiber Polypropylene hollow fiber Polypropylene hollow fiber Polypropylene hollow fiber Polypropylene hollow fiber Polypropylene hollow fiber

Kind of membrane Polypropylene hollow fiber

Table 1 Different applications of HF-LPME in environmental samples

Toluene

[C4MIM][BF4]

[C6MIM][PF6]

1-Octanol

[BMIM][PF6]

Toluene

Tributyl phosphate

1-Octanol

1-Octanol

Carbon tetrachloride

Extraction solvent Toluene

ETAAS

F AAS

ET AAS

HPLC-DAD

GC-FID

GC-ECD

CE-UV

HPLC-UV

ETV-ICPOES GC-ECD

Detection system HPLC-UV

0.06

0.7

0.02

0.0003–0.033

2.7–4.0

0.55–4.30

0.68–6.90

0.6–1.2

0.1–18

0.086, 0.071

LOD (ppb) 0.3–3.8

Lopez-Garcia et al. (2012)

Abulhassani et al. (2010) Zeng et al. (2012)

Payan et al. (2011)

Chiang and Huang (2007) Ma et al. (2011)

Li and Hu (2011)

Varanusupakul et al. (2007) Zhang et al. (2011)

Li and Hu (2007)

References Xia et al. (2007)

1214 H. Tabani et al.

Fish, rice

Water

Water

Hg (II)

Inorganic Sb

Tricyclic antidepressant drugs Pesticides

Atrazine

Natural water

Industrial and fresh orange juice Water, sludge water

Aqueous samples

Se (IV)

Environmental samples Aqueous samples

Estrogens

Aqueous samples

Aqueous samples

Endocrine disruptor

Anti-inflammatory drugs Pyrethroids pesticides Bisphenol A

Aqueous samples

Phenols

Polypropylene hollow fiber Polypropylene hollow fiber Polypropylene hollow fiber Polypropylene hollow fiber Polypropylene hollow fiber Polypropylene hollow fiber Polypropylene hollow fiber Polypropylene hollow fiber Polypropylene hollow fiber Polypropylene hollow fiber Polypropylene hollow fiber Polypropylene hollow fiber Toluene and ethyl acetate (85:15, v/v)

n-Dodecane

1-Octanol

Propylbenzoate

1-Octanol

1-Octanol

1-Octanol

1-Octanol

Toluene:octanol

Dihexyl ether

Dihexyl ether

HPLC–UV

HPLC-DAD

UPLC-MS/ MS GC–MS

HPLCfluorescence HPLC-UV

HPLC-DAD

300

3–350

0.08–0.2

1.1

0.012

0.02–0.1

0.2

0.002–0.012

0.5–1.25

0.055–1.46

0.52–0.54

0.14–0.29

Hu et al. (2009)

Margui et al. (2013) Ghambarian et al. (2012) Wang et al. (2012)

Ensafi et al. (2012)

Saleh et al. (2009)

San Román et al. (2012) Tan et al. (2012)

Zhang et al. (2013)

Villar-Navarro et al. (2012) Villar-Navarro et al. (2013) Chen et al. (2013)

50 Recent Advances in Membrane Extraction Techniques for Environmental. . . 1215

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Fig. 2 Schematic illustration of the setup for EME. (Reproduced with permission from Elsevier (Fakhari et al. 2013b))

Generally, efficiency and the selectivity of the EME are controlled by the chemical composition of the membrane and the type of organic extraction solvent. Based on published papers, basic analytes with log P values approximately higher than 2.0 have been extracted by nitro aromatic solvents such as 2-nitrophenyl octyl ether (NPOE) (Eibak et al. 2010). When pure NPOE is used as the organic solvent, basic analytes with polarities lower than 2.0 are not efficiently extracted. Addition of only a low concentration of hydrophobic ion-pair reagents such as DEHP to pure NPOE can promote the flux of basic compounds with low polarity and, on the contrary, decreases the extractability of compounds with high log P values (Eibak et al. 2010; Koruni et al. 2014). Up to now, several modifications have been proposed to improve the mass transfer of ionic analytes through the membrane. For example, crown ethers were added to NPOE for the extraction of potassium ions. The results showed that NPOE containing 1% w/v of dibenzo-18-crown-6 showed acceptable selectivity for extraction of K+ and had negligible recoveries for extraction of other cations (Slampova et al. 2014). In another study, an ionic liquid ([C6MIm] [PF6]) as a new extraction solvent was added in the membrane for the extraction of some basic drugs (Sun et al. 2014). Fakhari et al. proposed a new approach to improve the extraction of herbicides such as 2,4-DB and Dicamba from environmental samples, where nonionic

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surfactants were added into the sample to form a hydrophobic layer around the herbicides, so that the mass transfer of analytes towards the membrane was promoted (Bagheri et al. 2016). Continuing with the improvement of this method, nanoparticles such as carbon nanotubes (CNTs) (Hasheminasab et al. 2013), fullerene (Atarodi et al. 2017), and N-doped graphene (ND-G) (Atarodi et al. 2016) were used to decorate the membrane in order to improve of the mass transfer. The results showed EME recoveries were increased in shorter time. Another way for enhancing mass transfer could be obtained by decreasing the boundary layer around the membrane. Seidi et al., by increasing the stirring rate and decreasing the charge density around the membrane, reduced the thickness of the double layer at the DP/membrane interface (Rouhollahi et al. 2016). In another study, Tabani et al. reported a new method by using a rotating electrode for stirring the AP (Asadi et al. 2016). With this new apparatus, extraction improved with agitation of the AP due to decreasing thickness of the Nernst’s diffusion film around the membrane/AP interface. These days, the introduction of new kinds of membranes as alternatives to polypropylene hollow fibers is a hot topic. Román-Hidalgo et al. (2017) introduced a polar nano-structured (Tiss ®-OH) sheet containing high density of –OH groups for the determination of highly polar analytes. Also, polyvinyldifluoride (PVDF) was applied as a membrane material in a Parallel-EME (Pa-EME) method (Drouin et al. 2017). Most recently, some new membranes based on gels have been introduced by Tabani et al. as green EME techniques (Tabani et al. 2017; Sedehi et al. 2018; Asadi et al. 2018) (Fig. 3). Another interesting development in EME is the use of microfluidic devices. Microfluidic sample preparation techniques are very interesting, and the development of these systems will surely flourish in the future. The first paper to downscale the EME procedure was done by Petersen et al. (2010). After that several papers were reported to miniaturize EME with on-chip dynamicEME and lab-on-chip systems (Drouin et al. 2016; Karami et al. 2017; Abdossalami Asl et al. 2015). This design allows one to carry out analyses with low amounts of DP and fast extraction because of a very short diffusion path. Table 2 summarizes recent environmental applications of EME for extraction of analytes.

Emulsion Liquid Membrane Owing to their high contact surface area for mass transfer, selective or accumulative removal, and the need for low amounts of organic solvents, emulsion liquid membrane (ELM) methods have been recognized as suitable methods for the separation of pollutants from environmental samples. In the ELM process, the stripping and extraction steps are combined into one step, leading to the simultaneous preconcentration of the target species (Cahn and Li 1974). There is also significant saving on the amount of equipment and materials required in the

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Fig. 3 Schematic illustration of preparation process of gel membrane. (Reproduced with permission from Elsevier (Tabani et al. 2017))

various stages of the liquid-liquid extraction. The ELM process includes three steps and can be up to 40% more cost-effective than the solvent extraction process (Frankenfeld et al. 1981). The first step involves preparation of an emulsion by mixing the membrane phase and the internal phase, such as water in oil (W/O). The second step is diffusion or transmission of dissolved species through the membrane from the donor phase to an acceptor phase through surface contact between the emulsion and the continuous phase. The third step involves emulsion and external phases’ sedimentation due to emulsion breaking to recover the membrane phase. Since the invention of the ELM method by Li in 1968 (Li 1968), application of this method to hydrometallogically recover heavy metals has attracted the attention of numerous researchers. The acceptor phase emulates an immiscible fluid membrane. Then the emulsion is dispersed in the solvent or feed and the mass transfer occurs from the feed to the internal acceptor phase. Liquid membranes could be aqueous or organic solutions; however, studies have mainly addressed water-in-oil emulsions. An experimental setup for the ELM technique is shown in Fig. 4 (Malik et al. 2012).

Salicylic acid Ketorolac Ketoprofen Naproxen Diclofenac Ibuprofen Chloroacetic acid Trifluoroacetic acid Dichloroacetic acid Phenyl acetic acid p-Hydroxyphenyl acetic acid

Li

Seawater

4-Chlorophenol 2,4Dichlorophenol 2,4,6Trichlorophenol Pentachlorophenol Amlodipine

Wastewater

Standard solution Wastewater

Water

Matrix Amniotic fluid

Target analytes Pb (II)

Polypropylene hollow fiber

Polypropylene hollow fiber Polypropylene hollow fiber Polypropylene hollow fiber

Kind of membrane Polypropylene hollow fiber Polypropylene hollow fiber

Toluene

1-octanol

1-octanol

NPOE

1-octanol

Extraction solvent Toluene

66 55 62 58 100 74 87–106

98

83

74

Recovery % 81.6–86.3

Table 2 Overview applications of electromembrane techniques in environmental samples

HPLC-UV

(continued)

Alhooshani et al. (2011)

Nojavan and Fakhari (2010) Strieglerova et al. (2011) Payán et al. (2011)

References Basheer et al. (2008) Lee et al. (2009)

Recent Advances in Membrane Extraction Techniques for Environmental. . .

0.0196 0.0403

0.0193

0.16 0.18 0.12 0.08 0.23 3.36 0.0007 0.0043

0.063

CE-C4D HPLC-DAD

3

0.1–0.4

LOD (ppb) 20

CE-UV

HPLC-UV

Detection system CE-UV

50 1219

Polypropylene hollow fiber Polypropylene hollow fiber

Environmental samples

Seawater

Polypropylene hollow fiber

Polypropylene hollow fiber

Culm

Environmental samples

Wastewater

Wastewater

River water

Polypropylene hollow fiber

Wastewater

Ibuprofen Naproxen 2,4-D 2,4-DB Dicamba Uranium

2,4-D 2,4-DB Dicamba Cr (III) Cr (IV) Basic red-18 Basic red-46 Basic red-51 Thorium

Polypropylene hollow fiber

Wastewater

Salbutamol Serbutaline

Kind of membrane Polypropylene hollow fiber Polypropylene hollow fiber

Matrix Water

Target analytes Trimipramine

Table 2 (continued)

1-octanol +5% DEHP

NPOE

1-octanol

NPOE +1% DEHP 1-octanol

1-octanol

NPOE +10% DEHP +10% TEHP 1-octanol + CNTs

Extraction solvent NPOE

40 39 39 31.1 47.2 48 81 78 93

90 94 72 71 70 54

53 43

Recovery % 66

UV- Vis

HPLC-UV

HPLC-UV

CE-UV

Fluorescence

CE-UV

CE-UV

HPLC-UV

Detection system CE-UV

0.3 0.5 0.4 5.4 2.8 0.75 0.30 0.30 0.29

1.5 1 10 10 15 0.1

10 5

LOD (ppb) 7

Khajeh et al. (2015)

Nojavan et al. (2013)

Safari et al. (2013)

Davarani et al. (2013) Tabani et al. (2013b)

Tabani et al. (2013a)

Hasheminasab et al. (2013)

References Fakhari et al. (2012) Rezazadeh et al. (2012)

1220 H. Tabani et al.

Polypropylene hollow fiber

Polypropylene hollow fiber Polypropylene hollow fiber Polyvinylidene fluoride Polypropylene hollow fiber Polypropylene hollow fiber

Wastewater

Water samples

River water

River water Sea water River water

River water

River water

As(III)

Pb (II)

Hg (II)

Hg (II)

Au

Polypropylene hollow fiber

Polypropylene hollow fiber

Water samples

Fluoride Chloride Nitrate Nitrite Bromide Phosphate Sulfate Ag (I) Cd (II) Co (II) Cu (II) Zn (II) Parabens

1-octanol

1-octanol+2.5% DEHP 1-octanol +20% DEHP 1-octanol +2% DEHP+ 1-octanol +2% DEHP+

1-octanol

1-octanol +0.5% DEHP+ 0.5% TEHP

1-heptanol

95.6

41–43

86–93.8

73.3–80

93–96

49 70 46 47 40 51 46 67.6 81.4 63.3 66.8 86.5 87–101

Graphite furnace atomic absorption spectrometry (GFAAS) UV- Vis

Anodic stripping voltammetry (ASV) Square wave voltammetry (SWV) UV- Vis

HPLC-DAD

Atomic absorption spectroscopy

IC

4.5

0.5

0.7

–

0.18

0.6 0.6 1.5 1.5 3.0 7.5 4.5 – – – – – 0.98–1.43

Khajeh et al. (2018)

Villar-Navarro et al. (2016) Kamyabi and Aghaei (2016) Hamsawahini et al. (2016) Fashi et al. (2017) Kamyabi and Aghaei (2017)

Davarani et al. (2015)

Nojavan et al. (2014)

50 Recent Advances in Membrane Extraction Techniques for Environmental. . . 1221

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Fig. 4 Experimental setup for ELM technique. (Reproduced with permission from Elsevier (Malik et al. 2012))

Preparation of Emulsion Liquid Membranes Formulation is the main part in the liquid membrane process. This step involves selection of a carrier, surfactant stripping agent, and diluents. Proper selection of these components and their formulation will guarantee the success of extraction. Various surfactants have been evaluated, yet only a few like Span 80 and ECA 4360 J have shown high efficiency (Ahmad et al. 2011). There is no specific requirement for the selection of diluents; however, it is advantageous to have high dissolution capacities for the extractor, high boiling temperature, and low solubility in the feed and stripping phases, as well as being inexpensive and nontoxic. So far, high speed lab-scale mixers or homogenizers have been used for emulsion preparation. Surfactants: Surfactants are amphipathic organic compounds, meaning that they include both hydrophobic (their tail) and hydrophilic (their head) groups. Therefore, they can dissolve in both organic and aqueous solvents. Surfactants reduce the surface tension of liquids such as water. They are also known to decrease the oil-water surface tension at the liquid-liquid interface. Stripping agent: Acids or bases can be applied as a stripping phase in the ELM process depending on the solute to be extracted. For example, Chiha et al. (2006) used a NaOH solution as the internal phase for removal of chromium (VI) from sulfuric acid aqueous solutions. Also, an H2SO4 solution has been applied for the

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removal of Cr (VI) (Saravanan et al. 2006). The solute extraction speed improves with an increase in the amount of stripping agent present in the emulsion. Extractant: The extractant will be added to the membrane phase and serves as a carrier. This carrier facilitates the transmission of target species through the membrane. Carriers used in metals transport can include acidic carriers (-COOH, -SO3H, or chelating groups) as well as basic carriers (such as N-H or ammonium salts). Emulsion breaking: To break a W/O emulsion, electrostatic emulsion breaking techniques are mainly employed (Lu et al. 1997). Other emulsion breaking techniques are heating, phase dilution, and high stress (Devulapalli and Jones 1990).

Applications of the Emulsion Liquid Membrane Procedure Metal Extraction Removal of heavy and toxic metals is of crucial importance in environmental protection. ELM technology has been proposed to improve extraction procedures and reduce the consumption of solvents and carrier requirements in comparison with other conventional solvent extraction techniques. Numerous reports have been presented in the field of ELM application for separation of metals such as gold, silver, lanthanides, and rare earth elements. Table 3 provides several recent reports about the application of ELM in inorganic pollutant separation. The selective extraction of dysprosium by the ELM method has also been reported in which a new extractant (Cyanex 572) was successfully used as a carrier (Raji et al. 2018). In another study, more than 99% of Co and Ni were extracted from a chloride solution by use of Cyanex 301 (0.1 M) in a very short time (only 2 min) (Hachemaoui and Belhamel 2017). In order to resolve the problem of cyanide wastewater, cyanide removal by use of a W/O emulsion in the form of an emulsion liquid membrane was addressed by Xue et al. (2016). Recovery of rare earth elements during wet processing of phosphoric acid also has a great significance. When di(2-ethylhexyl) phosphate (D2EHPA) is used as the carrier in the ELM method, high selectivity was seen, but a satisfying extraction rate was not achieved. In this regard, Zhang et al. proposed a new ELM method in which aniline was employed as a carrier for the extraction of rare earth elements from the feed (Zhang et al. 2016). In another study, guanidine extractant (LIX 7950) was proposed as a carrier in Cu and cyanide removal from cyanide-contaminated wastewater (Lu et al. 2015). Bjorkegren et al. also introduced a new variation, through the use of environmentally friendly and nontoxic diluent (like palm oil) for Cr (VI) extraction from aqueous samples (Bjorkegren et al. 2015). Due to toxicity of nanosilver ions toward aquatic organisms, release of nanosilver ions into the environment has raised considerable concern. Logically, therefore, a study on using the ELM method for the extraction of ionized nanosilver from washing water has been conducted (Sulaiman et al. 2014). Mokhtari et al. used an emulsion membrane in which calixarene nano-baskets were used as the carrier and surfactant for the extraction of alkali metals (Mokhtari and Pourabdollah 2013). More applications of the ELM procedure can be found in Table 3 (Kassem et al. 2017; He et al. 2015; Elsayed et al. 2013; Mousavi et al. 2012; Gupta et al. 2011).

Matrix Acidic solutions Chloride solution

Wastewater Phosphoric acid

Cyanide solutions Water Wash water Water

Nitrate medium

Sulfuric acid solution

Phosphoric acid

Water Aqueous solutions

Solute Dysprosium Cobalt, nickel

Cyanide Rare earth

Copper, cyanide Cr (VI) Nano-silver Alkali metals

Paladium

Ce (IV)

Uranium

Arsenic (V) Mercury(II)

D2EHPA/ TOPO Cyanex 921 D2EHPA

Cyanex 471X D2EHPA

LIX 7950 TOMAC Cyanex 302 Calixarene

TOA Aniline

Extractant Cyanex 572 Cyanex 301

Span 80 Span 80

Span 80

Span 80

Span 80

Span 80 Tween 80 Span 80 Calixarene

Span 80 T154

Surfactant Span 80 Span 80

Na2SO4 Thiourea

TDDA

HCl/H2O2

KSCN

KOH NaOH Thiourea (NH4)2CO3

NaOH HCl

Stripping agent HCl HCl

Kerosene Toluene

Sulfonated kerosene Kerosene

Chloroform

Kerosene Sulfonated kerosene Kerosene Palm oil Kerosene Kerosene

Diluent Kerosene Kerosene

Table 3 Recent formulations of emulsion liquid membranes for extraction of metals and organic compounds

Mousavi et al. (2012) Gupta et al. (2011)

Elsayed et al. (2013)

He et al. (2015)

Lu et al. (2015) Bjorkegren et al. (2015) Sulaiman et al. (2014) Mokhtari and Pourabdollah (2013) Kassem et al. (2017)

References Raji et al. (2018) Hachemaoui and Belhamel (2017) Xue et al. (2016) Zhang et al. (2016)

1224 H. Tabani et al.

Span 80 Span 80 Span 80

– – Aliquat 336

Water

Aqueous solution Fermentation broths

Crystal violet, methylene blue Congo red Gibberellic acid

Span 80

4-Nitrophenol

–

Water

Rhodamine 6G 4-Chlorophenol Propylparaben

Butadiene styrene rubber Span 80

Span 80 Span 80

Span 80 Span 80

Span 80 LK-80 Span 80

Wastewater Water Aqueous solution

Acetaminophen

Acetic acid

Aliquat 336 D2EHPA

TOA TBAB

Aliquat 336 D2EHPA – TOPO

Pulping wastewater Sodium chloride solutions Alangium platanifolium root Water

Lignin Phenylalanine

Anabasine

Aqueous solution Wastewater

Ethyl paraben Diclofenac

Na2CO3 KCl

NaOH

Na2CO3

H2SO4 NaOH Na2CO3

KCl

H2SO4

NaHCO3 HCl

Na2CO3 NaOH

n-Hexane n-Heptane

Oil

n-Hexane

Kerosene Kerosene n-Hexane

n-Hexane

Kerosene

Kerosene Kerosene

n-Heptane Dichloromethane

DAas and Hamdaoui (2010) Berrios et al. (2010)

Chaouchi and Hamdaoui (2014a) Othman et al. (2013) Kargari (2013) Chaouchi and Hamdaoui (2015) Chaouchi and Hamdaoui (2014b) Agarwal et al. (2010)

Guo et al. (2015)

Kohli et al. (2018) Seifollahi and RahbarKelishami (2017) Ooi et al. (2016) Fang et al. (2016)

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Extraction of Organic Compounds Although the emulsion membrane method has been mostly used to remove and extract metals, application of this method in organic compound removal has also been investigated, and there exist several reports in this area. Recently, the ELM method was used for ethyl paraben (EP) removal from aqueous solution; this method involved the use of trioctylamine (TOA), n-heptanes, Span 80, and Na2CO3 as the carrier, diluent, surfactant, and stripping agent, respectively (Kohli et al. 2018). Through the use of a sustainable emulsion in optimal conditions, almost 90% of the EP could be recovered (Kohli et al. 2018). ELM can also be employed for treatment of pharmaceutical wastewater due to its high extraction efficiency in a short time. In this context, Diclofenac was separated from wastewater samples using the ELM method (Seifollahi and Rahbar-Kelishami 2017). Under optimal conditions, the maximum extraction from DCF was 99.65%. In another study, up to 95% of lignin was extracted from pastry wastewater (Ooi et al. 2016). Hence, ELM is an alternative technology for the separation of some organic materials from their solutions. Recently, phenylalanine was extracted from sodium chloride solutions using the ELM method as a selected model system in which sulfonated kerosene was used as a diluent (Fang et al. 2016). The flow chart of this extraction process is shown in Fig. 5. An experimental study was also conducted on the extraction of DL-anabasine from Alangium platanifolium root (APR) using the ELM system (Guo et al. 2015). In this study, an emulsion liquid membrane was prepared using an emulsifier (styrene-butadiene emulsifier) as an alternative to Span 80, and acetic acid and kerosene were applied as carriers and diluents, respectively. The results indicated that the extracted DL-anabasine was three times greater than that extracted by liquid-liquid extraction. Acetaminophen extraction from aqueous solution with the ELM method in which Aliquat 336 was employed as a carrier has also been reported (Chaouchi and Hamdaoui 2014a). Under favorable conditions, the practical extraction of all acetaminophen molecules from the external feed solution was shown to be possible. In another study, the extraction of rhodamine 6G (R6G) from liquid wastewater was addressed (Othman et al. 2013). Chlorophenols (CP) are hazardous compounds

Internal Phase

Membrane Phase

External Phase

Extract

Membrane Phase

Emulsion

Extraction

Raffinate Stratification

Concentrate Demulsification

Fig. 5 Flow chart for the extraction process of l-phenylalanine by ELM. (Reproduced with permission from Elsevier (Fang et al. 2016))

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produced during chlorination of phenol in water purification systems. These materials are considered as harmful substances in wastewater and exhibit toxicity even at very low concentrations. Therefore, recovery and recycling methods for treatment of CP-containing wastewater (rather than destructive methods such as chemical oxidation and garbage incineration) should be developed. In a recent study, 4-CP extraction from aqueous solutions was investigated (Kargari 2013). More applications of the ELM method for the extraction of organic compounds can be found in Table 3 (Chaouchi and Hamdaoui 2014b, 2015; Agarwal et al. 2010; DAas and Hamdaoui 2010; Berrios et al. 2010).

Polymer Inclusion Membrane The concept of polymer-based liquid membrane emerged about 50 years ago, but such membranes have been recently named polymer inclusion membranes (PIMs) (Nghiem et al. 2006). As PIM-based separation is a better alternative to solvent extraction, the popularity of these materials is exponentially increasing (Almeida et al. 2012). PIMs are a kind of liquid membrane consisting of a liquid phase and a base polymer. The base polymer (e.g., cellulose tri-acetate (CTA)) which is usually selected based on the extractant or specific application (Pereira et al. 2009) forms the membrane skeleton and provides mechanical strength. The liquid phase contains a carrier or extractant which is responsible for extracting the species through formation of an ion pair. PIMs are usually prepared by dissolving all components of the membrane in a small volume of a volatile solvent (e.g., Tetrahydrofuran) and pouring the solution on a special surface depending on the application. PIMs can be customized not only physically, but also chemically, since the chemical composition of the membrane can be well prepared for a particular analyte through selecting the proper extractants, plasticizers, modifiers, and polymers (Pereira et al. 2009). Similar to the other liquid membranes (supporting liquid membrane (SLMs)) (Chaouchi and Hamdaoui 2015), PIMs allow for simultaneous extraction and back-extraction on both sides of the membrane and facilitate the selective transfer of the target analyte along the membrane (i.e., from the donor to acceptor solution); therefore, the species would be extracted or preconcentrated. As no volatile and toxic organic solvents are used in this process, these membranes are environmentally friendly and hence separations with use of expensive reagents can be justified. In contrast to SLMs, in which the liquid phase is protected within pores of membranes, PIMs incorporate the membrane liquid phase in the composite chains of the base polymer, thus reducing its tendency to leach adjacent aquatic phases. Therefore, PIMs are substantially more stable and stronger than SLMs (Almeida et al. 2012). All these PIM advantages have attracted a lot of attention in chemical analysis as they can be helpful in improving the sensitivity and selectivity. Figure 6 shows how PIMs can be applied in each of these analytical operations (Almeida et al. 2017). Hence, recent applications of this method in purification of metals and organic pollutants will be discussed in the continuing sections.

Sample solution

Flow out

Receiving solution

Sample solution

Hollow PIM

Receiving solution

PIM

Sample solution

Flow-through approach

Pre-concentration

Sample solution

Transport cell

Separation

SAMPLE PRE-TREATMENT

Sample solution (e.g. creek)

PIM

Receiving solution

PASSIVE SAMPLING

Fig. 6 General schematic representation of PIM applications in chemical analysis. (Reproduced with permission from Elsevier (Almeida et al. 2017))

Sample solution

PIM

Flow in

Optical fibre

Light detector

Potentiometer

Optical fibre

Optode

Reference electrode

ISE

ISE body

PIM Receiving solution

SENSING

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Applications of Polymer Inclusion Membrane Procedure Metal Extraction PIMs have been used for extraction of various metals from different and complicated solutions. Recently, a study addressed the use of a PIM based on poly (vinylidene fluoride-co-hexa-fluoropropylene) in which Cyphos ® IL 101 (trihexyl (tetradecyl) phosphonium chloride) was used as carrier and NPOE was employed as plasticizing agent (in weight ratio of 55/35/10) for online extraction of Vanadium (V) (Yaftian et al. 2018a). In this study, xylenol orange was employed as a colorimetric reagent. It should be noted that the morphology and mechanical properties of the PIM could also affect the extraction procedure. This was first investigated by Witt et al. for the extraction of Zn ions from aqueous solutions (Witt et al. 2018). The possibility of selective separation of lanthanide ions from sulfuric acid solutions was first investigated through use of a PIM including 45% wt. D2EHPA and 55% wt. PVC (Croft et al. 2018). It was revealed that complete back-extraction of the same ions from PIM is achievable by sulfuric acid-containing solution in concentrations of 7.0, 1.0, and 0.3 M. In a review article, the increasing interest in PIMs was highlighted in analytical chemistry (Almeida et al. 2017). This review article presented an overall view about the analytical chemistry application of PIMs and revealed their adaptability for resolving challenging chemical analysis problems. Finding new carriers is an interesting field in PIM applications. A study has also comprehensively described the relationship between transport performance of Cr(VI) through ionic PIMs based on PVDF-co-HP (IPIM) and the length of alkyl chain of ionic liquids based on symmetrical imidazolium bromide at room temperature as a new carrier (Turgut et al. 2017). Aza [18]crown-6 and β-diketone derivatives are also among the other new carriers which have been employed for extraction of silver and zinc ions, respectively (Kolodziejska et al. 2017; Witt et al. 2016). Nowadays, graphene oxide plays a crucial role in extraction techniques. In a unique study, Cr(VI) removal by graphene oxide-based PIMs from Cr plating water was investigated (Kaya et al. 2016). A new approach has been created for determination of Al3+ in aqueous samples through use of optode membrane produced by the addition of aluminum-selecting reactants to PVC (Suah et al. 2015). Addition of Triton X-100 in the membrane phase is useful in improving Al3 + ions absorption from the liquid phase. Therefore, it can increase the absorption intensity of optode (Suah et al. 2015). Moreover, a paper-based sensor was proposed on the basis of PIMs which can be used for the determination of Cu (II) in water (Jayawardane et al. 2013). In this device, PIM contained PVC, di (2-ethylhexyl) phosphoric acid, dioctyl phthalate, and PAN, and the intensity of Pan-Cu2+ complex color was determined by a flatbed scanner. Passive diffusion is the major driving force in PIM-based extraction. However, there are some reports on the application of an electric field in PIMs, to improve extraction kinetics (Chaudhury et al. 2016; Mamat and See 2015; Schmidt-Marzinkowski et al. 2013; See et al. 2013; See and Hauser 2011). For example, the applicability of cellulose acetate membranes for the extraction of inorganic anions by an electric

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field has been investigated. Extraction of highly lipophilic ions (such as perchlorate) is also possible with membranes without the use of any ionic carrier. However, ions with lower lipophilic properties need a carrier (Schmidt-Marzinkowski et al. 2013). Many PIM applications can be found in Table 4 (Yaftian et al. 2018b; Arslan et al. 2017; Kagaya et al. 2017; Baczynska et al. 2016; Gaytan et al. 2014; Zawierucha et al. 2013).

Organic Compound Extraction PIMs have been usually employed for the removal or preconcentration of inorganic compounds. However, there are some particular reports about the use of PIMs for organic compound extraction. Malachite green is an organic compound which has been controversially used as a dye for coloring hair and as an antimicrobial agent in aquaculture as well. In a recent study, this compound was extracted from real and synthetic wastewater solutions with a PIM (Ling and Suah 2017). The PIM contained poly(vinyl)chloride (PVC) as a polymer, bis-(2-ethylhexyl) phosphate (B2EHP) as the extractant and dioctyl phthalate (DOP) as a plasticizing agent (Ling and Suah 2017). After optimizing the PIM composition, the average extraction efficiency was more than 98%. In a similar study, some acid dyes were extracted from environmental samples by PIMs consisting of cellulose triacetate as the base polymer, NPOE as plasticizer, and tricaprylmethylammonium chloride (Aliquat 336) as carrier (Salima et al. 2016). As a highly consumed antibiotic, sulfamethoxazole (SMX) has been commonly found in wastewater treatment plants and peripheral waters. It has adverse effects on living organisms and therefore needs to be removed from aqueous systems. A study focused on the development of PIM-based passive samplers containing Aliquat 336 as an extractant and carrier for tracking SMX in aqueous systems (Rodriguez et al. 2016). Also, other antibiotics have been extracted from wastewater samples based on PIMs (Rodriguez et al. 2015; PerezSilva et al. 2012). PIMs have several important properties. They are nonporous membranes and dry substances which can be prepared prior to application and stored without significant deterioration for extraction operations. In a study, an in-line coupling of the PIM-based microextraction method was shown for the first time with a separation method for clinical purposes, and formate (the main metabolite in methanol toxicity) was determined in undiluted human plasma and whole blood by capillary zone electrophoresis (CZE) (Pantuckova et al. 2015). Transport of organic acids, such as oxalic acid and lactic acid, was carried out through a PIM. As a polymer, cellulose tri-acetate (CTA) was used and 1-alkylimidazole (IMI-n, n = 10, 11, 12, 14) was employed as the carrier (Przewozna et al. 2014). As a pollutant, phenol can also be removed from water samples by the PIM method. The first study on the application of a PIM for phenol removal and transfer from aqueous solutions and contaminated water was reported by Perez-Silva et al. (2013). The results indicated that this process was highly discriminative for phenol even in heavily polluted waters. Other applications of PIMs for the extraction of organic compounds can be found in Table 4.

Water Chrome plating water Water Wastewaters

Water

Sulfate solutions

Water Acidic solution Chloride solutions HCl media Landfill leachate

Zn(II) Cr(VI)

Inorganic anions

V(V)

Zn(II) Cr(VI) Zn(II), Fe(II), Fe (III) Au(III) Pb(II), Cd(II), Zn (II)

Al3+ Cu(II)

Wastewater

Ag (I), Cu(II)

NPOE – NPOE NPOE NPOE

– Resorcinarene

CTA CTA

NPOE

Cyphos® IL 101 NaDDTC Aliquat 336 Cyphos® IL 101

NPOE

DOP DOP

ADO NPOE

NPOE

TEHP

DAO –

Aliquat 336

Triton X-100 D2EHPA

N-(diethylthiophosphoryl)-aza[18]crown-6/ N-(diethyloxophosphoryl)-aza (Payan et al. 2011) crown-6 β-Diketone derivatives Calix [4]arene derivative

RTILs

ACAC/D2EHPA D2EHPA

PVDFHFP CTA PVC CTA

CTA

PVC PVC

PVC CTA

PVDFHFP CTA

Water

Zn(II) La(III), Ga(III), Yb(III) Cr(VI)

Plasticizer NPOE

(continued)

Gaytan et al. (2014) Zawierucha et al. (2013)

Arslan et al. (2017) Kagaya et al. (2017) Baczynska et al. (2016)

Suah et al. (2015) Jayawardane et al. (2013) SchmidtMarzinkowski et al. (2013) See and Hauser (2011)

Kolodziejska et al. (2017) Witt et al. (2016) Kaya et al. (2016)

Turgut et al. (2017)

Witt et al. (2018) Croft et al. (2018)

References Yaftian et al. (2018a)

Analyte V(V)

Carrier Cyphos® IL 101

Table 4 Recent formulations of polymer inclusion membranes for extraction of metals and organic compounds

Polymer PVDFHFP PVC PVC

Recent Advances in Membrane Extraction Techniques for Environmental. . .

Matrix Dietary supplements Water H2SO4

50 1231

Citric acid

Phenol

Organic acids Phenol

Sulfamethoxazole Formate

Analyte Malachite Green Acid dye

Table 4 (continued)

Aqueous solution Water

Matrix Wastewater Aqueous solution Aquatic systems Human serum/ whole blood Water Polluted water

CTA

PVC

CTA CTA

CTA CTA

Polymer PVC CTA

1-Alkylimidazoles

N503

1-Alkiloimidazole Cyanex 923

Aliquat 336 Aliquat 336

Carrier B2EHP Aliquat 336

Przewozna et al. (2014) Perez-Silva et al. (2013) Meng et al. (2015) Gajewski et al. (2014)

– NPOE

–

N503

Rodriguez et al. (2016) Rodriguez et al. (2015)

References Ling and Suah (2017) Salima et al. (2016)

NPOE NPOE

Plasticizer DOP NPOE

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Environmental Toxicology and Air Pollution: A Comparative Analysis of Different Methods and Studies

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Gustavo Marques da Costa, Larissa Meincke, Darlan Daniel Alves, Ane Katiussa Siqueira Frohlich, Sandra Manoela Dias Macedo, and Daniela Montanari Migliavacca Osório

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Environmental Health . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Air Pollutants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Impacts of Air Pollutants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Environmental Monitoring: Air Pollution . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Methodologies Related to Environmental Toxicology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Active and Passive Biomonitoring . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Comparative Analysis of Different Studies and Parameters: Environmental Toxicology and Atmospheric Pollution . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Final Considerations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cross-References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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G. M. da Costa (*) Programa de Pós-graduação em Qualidade Ambiental, Universidade Feevale, Novo Hamburgo, RS, Brazil Universidade Feevale, Novo Hamburgo, Brazil e-mail: [email protected] L. Meincke Universidade Feevale, Novo Hamburgo, Brazil e-mail: [email protected] D. D. Alves · D. M. M. Osório Programa de Pós-graduação em Qualidade Ambiental, Universidade Feevale, Novo Hamburgo, RS, Brazil e-mail: [email protected]; [email protected] A. K. S. Frohlich UFMG, Belo Horizonte, Brazil e-mail: [email protected] S. M. D. Macedo Departamento de Farmacociências, UFCSPA, Porto Alegre, Brazil e-mail: [email protected]; [email protected] © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_170

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Abstract

The degradation of air quality is increasing in several regions of the world, making air improper or harmful to human health and the environment. The environmental toxicology aims to establish the safety limits of a chemical agent and assessment of environmental exposure, anticipating the consequences of damage to human health. The objective of this chapter will be to present concepts and methods related to environmental toxicology and atmospheric pollution, as well as a mapping of parameters and studies carried out involving the theme in question. In the atmosphere, the effects of pollutants include, in addition to damage to vegetation, soil, and materials, the increase of particles, thus reducing visibility and inhibiting solar radiation received; the increase of the concentrations of gaseous pollutants that absorb longwave radiation and increase the surface temperatures; and also the alteration of precipitation, reducing the amount of rain due to the increase of particles that function as a cloud condensation nucleus. In human health, exposure to air pollutants can affect the respiratory system, cardiovascular system, and nervous system, leading to increased mortality and morbidity. The importance of the use of environmental parameters is associated with its use as an instrument for planning and managing urban spaces, serving to make better use of natural resources and also as a preventive measure against environmental degradation. Keywords

Air quality · Human health and the environment · Pollutants · Urban spaces

Introduction The degradation of air quality is increasing in several regions of the world, making air improper or harmful to human health and the environment. Changes in air quality are more relevant in large urban centers or impact areas of industrial activities of high polluting potential, such as in the areas of mining and thermoelectric processes. In addition to industrial activities, vehicle emissions represent a significant percentage of the degradation of air quality in several regions. In this context, vehicle traffic can account for more than 50% of particulate emissions emitted into the atmosphere. The World Health Organization (WHO 2013) recommends that current limits on the emission of pollutants should be revised because even under current legislation, studies indicate that concentrations affect human health. Currently, air pollution can be considered as the major environmental risk factor and greatly responsible for affecting human health, leading to increased mortality and morbidity. The WHO warns that more than 90% of the world population breathes contaminated air. Epidemiological evidence about the health effects of air pollution is growing and evolving rapidly. Air pollution was responsible for the deaths of 7 million people worldwide and 14,100 new cancers in 2012, doubling compared to the previous 4 years (WHO 2014). Of the approximately 3 million deaths that occur annually as a result of air pollution, 94% occur as a

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result of noncommunicable diseases – mainly cardiovascular diseases, strokes, chronic obstructive pulmonary diseases, and lung cancers – besides also increasing the risk of acute respiratory infections (WHO 2016). The regions of the low- and middle-income are the main affected ones, reflecting 97% of these cities that do not comply with WHO guidelines. The serious environmental problem is reiterated by the study conducted in the last 2 years, and in 103 countries, covering more than 3000 cities, air pollution levels doubled (WHO 2016). The growing environmental and public health concern related to the particulate matter, particularly inhalable particulate matter, is associated with damage to human health, as it has been classified by the International Agency for Research on Cancer as a human carcinogen (Group I) (Bocchi et al. 2016). Thus, the concern to evaluate the chemical composition of the particulate material, mainly the inhalable particles with the particle size of less than 10 μm, is associated with an important adsorption characteristic of chemical substances on its surface and that can be transported to the lower respiratory tract, the alveoli. These particles can also affect climate and vegetation and reduce visibility (Basha et al. 2010). The toxicity of these particles is not only associated with the increase in mass but also its chemical composition and shape (Moreno et al. 2006). The atmospheric particles comprise a complex mixture of various elements and chemical compounds such as sulfates, nitrates, ammonia, organic compounds (PAHs, nPAHs, and others), marine salts (NaCl), soil elements (Al, Si, Ti, Ca, Fe), and metallic elements (Pb, Zn, Ni, Cu, V, Cr, Cd, and Hg, among others). Exposure to air pollutants can have severe consequences on the metabolism of individuals and their populations, ultimately leading to loss of viability. Considering that atmospheric contaminants tend to be precipitated by the rains, and through their percolation, they present themselves as a potential source of diffuse pollution. Therefore, the assessment of the environmental risks associated with its launch is essential, and the toxicity bioassays constitute an important tool for environmental monitoring, allowing the evaluation of complex mixtures to which the populations may be exposed (Alves et al. 2016). Bocchi et al. (2016) evaluated mutagenicity (Salmonella test) and genotoxicity (comet and micronucleus test in cell line A549) in the fine particulate matter (MP2.5 and MP1.0), also quantifying HPAs. The results showed a greater presence of mutagenicity in samples of particulate matter collected in winter and fall. However, the results with the comet test did not present seasonality. Thus, the use of different mutagenicity and genotoxicity assays applied to the particulate matter proved to be relevant. Palacio et al. (2016) evaluated the soluble fraction of total suspended particles and particulate matter (MP10 μm) in different samples collected in the state of São Paulo, Brazil. The results of the mutagenicity tests using the Salmonella assay showed positive results for all the samples evaluated. In this context, the objective of this chapter will be to present concepts and methods related to environmental toxicology and atmospheric pollution, as well as a mapping of parameters and studies carried out involving the theme in question. The presented methods can be applied to the elaboration of environmental diagnoses in regions with different magnitudes of environmental impacts.

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Environmental Health Environmental health is conceptualized as the relationship between the environment and the health standard of a population. This relationship incorporates all the elements and factors that potentially affect health. Thus, new theoretical approaches to the relation production/environment/health are studied, aiming at finding explanations about how this relationship occurs, pointing out and proposing interventions based on preventive actions to reduce risk, and having as a risk the likelihood of an adverse effect occurring during a given exposure time. While environmental toxicology aims to establish the safety limits of a chemical agent and assessment of environmental exposure, anticipating the consequences of damage to human health (Brilhante and Caldas 1999), the potential risk in environmental toxicology deals with the study of the likelihood of sources harmful to health and the environment, capable of causing damage, disease, or death to living beings when in concentrations higher than those they can assimilate under normal conditions, that is, to absorb, to distribute, to metabolize, and to eliminate the organism. Therefore, it is defined as: • Risk: the measured or estimated probability of damage, illness, or death caused by a chemical agent in an individual exposed to it. • Danger: qualitative term that expresses the agent’s harmful potential for health and/or the environment. • Risk assessment: the first step in triggering decision-making processes comes from knowing the cause-effect relationship and possible damages caused by exposure to a particular chemical agent. The stages of risk assessment are: – Danger identification: this is the identification of the dangerous agent in its essence, its effects, the exposure conditions, and the target population. – Exposure assessment: refers to the quantification of the concentration of the harmful agent in an environment, for an individual or group. – Risk estimation: relates the quantification of the dose-response or dose-effect relationship to a given environmental agent, demonstrating the likelihood and nature of its effects on health and the environment. – Exposure or dose: it is the quantitative definition of the concentration of chemistry that reached (external dose) the individual or that which was absorbed (internal dose) by him. – Risk characterization: this is the meeting of the previous stages that, in possession of all available data on the subject, characterizes the specific use or occurrence of damage, illness, or death caused by exposure to a particular concentration of a chemical agent. – Management or risk management: thus conceived, it refers to the comparison of calculated risk or public health impacts, environmental exposure to the agent, as well as the possible contribution of social and economic factors that also include the benefits associated with these factors. Ultimately, in this process, it can be established that under the proposed conditions, the risk may be acceptable (Brilhante and Caldas 1999; Amaral e Silva 2004).

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Air Pollutants The rapid growth of the population, as well as the processes of industrialization and urbanization associated with the high consumption of energy and the delay in measures to control atmospheric emissions, especially in emerging countries, has led to an increase in pollutant emissions and, consequently, serious problems of air pollution (Migliavacca et al. 2012; Rodriguez et al. 2015). The main anthropogenic sources of air pollutants are industries (production of both raw materials and other products), automotive vehicles (gasoline, diesel, alcohol), thermoelectric plants, biomass, and fossil fuel burning. These emitting sources may be classified, by their nature, into mobile and fixed sources. Mobile sources are those that are not located in a fixed place, being able to move (automobiles), and fixed sources are located in a specific, fixed point (chimney). Air pollutants can be classified according to their origin (primary and secondary), composition (organic and inorganic), and physical state (particulate matter, gases/vapors). Primary pollutants are those emitted directly by sources of emission into the environment, for example, sulfur dioxide, nitrogen oxides, carbon monoxide, volatile organic compounds, and particulates. Secondary pollutants are those formed by chemical reactions (hydrolysis, oxidation, or photochemical reactions). Among the primary pollutants and natural components present in the lower atmosphere and fractions of solar radiation, for example, emitted sulfur dioxide reacts with oxygen from the air to form sulfur trioxide, which reacts with water vapor to form sulfuric acid, one of the components of acid precipitation. The sources of air pollutants can be originated through natural or anthropogenic events. Natural sources can be volcanic eruptions with the emission of ashes and gases, decomposition of animals and plants, resuspension of soil dust by the winds, formation of methane gas in swamps, marine aerosols, plant pollen, and natural forest fires, among others. Air pollution can be classified into levels: local, urban, regional, transboundary, and global. At a local level (5 km radius), pollution is characterized by high concentrations of specific pollutants that may originate from automobiles or industrial activities in that region. The lower the launch height of a source, the greater the potential impact for a given release. At the urban level, three main types of pollution are characterized: high average of annual concentrations of various pollutants (from industrial and automobile activities); winter-type smog, when low temperatures and mists in the year are observed (from industries); and summer smog, or photochemical smog (from motor vehicles), favored by sunlight and high temperatures. However, regional air pollution (50–1000 km) can be attributed to the transport and dispersion of urban pollutants in large areas and the transport and transformation of primary pollutants to secondary levels at a regional level. Associated with these mechanisms are problems such as the formation of tropospheric ozone and acid rain. Transboundary air pollution is related to the transfer of air pollution from one country to another, and problems that occur at a regional level may also occur between adjacent countries. Global pollution is related to two known problems which are the greenhouse effect and the decrease of stratospheric ozone. It is derived

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from anthropogenic actions contributing to the other levels mentioned above, and the accumulation of pollution from these levels can contribute to the increase of pollution at the global level. Thus, high concentrations of toxic gases and other pollutants emitted into the atmosphere, mainly through anthropogenic sources, have been responsible for causing adverse effects on human health, climate, and the environment (Seinfeld and Pandis 2006; Li and Zhang 2014). Recent greenhouse effect gas emissions are the largest in history, making them the main causes of global warming. In addition, since each part of the planet is interconnected through atmospheric transport, emissions from one continent can reach the other, causing impact on the climate on a global scale and not just locally (Seinfeld and Pandis 2006). In order to understand the harmful effects of atmospheric pollutants on humans and the environment, its constituent is studied, the particulate matter (PM), which is composed of a set of adsorbed pollutants composed of dust, smoke, and all kinds of solid and liquid material that remain suspended in the atmosphere because of its small size. Particulate matter may be classified as: (a) Total suspended particles (PTS), which can be defined in a simplified manner as those whose aerodynamic diameter is less than 50 μm. Some of these particles are inhaled and can cause health problems and can adversely affect the quality of life of the population, interfering with environmental conditions and harming normal community activities. (b) Inhalable particles (MP10) are those whose aerodynamic diameter is less than 10 μm. Inhalable particles can also be classified as fine inhalable particles – MP2.5 (90% degradation

Photocatalysis

Fouad and Mohamed (2011)

Acidic condition: Chloridazon> metribuzin Alkaline condition: Metribuzin > Chloridazon 99%

Photocatalysis

Khan et al. (2012)

Adsorption

Dehaghi et al. (2014)

Photocatalysis

Mir et al. (2013)

Photocatalysis

Khan et al. (2010)

Photocatalysis (UV)

Zabar et al. (2012)

Photocatalysis (UV)

Sharma et al. (2009)

Ultrasonication

Lopes et al. (2008)

TiO2 P25 is more efficient photocatalyst than UV100 and PC500 More efficiently degraded in alkaline pH and all electron acceptors enhanced the degradation rate Mineralized within 2 h with k = 0.035, 0.019 and 0.021 min1, respectively At neutral pH, degradation within 90, 240, and 120 min, respectively Degradation >90% after a reaction time of 30 min via initial NO2 – > NH2 reduction

(continued)

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Table 8 (continued) Pesticide(mgL1) Thiamethoxam, (20 mL; 50 mg L1) Imidacloprid (20 mL; 35 mg L1) Atrazine (2 g)

Atrazine

Bulk/ nanoparticles Fe /Fe3O4 (0.06 mol L1) With/or H2O2 Ag-chitosan (1 L of 1 mg L1) Carbon nanotubes

Brief summary > 90% in 30 min, highly efficient in acidic conditions

Mechanism Redox degradation

Reference De Urzedo et al. (2009)

98% degradation in 65 min

Adsorption

>90 degradation

Adsorption

Saifuddin et al. (2011) Yan et al. (2008)

Fig. 7 Mechanism for the working of zero-valent iron (ZVI) under different conditions

Nanocomposite of Fe3O4 supported on polystyrene degraded aldrin, endrin and lindane via adsorption mechanism (Jing et al. 2013) while combination of C/ZnO/ CdS photocatalytically degraded 98.0% of 4-Chloro phenol (Lavand and Malghe 2015). Scattered information is also available in literature on the use of Cu and supported on reduced graphene for the degradation of endosulfan (Mitra and Varshney 2013) and lindane (Gupta et al. 2015), respectively.

Degradation of OP Pesticides After OCs, the pesticides having great environmental concern are OPs because of their abundant use, toxicity and residue detected in environment. The complete degradation of organic pollutants is not possible by conventional approaches such as digestion (activated sludge, anaerobic etc.) and physicochemical treatment (Galindo et al. 2001; Kuo and Ho 2001; Tang and An 1995). Sud and Kaur (2012) reviewed the importance of heterogeneous photocatalytic degradation of several OP pesticides. These photocatalysts do not operate at room temperature without aid of irradiation system. Therefore, in addition to photocatalysis,

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Fig. 8 Rapid reactive adsorption of parathion methyl on MnO2 (Stastny et al. 2015)

rapid reactive adsorption were highly proposed for remediation of OPs in wastewater. Like OCs, TiO2 as such, modified, doped and composites were investigated along with ZVI and other noble metals for degradation of OPs such as methamidophos malathion Parathion etc. by heterogeneous photocatalysis (Doong and Chang 1997; Xu et al. 2002; Moctezumaa et al. 2007; Zhang et al. 2006; Kralj et al. 2007). Under UV exposure and aqueous medium, TiO2 could complete mineralized (99% photodegradation) acephate (Han et al. 2009) while terbufos could converted into eight intermediates within 90 min (Fig. 12) (Wu et al. 2009). These findings have showed that OP’s degrade rapidly and non-selectively with TiO2. Also, the oxidation strength of the UV/TiO2 photocatalysis is enhanced by the addition of H2O2 than UV/TiO2 photocatalysis (Fig. 6) in degradation

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Fig. 9 Proposed degradation pathway of lindane using N-doped TiO2 (Senthilnathan and Philip 2010) and zero-valent iron (Elliott et al. 2009)

Fig. 10 Role of magnetite in oxidative degradation of pesticides (Rodriguez et al. 2011)

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Fig. 11 Tentative degradation pathway of DDT using Ni/Fe bimetallic nanoparticles (Tian et al. 2009)

of pesticides chlorpyrifos, cypermethrin and chlorothalonil in aqueous solution (Fig. 13). In case of mixed photocatalyst (ZnO/TiO2), the maximum degradation of monocrotophos and quinalphos has been observed when ZnO and TiO2 were in the ratio of 7:3 and 8:2 respectively (Kaur et al. 2013). The degradation efficiency with synthesized heterostructured nanophotocatalyst was found to be comparable with TiO2. Solar photocatalytic activity of WO3/TiO2 was evaluated using malathion as a model contaminant (Fig. 14). Best results were obtained using 2% WO3/TiO2(complete degradation) after 2 h, whereas, 5% WO3/TiO2 and bare TiO2 achieved 28% and 47% mineralization, respectively (Ramos-Delgadoa et al. 2013). Degradation of malathion increased by 172% when modified TiO2 (Au-Pd-TiO2 nanotube film) was used (Fig. 13) (Yu et al. 2010).

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Fig. 12 Proposed photocatalytic degradation pathway of terbufos (Wu et al. 2009)

The feasibility of nanocomposite such as TiO2 SiO2 beads, IO3 doped TiO2, La doped ZnO, TiO2 supported on Zeolite and La-doped TiO2 were employed for eradication of monocrotophos with dichlorvos (Shifu and Gengyu 2005; Gomez et al. 2015) and phoxim (Dai et al. 2009). Doping of TiO2 latice with metal ions might cause an increase in the formation of Ti3+ ion and more surface defects (which depend on the doping identity and concentration) and facilitate efficient adsorption of oxygen on titania surface. Metal ion work as charge carrier traps and effectively enhances the charge separation of electrons and holes resulting in increase in the quantum yield of surface photo reaction. Since the redox energy state of several metal ions lie within the band gap of TiO2, addition metal ions into TiO2 introduces new energy levels between CB and VB edge and absorption was due to the charge

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Fig. 13 Photodegradation of malathion in presence of TiO2/WO3 (Ramos-Delgadoa et al. 2013) and Au-Pd-TiO2 nanofilm (Yu et al. 2010)

transfer transition between the d electrons of dopant and CB or VB of TiO2 (Ramacharyulu et al. 2014) (Fig. 14). Addition of iron compounds (Fe0, Fe2+) could able to enhance the degradation efficiency of OPs similar to OCs via electrochemical reaction (Joo et al. 2004; Hung and Hoffmann 1998; Zhang 2003). In contaminated soil (1–10 μg g–1), quantitative oxidation of malathion into metabolits O-dimethyl phosphorodithioic (non-toxic to human or any other living beings) was achieved with n-ZVI (33–78 nm) within 8 min in ambient environment (Singhal et al. 2012). Degradation kinetics was surface area dependent clearly pointed out that higher specific surface area of iron particles is desired for faster degradation. For example, chlorpyrifos, cypermethrin and chlorothalonil was completely and rapidly degraded in 1 min by coating of Fe-granular AC and H2O2(FeGAC/H2O2), via oxidative mechanism (Fig. 15) (Affam et al. 2016). Here, presence of AC increase the adsorbing properties of iron because of its high surface area. AC as an electron donor reduces the ferric ion to ferrous ion and

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Fig. 14 Photodegradation of chlorpyrifos, cypermethrin and chlorothalonil with Fe/GAC (granular activated carbon) (Bach and Semiat 2011; Affam et al. 2015)

Agn+

Agn+

Agn+

Agn+

Agn+

Fig. 15 Degradation of chlorpyrifos (CP) on Ag nanoparticles (Bootharaju and Pradeep 2012)

combination of ferrous/ferric ions and H2O2 are used to generate excess of OH• (Bach and Semiat 2011). Another metals employed for the degradation of various OPs are Ag and Au supported or modified with other materials. Under solar as well as UV, pesticides containing phenol were effectively removed by Ag@TiO2 at pH~ 3 due to the interactions between positively charged TiO2 surface and negatively charged

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phenols and intermediates (Shet and Shetty 2015). Chlorpyrifos (CP) was degraded using Penicilliumpinophilum derived Ag nanoparticles with consistently highest degradation observed in mild acidic condition (pH 6) irrespective of incubation time (Deka and Sinha 2015). Noble metals have also efficiency to degrade OPs at room temperature. For example, CP is shown to decompose to 3,5,6-trichloro-2pyridinol(TCP) and diethyl thiophosphate over Ag and Au NPs supported on neutral alumina (Fig. 15) Bootharaju and Pradeep (2012). Complete degradation of CP occurred within 3 h through the formation of AgNPS surface complex where PO bond cleaves to yield a stable aromatic species, TCP. In addition to these catalysts, some other catalysts were also proved to be effective in the degradation of hazaradous OPs like ZnO and PbO2. ZnO nanoparticles were proved to be a time saving photocatalyst for the degradation of Chlorpyrifos with degradation rate increased several times when photocatalytic reactor was combined with membrane filteration (Khan et al. 2015). Using Bi-doped PbO2 electrodes several pollutants from wastewater were removed successfully, particularly Tebuconazole-TBC into less toxic products than TBC (de Figueredo-Sobrinho et al. 2015). Overall, due to high adsorbing properties and semiconducting nature, nanophotocatalysts (TiO2, Ag, Au, ZnO, PbO2) become efficient adsorbent and generate free radical upon irradiation. This free radical is responsible for the degradation of hazardous OPs pesticides in a short duration of time. Similarly, the chitosan based nano-biocomposites were proved to show the higher removal of organophosphate pesticides (Jaiswal et al. 2012; Dehaghi et al. 2014). Higher removal of dichlorvos by CuO-montmorillonite-chitosan and CuO-montmorillonite gumghatti nano-biocomposites suggested that biopolymers can play major role in the adsorption of pesticide (Sahithya et al. 2015). Nanocomposites ZnO – Zeolite and CuO-Chitosan completely degraded Monocrotophos (Tomasevic et al. 2010) and Malathion (Jaiswal et al. 2012) via Photocatalytic and Adsorption mechanism, respectively. Another most extensively field used for fast destruction of toxic pesticides into less toxic residuals is use of potential reactive sorbents such as nanostructured metal oxides of Mg, Ca, Ti, Mn, Fe, Zn, Al, and Ce (Houskova et al. 2007; Chen et al. 2010; Štengl et al. 2016). MgO, Al2O3 and CaO titania nanotubes, manganese oxide nanotubes remove the agent rapidly owing to their high surface area, strong adsorbability and potential reactivity (Talmage et al. 2007). MnO2 and various oxides were effectively used in 90% decomposition of parathion methyl within 2 h by cleavage of the P–O-aryl bond in the pesticide (Fig. 8). These promising materials are prepared via homogeneous hydrolysis and exhibit a capability to capture and degrade toxic OP compounds via nucleophilic substitution SN2on their surfaces even at ambient temperature i.e., rapid reactive adsorption (Khaleel et al. 1999; Mitchell et al. 2004). They are first time developed by Klabunde, and his research group and used for the destruction of chemical warfare agents and analogues OPs (Klabunde et al. 1996; Khaleel et al. 1999; Lucas and Klabunde 1999). It was found that exhibit a comparatively higher degradation activity. Mixed or doped oxides with unusual structural features and properties are promising materials for preparing reactive sorbents than pure oxides. Ti/Ce mixed oxides or composites are used in

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heterogeneous catalysis (Gao et al. 2010) due to unique redox properties (Rynkowski et al. 2000). Henych et al. (2016) studied there active adsorption of toxic parathion methyl and DMMP on nanostructured Ti/Ce oxides and their composites. Nano-ceria exhibited substantially higher degradation ability compared to nano-titania. The highest degradation efficiency was achieved with the composites with Ti:Ce 2:8 and 1:1 molar ratio. Interestingly, the degradation efficiency of the TiO2/CeO2 composites is moderately higher than that of pure CeO2.The strong interaction of Ti with Ce led to an increase of Ce3+ and formation of Ti4+ states and changed surface area and porosity which may cause improved degradation efficacy for both DMMP and parathion methyl. In addition to metal oxides, the use of metal-organic frameworks (MOFs) and organocatalysts as reactive adsorbent are also in progress. MOFs are crystalline metallo-organic compounds with a well-defined geometric structure consisting of metal ions or clusters coordinated to organic ligands. Some of the Zr-based MOFs and their rare earths analogues were tested for the degradation of organophosphate pesticides. The study of Barba-Bon (2015), who examined a series of different amines, aminoalcohols and glycols as potential organocatalysts for DCNP (diethylcyanophosphate) degradation, may serve as an example. It was found that some of the functionalized sorbents are capable to destroy parathion methyl and convert it to 4-nitrophenol in a similar way as the metal oxide-based reactive sorbents. On contrary to the metal oxide-based reactive sorbents, the degradation proceeds in aqueous solutions, although the rate of degradation is lower.

Degradation of SU Pesticides In spite of a less persistence in soil and water (Zhou et al. 2009), their chronic exposure and long-term toxicity effects were observed via food chain. Chemical hydrolysis and microbial degradation of SU in soil is influenced by pH, moisture content and microbiological activity (Saha and Kulshrestha 2008; Si et al. 2005). The dissipation of two SU herbicides, chlorsulfuron(t1/2 = 6.8–28.4 days)and imazosulfuron(t1/2 = 6.4–14.6 days) was faster in acidic medium(Wang et al. 2010). Saha and Kulshrestha (2008) carried out pH dependent hydrolytic degradation of sulfosulfuron with faster hydrolysis rate in acidic condition (t1/2 = 9.24 days; pH 4.0) than alkaline (t1/2 = 14.14 days; pH 9.2). Under abiotic conditions, the major degradation mechanism of the compound was the breaking of the sulfonylurea bridge yielding corresponding sulfonamide and aminopyrimidine (Saha and Kulshrestha 2008). The ultraviolet (UV) radiation in the sunlight is one of the most powerful forces for pesticide degradation. Photolysis of imazosulfuron was reported inaqueous solution in UV light under aerobic (chemical cleavage) and anaerobic (microbial degradation)conditions (Wang et al. 2010). Diflubenzuron is quickly degraded in the environment mainly by hydrolysis and photodegradation producing major metabolites: 2, 6-diflurobenzamide,

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4-chlorophenylurea, 4-chloroacetanilide, 4-chloroaniline and N-methyl-4chloroaniline, which are classified as mutagens (Rodriguez et al. 1999). Senthilnanthan et al. (2010) carried out photocatalytic degradation of metsulfuronmethyl in aqueous phase under visible light with using TiO2 (40%) and Ru doped TiO2 (80%). Use of nanoparticles for the degradation of SU pesticides is still awaited. No doubt these pesticides have comparatively lower half-lives than the OCs or OPs, but once the threshold limit is exceeded, these can cause severe damage to the environment due to their toxicity. Keeping in view the future aspects, their use should be limited or banned. Presently, to eliminate the persisting SU from the environment, various nanomaterials can be employed with high degrading efficiencies. This opens a research gap for the researchers working in this field.

Degradation of Carbamate Pesticides Because of the variety of different groups present in carbamateinsecticides, the metabolism of these compounds is often complex. Carbaryl, for example, is a relatively simple compound, yet it is metabolized to at least 15 different compounds in mammals through a variety of oxidative and hydrolytic reactions (Leeling and Caisida 1966). Scattered informations on biodegradation on carbamate are moving slightly to use of nanophotocatalyst. A study showed that the urine from carbaryl-treated rabbits contained at least four or five other watersoluble metabolites (glucuronides or sulfates). Tien et al. (2013) used natural river biofilms to degrade the carbamate pesticides methomyl, carbaryl and carbofuran. Certain metabolites of carbamates such as 1-naphthol are found in the urine as 1-naphthyl glucuronide and/or 1-naphthyl sulfate. Sullivan et al. (1972) identified one of these metabolites, 5,6-dihydroxy carbarylglucuronide, in the urine of rats. Doddamani and Ninnekar (2001) isolated Micrococcus species from garden soil and used for the degradation of carabaryl (Fig. 16) into 1-naphthol and methylamine, further metabolized by oxygen uptake. Degradation of aldicarb under UV radiation was carried out by Dixit et al. (2009) using TiO2/polyacrylonitrile nanofiber into lesser toxic third oxidation oxime product (Fig. 17). Photo catalytic oxidation of Methomyl by Fe supported on Zeolite resulted in complete removal (Tomasevic et al. 2010). It is evident from the above the information that the literature available on the degradation of carbamates using nanoparticles is very scarce. Since these are a new class of pesticides, therefore, not much attention has been paid to their degradation using advanced technologies. After one or two decades, these pesticides may cross their threshold limit and ultimately will affect the environment. To avoid this situation, prior efforts must be made. In this context, use of nanotechnology is highly recommended.

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Fig. 16 Degradation pathway of carabaryl by Micrococcus species (Doddamani and Ninnekar)

Fig. 17 Proposed pathway showing third oxidation of aldicarb using TiO2 catalyst (Dixit et al. 2009)

Degradation of Miscellaneous Pesticides Heterogeneous photocatalysis using TiO2 as photocatalyst appears as the most emerging destructive technology for all pesticides including well known OCs and OPs. Complete photodegradation of imidacloprid (90 min), isoproctum (240 min), phosphamidon (120 min) was achieved using 5 wt% TiO2 supported on porous nanosilica (under neutral pH conditions) (Mangalampalli et al. 2009). Mineralization rate was enhanced in the presence of TiO2 containing H2O2 due to the

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efficient generation of hydroxyl radicals. Degussa P25, was found to be more efficient catalyst for the degradation of both herbicides chloridazon and metribuzin as compared with two other commercially available TiO2 powders like Hombikat UV100 and PC500 (Khan et al. 2012). Chloridazon was found to degrade more efficiently under acidic condition, whereas metribuzin degraded faster under alkaline medium. Similar observation were found for photocatalytic degradation of Thiamethoxam. Photocatalysis of Thiamethoxam follows pseudo-first-order kinetics. An increase of the degradation percentage was observed for removal of chloridazone on using Au/ZnO and Au/TiO2 and also a reduction of the radiation rate time, which is due to the presence of the gold which prevents the electron hole recombination and accelerates the photodegradation (Fouad and Mohamed 2012). In addition to TiO2 zero-valent metals in the presence of ultrasonic irradiations (Lopes et al. 2008) and Fe /Fe3O4 composite were found efficient in promoting the degradation of thiamethoxam and imidacloprid in acidic aqueous solution. Dual behavior of Fe /Fe3O4 composite as a reducing or an oxidizing agent was found depending on absence or presence of H2O2, respectively. Fe3O4@Au has turned out to have high efficiency in photodegradation of chloridazon under UV and sunlight in comparison to the Fe3O4 due to the presence of the gold which has the plasmonic phenomena having a great influence in accelerating the oxidation of the pesticides (Fouad and Mohamed 2011). Silver supported on chitosan adsorbed 94.0% of Atrazine (Saifuddin et al. 2011) and demonstrated the high removal capacity of synthesised Ag/chitosan nano-biocomposite. During the removal of pesticide atrazine on CNTs, Yan et al. (2008) inferred that the surface area might not be the only factor that determines the adsorption capacity; other factors such as surface chemistry would also influence the adsorption processes. Carbon nanotubes (CNTs) are most efficient adsorbents because of their unique chemical structure and intriguing physical properties (high surface area, high reactivity) that helps in formation of strong interaction between organic chemicals and CNTs (Long and Yang 2001; Fugetsu et al. 2004; Mauter and Elimelech 2008).

Utilization of Green Synthesized Nanomaterials in Remediation of Environment Nanomaterials generated via green route are cheap. Currently, a lot of researchers worldwide are working on synthesis of nanoparticles using by employing sunlight, or plant-based surfactants or microorganisms and water or combination of those. Application of these nanoparticles in the degradation of harmful pesticides will help in reconstructing the polluted environment. There are some examples highlighting the use of green synthesized nanomaterials in removal of various organic pollutants and provide the feasibility in treatment of pesticides in wastewater. Recently, sunlight assisted green synthesis of various transition metal oxide (ZnO, CuO, Co3O4, NiO and Cr2O3) nanoparticles and their catalytic application in removal of organic colorants were studied (Shanker et al. 2016b). Interestingly, in

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a short period of 180 min, 88% of the dyes was more or less completely degraded using Cr2O3, ZnO followed by >87% (CuO) >86% (NiO) >80% (Co3O4), depending on the sizes of respective nanoparticles. Employing a green route Jassal et al. (2016b) synthesized different potassium metal hexacyanoferrate (KMHCF) nanoparticles using Aegle marmelos as a natural surfactant and utilized them for the degradation of various dyes. 94.72% of the dye was degraded using KCuHCF followed by 91.35% using KNiHCF and 89.28% using KCoHCF nanoparticles (Jassal et al. 2016b). Using Morinda tinctoria leaf extract silver nanoparticles were synthesized and effectively degraded 95% of dye in 72 h under sunlight irradiation (Vanaja et al. 2014). In addition to this, Polygonum Hydropiper was also used as a biogenic source for synthesizing silver nanoparticles with high catalytic efficiency in degrading MB completely within 13 min (Bonnia et al. 2016). Yeast (Saccharomyces cerevisiae) extract was also utilized for synthesizing silver nanoparticles which were effectively applied for degrading ~90% of MB within 6 h (Roy et al. 2015). A high degradation of 98% for Orange II was reported using green synthesized bimetallic Fe/Pd nanoparticles, whereas, only 16% of the dye was degraded using Fe nanoparticles (Luo et al. 2016). Green synthesized copper oxide nanoparticles and ZnO/Graphene Oxide (GO) nanocomposite were also used for decolorization of the dye (Sankar et al. 2014; Lellala et al. 2016)

Conclusions and Future Scope This review covered the encouraging performances of a variety of nanomaterials and nanocomposites as adsorbent, heterogeneous redox and photo-catalyst or reactive sorbents for the effective degradation of pesticides (Rani et al. 2017b; Rani and Shanker 2017b; Jassal et al. 2015a, b, c; Jassal et al. 2016a). Nanomaterials not only adsorb the pesticides, but also degrade them into non-toxic by-products. Being inexpensive, commercially available, non-toxic in nature, and of high efficiency, TiO2 and nZVI have emerged as a promising wastewater treatment technology in the last few years (Shanker et al. 2016a, b, 2017a, b, c; Rani et al. 2017b). Majority of the nanoadsorbents are able to degrade more than 80% of the pesticide and various even achieved complete degradation in a short span of time. Further, the band gap of TiO2 can be altered via the insertion of other metal oxides or electron rich elements. Addition of H2O2 enhanced the oxidative efficiency of nanomaterials because of production of more •OH radical. Nanocrystalline reactive adsorbents such as metals (Mg, Ca, Ti, Mn, Fe, Zn, Al, Ce) by itself or in mixture(e.g., Ti/Ce) and their oxides are found to have potential against OPs at ambient temperature. Further, nanocomposites such as ZnO/clay, Fe/zeolite and TiO2/zeolite and nanobiocomposites like TiO2/chitosan were being used due to their enhanced activity and biocompatibility. OCs and OPs are frequently degraded by nanomaterials but carbamates and SU (future contaminants) have not been explored extensively yet. Nanomaterials mediated studies on different traditional or upcoming pesticides are imperative in order to ban or restrict them. Finally review ends with suggested future scope of

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using green synthesized nanomaterials that would be cheaper and effective catalysts. For sustainable developments, biopolymers-based nano-biocomposites should be developed.

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R. M. S. Radin Mohamed, Adel Al-Gheethi, Muhammad Shabery Sainudin, and M. K. Amir Hashim

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Grey water Characteristics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Grey water Treatment by Filtration Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Ceramic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Size and Surface of Media . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cross-References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

This review described, summarized, evaluated, and clarified the related theories and previous research relevant to bathroom gray water quality and its design treatment system to reduce the contamination of gray water. It provides a theoretical base for research and helps in determining the nature of the research. In addition, this chapter may also serve as a handy guide on the use of ceramic filtration for bathroom gray water. Keywords

Bathroom gray water · Design treatment system · Contamination R. M. S. Radin Mohamed (*) · A. Al-Gheethi (*) · M. K. Amir Hashim Micro-Pollutant Research Centre (MPRC), Department of Water and Environment Engineering, Faculty of Civil and Environmental Engineering, University Tun Hussein Onn Malaysia (UTHM), Parit Raja, Johor, Malaysia e-mail: [email protected]; [email protected] M. S. Sainudin Micro-Pollution Research Centre (MPRC), Department of Water and Environmental Engineering, Faculty of Civil and Environmental Engineering, Universiti Tun Hussein Onn Malaysia, Parit Raja, Johor, Malaysia © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_34

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Introduction The water pollution is a major global problem and requires ongoing evaluation and revision of water resource policy at all the levels. The increasing of water pollution due to the direct discharge of wastewater into the natural water systems is associated with the growth rate of the population as well as the rapid construction development (Mohamed et al. 2014, 2016). Therefore, it can be stated that most of the environmental issues are arising from human, commercial, domestic, or industrial activities. Grey water in the rural regions is one of the major sources of pollution because it contains different chemical substances such as xenobiotic and detergents as well as nutrients which are directly discharged into rivers without any treatment due to the absence of central wastewater treatment plants (Figs. 1 and 2). These wastes are also defined as the wastewater that resulted from the urban activities which include baths, showers, handbasins, washing machines, dishwashers, and kitchen sinks, but not the backwater from the toilets (Jefferson et al. 1999; Al-Gheethi et al. 2016). According to Friedler and Galil (2003), around 60–70% of gray water are transformed from the urban water demand of industrialized countries. However, the real quantities of the gray water are depending on the lifestyle and household activities which are differing from one country to another (Al-Mashaqbeh 2012). Bathroom gray water has low contaminations in comparison with the blackwater especially in terms of organic contents (Mohamed 2011). In some definition, these waste are called the light gray water (Friedler and Hadari 2006); it represents 30% of the total wastewater generated from each house (Lesikar 2005). However, it still contains viruses and pathogenic bacteria from cleaning activities in the bathroom

Fig. 1 Flow of bathroom gray water to the drains (Photo was taken on 12 October 2016)

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Fig. 2 Main drainage conditions (Photo was taken on 12 October 2016)

such as body washing, laundry, and nappy washes. Mohamed et al. (2013) stated that the bathroom gray water has nutrients as phosphorus and nitrogen which is coming from the detergents used in laundries activity. The discharge of untreated bathroom gray water might cause unpleasant odors and increase the insect attraction; the presence of nutrients in the gray water creates ideal surroundings for bacteria growth and thus attracts for insect pests to breed (Mohamed et al. 2014). Filtration method is one of the most efficient technologies used for treating of gray water. The filtration media from natural sources such as sand, clamshell, charcoal, and gravel have been investigated to determine their efficiency in removing pollutants from the wastewater. These materials are more applicable because they are easy to be handling, sustainable with low cost (Siracusa and La Rosa 2006). The combination between different natural materials in the design a filtration system for the treatment of gray water has been reported in literature (Mohamed et al. 2016). The most common natural materials used as filter are fine particles such as ceramic  (fired  clay) which exhibited high efficiency in removing lead (Pb) by 100%, sulfate SO2− 4 by 76%, iron (Fe) by 47%, and fluoride by 75% from the wastewater (Erhuanga et al. 2014). The present chapter has focused on highlighting a sustainable treatment process of bathroom gray water in the village house by using filtration media. Ayer Hitam is one of the towns in Johor and is well known for ceramic production industries. The town is rich in kaolinite which is the main raw material in the ceramic industries. All kinds of ceramics, including garden decorative items, kitchenware, home furnishings, and souvenirs, are available in this town. Figure 3 shows the ceramic waste disposed at the side of the store. The improper dumping of these wastes produces a significant visual impact and environmental degradation (Juan et al. 2010). Therefore, the importance of ceramic waste recycling as filtration system for gray water lies in introducing a new filtration media. Although the bathroom gray water water are less polluted, it also needs to be treated due to the

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Fig. 3 Ceramic wastes from “AS Highway Enterprise Pusat Perkahwinan Kraftangan Bunga”

presence of several various chemical agents such as detergents, solvents, organic acids, and bleaching agents which might be the reason for serious health problems. These pollutants are qualitative pollutants and might be more serious than the quantity pollutants such as nutrients and pathogens.

Grey water Characteristics Grey water is wastewater generated from the households or building activities without fecal contamination. It is also defined as the urban wastewater from bathrooms, handwashing basins, laundry, and kitchen excluding toilet waste (Jefferson et al. 1999; Ottoson and Stenström 2003). Grey water are derived from the residential area, school, and office building; it contains traces of dirt, food, grease, oil, and certain household cleaning products depending on the source. The quantity and quality of gray water are quite different from one house to another as well as from one time to another due to the type of personal care that has been used, shampoo brand, and inclusion of urine and diaper washing in the bath (Mohamed et al. 2014). Grey water is divided into three categories including laundry, bathroom, and kitchen wastewater. In the residential, bathroom gray water represent 60% of the total gray water; 40% are dishwasher, sink, and laundry gray water (Jamrah and Ayyash 2008). The characteristic of gray water is highly different (Table 1). Among these wastes the laundry gray water contains high concentrations of the pollutants due to presence of powdered detergents which has high salt concentrations and phosphorus. The characteristic of bathroom gray water produce in household activities depends on the number of occupants, age, health status, lifestyle, and water usage patterns (NSW Government 2008). As a result, the bathroom gray water may vary greatly in physical, chemical, and biological characteristics. Moreover, the quality of the bathroom gray water is characterized by the water quality test. According to the previous studies presented in Table 1, it can be noted that the physical, chemical, and microbiological characteristics of bathroom gray water

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Table 1 Characteristic of bathroom grey water from previous studies Parameters pH TSS Turbidity BOD COD Ca Na E. coli TN TP

Units – mg/L NTU mg/L mg/L mg/L mg/L Cell/mL mg/L mg/L

Mohamed et al. (2014) 6.1  0.06–6.4  0.21 78  4.55–163  7.12 NR 40  0.25–105  0.42 445  2.52–621  4.02 6  0.63–20  0.45 72  0.78–85  1.03 NR 10  2.90–38  0.56 3  0.87–20  1.76

Ghaitidak and Yadav (2013) 7.1–7.6 58–78 59.8 129–173 230–367 NR 112 82.7 6.6 NR

Eze et al. (2015) 6.30  0.42 305.0  6.0 41.0  1.0 9.7  0.10 15.0  0.8 40.10  2.0 57.2  2.8 NR NR NR

TSS total suspended solids, BOD biological oxygen demand, COD chemical oxygen demand, TN total nitrogen, TP total phosphorus, NR non-reported

exhibited high differences. The total suspended solids (TSS) ranged from 58 to 305 mg/L, while biological oxygen demand (BOD) and chemical oxygen demand (COD) were between 9.7 and 173 mg/L as well as between 15 and 621 mg/L, respectively. The total nitrogen (TN) ranged from 6.6 to 38 mg/L, and total phosphorus (TP) ranged from 3 to 20 mg/L. However, these wastes have low load of E. coli (82.7 cell/mL). The bathroom gray water contains various elements generated from the utilization of cleaning substances such as detergents. The detergents are known as the primary source of phosphates in the gray water but not yet had been prohibited as phosphorus-containing detergents (Eriksson 2002). The most common elements in the gray water are Na, Al, and Zn. The Na concentrations in many of powder detergents ranged from 200 to 700 mg/L (Handreck and Black 2002). The common chemical contaminants from bathroom gray water include soap, shampoo, hair dye, toothpaste, and cleaning products (Mohamed et al. 2013). These substances represent a source for nitrogen; however, the high concentrations of nitrogen in the gray water from urban regions might be due to pass urine during bathing (Mohamed et al. 2014). Gross et al. (2008) indicated that the detergents might also are source for boron which is one of the dangerous elements. The dispose of gray water with boron effects negatively on the plants even in a small amount because it was toxic. Ahmad et al. (2009) stated that the boron reduces the reproductive growth of plant and leading failure of newly divided cells to enlarge.

Grey water Treatment by Filtration Systems The bathroom gray water with nutrients and pathogenic organisms introduced from several sources should be managed probably to avoid the creation of bad odors, stagnant water, and breeding sites for mosquitoes, as well as to prevent the occurrence

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of eutrophication in the natural water systems and reduce the health risk for human. The filtration system is among several technologies used for the treatment of gray water; the filter is defined as a device or an instrument used for removing of pollutants from the liquid wastes. It used as a primary treatment for the wastewater or an advanced process depends on the filtration media. The filtration system is consisted of gravel filter medium, mostly crushed, hard limestone which is efficient for removing lint, fats, hair, and grease (Allen et al. 2010). However, a specific filtration system is required from removing soluble pollutants such as nitrogen and phosphorus from the wastewater (Dunets and Zheng 2013). Mohamed et al. (2014) examined peat filter media for the treatment of bathroom gray water. Based on the method that has been used, the designed system exhibited a significant removal of BOD within 7–28 days; the removal percentage ranged from 52% to 74%. However, the multicomponent filters consisted of clamshell, steel slag, limestone, and sand occurred 98% of turbidity, 96.2% of TSS, and 80% of COD within 3 days.

Ceramic The utilization of ceramic as filters becomes popular for water treatment system, but most of the ceramic are being used for point of applications only (National Academy of Sciences 2008). Ceramics have high biocompatibility, high resistance in corrosion, high resistance in compression, and low electrical and thermal conductivities (Sáenz et al. 1999). The filtration made from the ceramic is a simple technology which made by mixing of clay with organic materials locally and sustainably sourced through burning process and leave small pores throughout the filter (Wald 2012). Ceramic is made from clay and it is widely used in the ceramic industries; the clays have high ion exchange capacity and can adsorb certain anions and cations exchanged for other anions and cations in an aqueous solution (Obaje et al. 2013). OyanedelCraver et al. (2014) indicated that the impregnated clays which contain traces of crystalline albite or crystalline pyroxene with silver compounds showed high adsorption efficiency for metal removal. One more property of the ceramic is the small pores in the filter which can be used to remove small organism such as bacteria while the colloidal silver acts as a chemical biocide to kill microbes (Wald 2012). The ceramic filtration media used in the design of filtration system are illustrated in Table 2. Erhuanga et al. (2014) studied the efficiency of ceramic water filtration as medium to removal of microbes or other contaminants from water. The study revealed that the efficiency removal reaches to 100% of the metals such as Pb. Abiriga and Kinyera (2014) at Kampala, Uganda, stated that a double filtration made from different mixture ratios of clay powder, fine sawdust, and powder of grog removes E. coli by 100%. Nasir and Faizal (2016) found that the treatment system using mixture of natural clay, rice bran, and iron powder to form ceramic filters removed cadmium (Cd) by 99.0%. Subriyer (2013) indicated that the ceramic filter from natural clay and fly ash can remove Fe and Zinc ions by 95%. Based on these studies, it can be proven that ceramic water filter can remove the contamination of

Location Akure, Nigeria

Kampala, Uganda

Indralaya, Indonesia

Sriwijaya University, Indonesia

Researcher Erhuanga et al. (2014)

Abiriga and Kinyera (2014)

Nasir and Faizal (2016)

Subriyer (2013)

Contaminated water

Pulp industry effluent

Tap water

Ceramic filters made from a mixture of natural clay, rice bran, and iron powder in removing cadmium

Ceramic filter made of natural clay, fly ash, and iron powder

Greywater sources Wells, boreholes, and stream

Double filtration on the rate of water percolation and E. coli removal efficiency of ceramic water filters (clay powder, fine sawdust, and powder of grog)

Greywater treatment system Ceramic water filtration used of a porous ceramic (fired clay) as medium to filter microbes or other contaminants from water

Table 2 Efficiency of using ceramic in different treatment system

Ball clay mineral and hard wood sawdust (1 mm sieves) to obtain powder Natural clay (), rice bran (500 μm), iron powder (500 μm) 77.5% natural clay, 20% fly ash, and 2.5% iron powder

Particle size (mm) –

Fe ions (99%) Zinc ions (96%)

Cadmium (99.0%)

Pollutant removal Lead (Pb) (100%) Sulfate (SO4) (76%) Iron (Fe) (47%) Fluoride (75%) E. coli (100%)

–

Effective in removing cadmium

–

Advantages Effective in the treatment of physiochemical contaminants detected in the water samples

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water and wastewater. The efficiency of this material very surprised and hold the potential to be valuable sources for this treatment in removing many of pollutants from the wastewater.

Size and Surface of Media The removal efficiency of the ceramic in the pollutant removal depends on the proper choice of filter depth, sand type, sand size, and filtration (Abudi 2011). The smaller grain sizes fill between the large grain sizes and make the sand easier for the filtration process (National Small Flows Clearinghouse 1997). In contrast, the fine particles might facilitate the breakdown of organic compounds and recovery of nutrients. Torrens et al. (2009) stated that the effective desired pollutant removal depends on particle size and distribution, conditions of influent water, quality of effluent, the filtration rate, and dosing regime and resting period duration. Culp et al. (1978) claimed that the separating suspended and colloidal impurity process of water takes place from water through the filtration media material. The water fills the pore of the filter medium, while the impurities are adsorbed on the surface or trapped in the openings. Crites and Techobanoglous (1998) indicated that the filtration performances of media characteristics are depending on effective grain size and coefficient of uniformity. Thus, it affects the retention time of liquid or water and thus the potential for clogging inside the pore surface of grains. The small material size has a large surface area which might exhibit more efficiency for adsorption of particles and adequate pore space which can promote aeration and unsaturated flow in media (Ball 1997). According to Erhuanga et al. (2014), ceramic used in the design treatment systems are available in three sizes which are 1.18, 0.60, and 0.25 mm.

Conclusion In a nutshell, this study provides a theoretical base for research and helps to determine the nature of the research. This chapter may also serve as a handy guide for ceramic filtration systems. For future studies, the extension of time should be improved in order to identify the effectiveness of ceramic filter systems in the removal of gray water contamination.

Cross-References ▶ Development In-House: A Trap Method for Pretreatment of Fat, Oil, and Grease in Kitchen Wastewater Acknowledgments The authors gratefully acknowledge assistance provided under the contract grant, FRGS vot 1453, to support our research needs.

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References Abiriga F, Kinyera SO (2014) Water purification by double filtration using ceramic filters. Environ Nat Resour Res 4(2):92 Abudi Z (2011) The effect of sand filter characteristics on removal efficiency of organic matter from gray water. Al-Qadisiya J Eng Sci 4(2):143–155 Ahmad W, Niaz A, Kanwal S, Rasheed MK (2009) Role of boron in plant growth: a review. J Agric Res 47(3):329–338 Al-Gheethi AA, Radin Maya Saphira RM, Efaq AN, Amir HK (2016) Reduction of microbial risk associated with gray water utilized for irrigation. Water Health J 14(3):379–398 Allen L, Christian-Smith J, Palaniappan M (2010) Overview of gray water reuse: the potential of gray water systems to aid sustainable water management. Pacific Institute, Oakland, p 654 Al-Mashaqbeh OA, Ghrair AM, Megdal SB (2012) Grey water reuse for agricultural purposes in the Jordan Valley: household survey results in Deir Alla. Water 4(3):580–596 Ball HL (1997) Optimizing the performance of sand filters and packed bed filtered through media selection and dosing methods. Proceedings ninth northwest on-site wastewater treatment short course. College of Engineering, University of Washington, Seattle, pp 205–213 Crites R, Technobanoglous G (1998) Small and decentralized wastewater management systems. McGraw-Hill, New York Culp RL, Wesner GM, Culp GL (1978) Handbook of advanced wastewater treatment (No. Ed. 2). Van Nostrand Reinhold Co. Ltd Dunets A, Zheng Y (2013) Greenhouse and nursery water treatment information system. Media filter. Retrieved on 7 Nov 2016, from www.ces.uoguelph.ca Erhuanga E, Kashim IB, Akinbogun TL (2014) Development of ceramic filters for household water treatment in Nigeria. Art Des Rev 2(01):6 Eriksson E (2002) Potential and problems related to reuse of water in households. Ph. D. Thesis. Technical University of Denmark Danmarks Tekniske Universitet, Department of Environmental Science and Engineering Institut for Miljøteknologi Eze VC, Onwuakor CE, Mgbeokwere EU (2015) Comparative analysis of the microbiological and physicochemical characteristics of gray water sources in off-campus hostels at Michael Okpara University of Agriculture, Umudike, Abia State, Nigeria. Int J Curr Microbiol App Sci 4(8): 196–205 Friedler ERAN, Galil NI (2003) On-site gray water reuse in multi-storey buildings: sustainable solution for water saving. In: Efficient 2003–2nd international conference on efficient use and management of urban water supply, Tenerife, Canary Islands: Spain Friedler E, Hadari M (2006) Economic feasibility of on-site gray water reuse in multi-storey buildings. Desalination 190(1–3):221–234 Ghaitidak DM, Yadav KD (2013) Characteristics and treatment of gray water – a review. Environ Sci Pollut Res 20(5):2795–2809 Gross A, Wiel-Shafran A, Bondarenko N, Ronen Z (2008) Reliability of small scale gray water treatment systems and the impact of its effluent on soil properties. Int J Environ Stud 65(1): 41–50 Handreck KA, Black ND (2002) Growing media for ornamental plants and turf. UNSW Press, Sydney Jamrah A, Ayyash S (2008) Grey water generation and characterization in major cities in Jordan. Jordan J Civ Eng 2(4):376–390 Jefferson B, Laine A, Parsons S, Stephenson T, Judd S (1999) Technologies for domestic wastewater recycling. Urban Water 1, 285–292 Juan A, Medina C, Morán JM, Guerra MI, Aguado PJ, De Rojas MIS, Frías M, Rodríguez O (2010) Re-use of ceramic wastes in construction. In Ceramic Materials. InTech Lesikar B (2005) Chapter 1. Introduction to onsite wastewater treatment systems. OWTS 101: Basics of Onsite Wastewater Treatment Systems. p 4

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Mohamed (2011) Environmental impacts of gray water use for irrigation on home gardens. Ph.D. Thesis, Murdoch University, Western Australia Mohamed RMSR, Kassim AHM, Anda M, Dallas S (2013) A monitoring of environmental effects from household gray water reuse for garden irrigation. Environ Monit Assess 185(10): 8473–8488 Mohamed R, Saphira RM, Chan CM, Senin H, Kassim M, Hashim A (2014) Feasibility of the direct filtration over peat filter media for bathroom gray water treatment. J Mater Environ Sci 5(6):2 Mohamed RMSR, Al-Gheethi AA, Jackson AM, Amir HK (2016) Multi component filters for domestic gray water treatment in village houses. J Am Water Works Assoc 108(7):405–414 Nasir S, Faizal S (2016) Ceramic filters and their application for cadmium removal from pulp industry effluent. Int J Technol 7(5):786–794 National Academy of Sciences (2008) Ceramic filtration. Retrieved from http://drinking-water.org/ html/en/Treatment/filtration-Systems-technologies National Small Flows Clearinghouse (1997) In summer 1997 pipeline: sand filters provide quality. Low Maint Treat 8(3):1–8 NSW Government (2008) Water for life. Department of Water & Energy, Sydney Obaje SO, Omada JI, Dambatta UA (2013) Clays and their industrial applications: synoptic review. Int J Sci Technol 3(5):264–270 Ottoson J, Stenström TA (2003) Faecal contamination of gray water and associated microbial risks. Water Res 37(3):645–655 Oyanedel-Craver V, Narkiewicz S, Genovesi R, Bradshaw A, Cardace D (2014) Effect of local materials on the silver sorption and strength of ceramic water filters. J Environ Chem Eng 2(2):841–848 Sáenz, Muñoz, Brostow, Castaño (1999) Ceramic biomaterials: an introductory overview. J Mater Educ 21(5–6):297–306 Siracusa G, La Rosa AD (2006) Design of a constructed wetland for wastewater treatment in a Sicilian town and environmental evaluation using the emergy analysis. Ecol Model 197(3): 490–497 Subriyer N (2013) Treatment of domestic water using ceramic filter from natural clay and fly-ash. J Eng Stud Res 19(3):71 Torrens A, Molle P, Boutin C, Salgot M (2009) Impact of design and operation variables on the performance of vertical-flow constructed wetlands and intermittent sand filters treating pond effluent. Water Res 43(7):1851–1858 Wald I (2012) Modeling flow rate to estimate hydraulic conductivity in a parabolic ceramic water filter. Undergrad J Math Model One+ Two 4(2):6

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R. M. S. Radin Mohamed, Adel Al-Gheethi, A. N. Welfrad, and M. K. Amir Hashim

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Kitchen Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Fat, Oil, and Grease (FOG) from Kitchen Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Effects of Kitchen Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Technology of FOG Trap Pretreatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Gravity Separators . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Electrocoagulation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Bioaugmentation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Model and Manufacture of FOG Trap Available in Market . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Previous Studies Conducted on the Kitchen FOG Wastewater Treatment . . . . . . . . . . . . . . . . . . . . Challenge and Development of Interceptor for FOG in Malaysia . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cross-References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

Most of the commercial restaurant and domestic house kitchen does not equip with the proper kitchen wastewater treatment. These wastes contain high concentrations of fat, oil, and grease (FOG) which are disposed into the sewerage network and increase the sanitary sewer overflow (SSO). The present chapter discusses the individual treatment systems which are used for the removal of FOG from the kitchen wastewater. The most common and cheapest FOG treatment system is by using the gravity separation. However, in order to increase the efficiency of these systems, some of the researchers have been using natural R. M. S. Radin Mohamed · A. Al-Gheethi (*) · A. N. Welfrad · M. K. Amir Hashim Micro-pollution Research Centre (MPRC), Department of Water and Environmental Engineering, Faculty of Civil & Environmental Engineering, Universiti Tun Hussein Onn Malaysia, Parit Raja, Johor, Malaysia e-mail: [email protected]; [email protected] © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_35

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materials to enhance the adsorption process of FOG. The establishment of a reliable FOG treatment system for each house based on the separation system might contribute effectively in the FOG removal concentration in the kitchen wastewater. Keywords

Fog · Kitchen wastewater · Removal · Individual usage

Introduction Nowadays, fat, oil, and grease (FOG) blockage becomes a worldwide concern due to their role in the increasing water pollution. According to American Environmental Agency (USEPA 2004), 10,350–36,000 sanitary sewer overflow (SSO) cases are occurring every year (Husain et al. 2014). In Malaysia, it has been estimated that the total quantity of wastewater from municipal and industrial sectors which enter sewerage systems is 2.97 billion cubic meters per year (Mat et al. 2012). Moreover, the population growth rate in Malaysia is 1.5% (Labourforce Survey 2015). The increasing of the population is associated with the increasing of the quantities of the wastewater and thus the level of risk in terms of the severity of the problem and the likelihood of the blockage of sewerage system. Therefore, a proper planning to design and manage the FOG wastewater is required to maintain the sewer system efficiency. Nonetheless, the high load of FOG in the wastewater treatment plant reduces the efficiency of the sewage treatment process. For this reason, the Department of irrigation and drainage (DID) in Malaysia provides a guideline for the FOG trap, to ensure the entire restaurant kitchen equip with sufficient design of FOG trap system. However, as a result of the absence of the awareness on the nature conservation issue, most of the restaurants in Malaysia have not equipped with a proper kitchen wastewater treatment unit to treat the FOG contents in kitchen wastewater. The FOG trap represents one of the most effective pretreatment methods to separate and remove FOG from the kitchen wastewater. There are many different types of FOG trap that are available in the markets. In the Economic Transformation Program (ETP) introduced by Malaysia Prime Minister in 2013, the grease trap system upgrading work is only focusing in the Kuala Lumpur and Klang Valley. The area is chosen because it is the country capital city to a world-class city and globally competitive economic hub (Annual Report, Malaysia Federal of Governor of the Federal Reserve System). High installation cost and operation and maintenance cost lead the restaurant owner in the rural and suburban area to take least effort to properly maintain the system. FOG from kitchen wastewater receives the least concern in wastewater treatment, especially for domestic and commercial pretreatment process. However, the untreated kitchen wastewater that is discharged into the sewer trunk results in sewer blockage and causes SSOs. The first process of kitchen wastewater treatment at the restaurant is by sorting and removing the large debris such as the bone, food waste, and kitchenware. The washing process of plate and kitchenware using

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Fig. 1 Improper waste water pretreatment and filtration by a local restaurant in Malaysia (Photo was taken on 10 November 2016)

Fig. 2 Drainage blockage due to kitchen wastewater (Photo taken on 10 November 2016)

detergent and water washes the FOG and the leftover food waste into the filtration mesh screen. In some restaurant, the screening is conducted by using a plastic basket with 3 mm of diameter which is able to screen a very few amounts of the kitchen waste; meanwhile the fine debris and FOG contamination will flow into the drainage sewer system without the proper filtration system (Fig. 1). The observation made at the premise found that the drainage system is a drainage blockage due to solidified FOG and the kitchen waste as shown in Fig. 2. The improper FOG filtration system does not only result sewer blockage but also cause the bad odor around the restaurant. Untreated kitchen wastewater disposal also leads to the increase of the contamination load at the nearby river.

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FOGs in wastewater create problems including the production of foul odors, interference with proper operation of wastewater treatment works, and the blockage of sewer lines leading to sanitary sewer overflows (SSOs). About 50% of SSOs occur as a result of line blockings with the largest source at 47% of these blockages attributed to FOG deposits that accumulate in sewer lines (US EPA 2004). SSOs can potentially release high pathogens, nutrients, and solid loadings that result in harm to public health and the environment. This chapter aimed to highlight and review the previous kitchen wastewater treatment effectiveness used for the treatment of kitchen wastewater.

Kitchen Wastewater Kitchen wastewater is defined as a heavily polluted wastewater with food particles, oil, fat, and grease. The FOG concentrations in the commercial restaurant kitchen wastewater range from 50 to 2100 mg/L. These concentrations play an important role in the growth of microorganisms in the sewer line and create odor for open trench drainage (Chen et al. 2000). The solid food waste particles are combined with the FOG molecule and result the FOG to harden and solidify, thus causing sewer blockage. In the United States, FOGs are generated from commercial food and beverage service establishments which are accumulated in sewer line systems in the form of hardened solids and are responsible for approximately 50–75% of annual sanitary sewer overflows (SSOs) (Ducoste et al. 2008). Kitchen wastewater sources that discharge to the sewer system include garbage grinders, prewash sinks, prep sinks, dishwashers, floor drains, tilt kettles, steam trays, floor sinks, floor drains, and hood-cleaning residue.

Fat, Oil, and Grease (FOG) from Kitchen Wastewater Domestic wastewater with high contents of organic matter is mainly derived from kitchen wastes, which consisted of the remnants of vegetable oils and animal fats as well as proteins and suspended solids. The composition of these wastes varies considerably depending on the food restaurant activity that produces the different types of food (Alves 2013). FOG is a term used to represent the lipid layer formed on the surface of the wastewater generated from the cooking and food processing. These FOGs might be accumulated on the pipe walls and form hardened deposits as a result of the chemical reaction with the other chemical substances in the wastewater or due to the physical aggregation process (He 2012a). The deposits are the main reason for the reduction of conveyance capacity and the occurrence of sanitary sewer overflow. The annual production of collected grease trap waste and uncollected FOG entering sewerage treatment plants in the United States ranges from 350 to 7700 kg/year per restaurant. EPA stated that the combined sewer overflow and sanitary sewer overflows FOG from homes and restaurants as well as from the industrial sources are the

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most common reasons for sewer blockages. Grease is problematic when it solidifies and affects the conveyance capacity and blocks the sewer line. FOG is presented in the wastewater as liquid or solid substances and is characterized by a greasy texture in its pure state. FOG is colorless, odorless, and tasteless. Moreover, FOG is soluble in organic solvents such as ether, hexane, and chloroform but not in water (Husain et al. 2014). Oils and fats are a subsection of lipids that are composed of fatty acids, triacylglycerols, and lipid soluble hydrocarbons that are minor but important components of FOGs. FOG has a density less than water (specific gravity is less than 1), and thus, it floats on the water surface. However, FOG forms emulsions with aqueous media in the presence of soap or other emulsifying agents. Gravity separates free-floating oils from water because their specific gravities are less than 1. Petroleum-based oils typically can be removed from wastewater by skimming them off the top of sedimentation basins. Kitchen wastewater is physically and chemically polluted as it also contains detergent and cleaning agent, and where dishwashers are used, the kitchen wastewater is in alkaline form. According to the Arizona department of environment quality guideline, there are five types of oil that are found in the wastewater (Hess et al. 2016). The most common type is free oil where oil is present in water having little if no water associated with it and can be easily separated by gravity separation method because the oil will rise rapidly to the surface of the water tank under calm conditions. Mechanically emulsified oil is categorized as oil dispersed in water in a stable form as 20–150 μm droplets. It is stabilized by electrical charges and other forces that result in the coating of suspended solids. Such oils mix with water due to shear that can result from the wastewater travelling through a pump, wastewater splashing into a tank, and anything that will break up and disperse larger oil droplets. Oil dispersed in the water droplets less than 2.4-DNP > PNP > MNP. The final conclusion is that hydrophobic FAU zeolite is a good adsorbent and can be easily regenerated (Kobeissy et al. 2008). BEA (β) and Zeolite Socony Mobil-5 (ZSM-5) zeolites as adsorbents for removal of phenol and the performance of the adsorbents was compared with activated carbon, studied by heat-flow micro-calorimetry. The values obtained of heats evolved during phenol adsorption indicate the heterogeneity of active sites present on the investigated systems for the adsorption of phenol. The equilibrium adsorption isotherm data were interpreted using Langmuir, Freundlich, Dubinin-Astakov, and Sips’ equations. The Sips equation was found to express a high level of agreement with experimental data. The results reveal that the adsorption of phenol on zeolites depends on both the Si/Al ratio and pore size. Hydrophobic zeolites that possess higher contents of Si show higher affinities for phenol adsorption. Among two zeolites, zeolite β possesses the highest capacity for adsorption of phenol (Damjanović et al. 2010). Using natural attapulgite and zeolite, a nano-structure adsorbent was manufactured, namely attapulgite–zeolite composite, and used as adsorbent for removal of phenol, and it was concluded that these zeolites are good adsorbents (Wang et al. 2010).

Siliceous Materials Natural siliceous materials such as bentonite and perlite have been utilized as adsorbents to remove phenol from an aqueous solution (Koumanova and PeevaAntova 2002). The study investigated the effect of initial adsorbate concentration and adsorbent mass on adsorbent using the batch adsorption experiments. The adsorption equilibrium of p-chlorophenol (p-CP) on bentonite and perlite was described by the Langmuir and Freundlich models. A higher adsorption capacity was observed for bentonite (10.63 mg g1) than for perlite (5.84 mg g1). The kinetic results reveals that among the adsorbents studied HiSiv 1000 has the highest adsorption capacity. The effect of particle size, temperature, and thermal regeneration were also studied on adsorption of phenol onto HiSiv1000. The adsorption capacity decreased by increasing the particle size and adsorption capacity decreased with increasing temperature. The effect of a mesoporous molecular sieve (M) AlMCM-41 (M = Na+, K+, Cu2+, Cr3+) was examined with regards to the adsorption capacity of chlorinated phenol at a temperature ranging from 303 to 323 K. Different isotherm models were studied to determine the best fit of the curve and the Freundlich model was found to be the best fit. The adsorption capacity of pentachlorophenol (PCP) increased in the order (K) Al-MCM-41 < (Cr) Al-MCM41 < (Na) Al-MCM-41 < (Cu) Al-MCM-41. The isosteric analysis was carried out on all the specified adsorbents, but (Cu) Al-MCM-41 and (Na) Al-MCM-41 presented a heterogeneous profile, whereas (K) Al-MCM-41 and (Cr) Al-MCM-41 did not (Marouf-Khelifa et al. 2004). Chitosan-coated perlite (CCP) beads, which

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contain 23% chitosan, were used as an adsorbent for removal of phenolic compounds (phenol, 2-chlorophenol, and 4-chlorophenol) by Kumar et al. (2010). Batch and column flow experiments were conducted to determine the adsorption capacity of the beads and the effect of different parameters, and isotherm models and kinetic models were studied to identify the adsorption capacity as well as the rate of reaction of the adsorbent. The experimental data best fits the Langmuir model and pseudo first-order kinetic model.

Biosorbents Biosorption occurs due to a number of metabolism-independent processes (physical and chemical adsorption, electrostatic interaction, ion exchange, complexation, chelation, and micro-precipitation) that essentially take place in the cell wall, where the mechanisms responsible for the pollutant uptake will differ according to the biomass type (Vijayaraghavan and Yun 2008; Aksu 2005). The main advantages of biosorption are high selectivity and efficiency, cost effectiveness, and good removal performance. Biosorbents mainly come under the following categories: bacteria, fungi, algae, industrial wastes, agricultural wastes, and other polysaccharide materials. Raw materials such as seaweeds or wastes from other industrial operations (fermentation wastes, activated sludge process wastes) can also be used as biosorbents. Their performances are often comparable with those of ion exchange resins. Biosorption is a method of fast and reversible binding of pollutants from water and wastewater onto functional groups that are present on the surface of biomass. Tsezos and Bell (1989) observed that the uptake of PCP was found to be adsorbed more by dead cells of Rhizopus arrhizus (4800 μg/g) than the live cells (800 μg/g). Uptake of a compound by live biomass is principally due to biodegradation, which in turn depends upon the biodegradability of the compound in question. Chitin is a natural homopolymer comprising β-(1-4)-linked N-acetyl-D-glucosamine. Chitin is the most abundant and renewable natural polymer. It is found in the exoskeletons of crabs and other arthropods and in the cell walls of some fungi and is a waste product of the crab meat canning industry. Due to the presence of one linear amino group per glucose ring, chitin can be used as a metal ion adsorbent. Dursun and Kalayci (2005) studied the adsorption of phenol onto chitin as a function of initial pH, temperature, and initial phenol concentration. The optimum pH value was determined to be 1.0 and they also found that adsorption capacity increased as the temperature increased to 40  C. The experimental data fits the Langmuir and Freundlich adsorption models. Aksu et al. (2002) has studied adsorption of phenol onto the Mowital B30H resin immobilized dried activated sludge as a biosorbent. Freundlich and Langmuir adsorption models were used to represent the column equilibrium data. It was seen that the adsorption equilibrium data fitted very well to the Langmuir adsorption model.

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Soil as Adsorbent In recent decades, studies of the interaction between soils and organics have been numerous; however, their conclusions have often been contradictory. This is most likely due to disparity in the different experimental methods adopted and the different adsorbent/adsorbate systems investigated. Studies have investigated the adsorption isotherms of benzene and polychlorinated benzenes on dry Woodburn soil (Chiou and Shoup 1985). The soil environment is a complex system; when agricultural pollutants enter into the soil system, the concentration present in the soil system depends on the retention action of the applied agricultural chemicals. The potential mobility of dissolved chemicals in soils has been described by several models. The most commonly used isotherm models such as Freundlich and Langmuir are used to study chemical waste distribution or segregating between the soil matrix and the sorbate. Various studies have investigated soil heterogeneity and kinetics of adsorption–desorption reactions models (Zhu 2002). The soil matrix system showed widespread hysteretic behavior causing inconsistencies between adsorption and desorption isotherms. Hysteresis was more distinct with an improved adsorption reaction time. Johnson and Sims (1993) conducted batch equilibrium studies of five 14C-labeled herbicides, including metolachlor, in selected soils The outcomes showed retention of metolachlor was best correlated with the organic matter content and cation exchange capacity of the soil. Peat has been widely used for the adsorption of heavy metals and dye house effluent from wastewaters due to its excellent ion exchange properties (Leslie 1974; Belkevich et al. 1976). McKay et al. (1981) conducted experimental batch studies to determine the parameters that affected surface mass transfer of Astrazone Blue onto peat during the initial stages of adsorption. They proposed a three-step model: (i) the mass transfer of dye from the bulk solution to the surface of the adsorbent; (ii) external diffusion of the adsorbent; and (iii) intra-particle diffusion. McKay and Allen (1984) used peat as an adsorbent to remove dyes (such as Acid Blue 25, Basic Blue 69, etc.) from aqueous solutions. They also developed a model based on external mass transfer irreversible adsorption and internal pore diffusion. Cardoso et al. (1985) investigated the effectiveness of two peat samples for adsorbing selected heavy metals from aqueous solutions. It was found that peat had a good uptake capacity for removal of heavy metals from wastewater as well as from the aqueous solution. The adsorption capacity of copper(II) uptake by peat was limited to 12.6 mg/g. Allen et al. (1992) has investigated sphagnum peat to adsorb copper, cadmium, and zinc from aqueous solutions. Boyd et al. (1988), Brigatti et al. (1995), Gutierrez and Fuents (1996), and Lo et al. (1997) have studied exploiting naturally available clay minerals such as montmorillonite, kaolinite, and illite for the removal of Zn2+, Pb2+, and cesium and also pental-chlorophenols from aqueous media. Viraraghavan and Kapoor (1994) investigated removal of mercury form wastewater using bentonite and reported that 34% of mercury was removed with an initial concentration of 1 mg/l. Ghiaci et al. (2004) studied the potential of adsorption to remove hazardous liquids (such as benzene, toluene, and phenol) by organo-zeolites,

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which were prepared from synthetic ZSM-5 and natural zeolites. Tayim and Al-Yazouri (2005) investigated local soil (from the Emirate of Abu Dhabi in the United Arab Emirates) efficiency in removing heavy metals (such as Pb, Zn, Fe, Cu, and Mn) from industrial wastewater. The authors concluded that using locally available soils in an effective and low-cost technique may encourage polluting industries to be more environmentally sustainable. The results indicated that montmorillonite showed a general higher sorption capacity with respect to kaolinite and that for both the reference clays, in the concentration range investigated. Maiti et al. (2007) investigated natural laterite as an adsorbent for uptake of arsenite [As(III)] from aqueous solution. A speciation study revealed that about 20% of As(III) is converted to As(V). Therefore, natural laterite seemed to play the role of a natural oxidant for arsenite and the extra step of pre-oxidation of As(III) is not required. Instead of natural virgin soils, modified soils (chemically and physically) were used as adsorbents for removal of pollutants from wastewaters in most of the research. Boyd et al. (1988) used modified smectite with cetyltrimethylammonium bromide (CTAB-smectite) to adsorb trichloroethylene and benzene and found that sorption is affected by partition action. Sameer et al. (2003) examined various types of activated bentonite for removal of phenol from aqueous media, because of activation bentonite renewed into cationic surfactant bentonite, exhibits maximum adsorption of phenol. Juang et al. (2003) used surfactant-modified montmorillonite (treated with cetyltrimethylammonium bromide) as adsorbent for removing phenol, m-nitrophenol (m-NP), and o-cresol. Akal (2005) studied the effect of treated pumice on phenol and 4-chlorophenols. The results show a higher adsorption capacity than that of treated bentonite and vermiculite.

Agriculture and Industrial Residue/By-Products Agricultural residue, especially cellulose, exhibits adsorbent qualities to treat water and wastewater. Agricultural residues are economic and eco-friendly due to their unique chemical composition, abundant availability, renewable nature, and low cost, making them a viable option for water and wastewater remediation. Through chemical and thermal treatment, agricultural residues are converted into activated carbon and are potential adsorbents (Ahmedna et al. 2000). Using this method one can solve environmental problems and also reduce the preparation costs.

Carbons Prepared from Residues/Wastes A comparative study of the sorption of different pollutants from water has been carried out using novel adsorbents prepared by carbonization and subsequent activation of straw and used rubber tires as well as conventional activated carbons based on coal, coconut shells, and wood. Pseudo-equilibrium sorption of sorbate equilibrium data fitted Freundlich isotherm well and measured values of k and n indicate that the sorption of sorbate onto derived active carbons is almost identical to that for

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S. Busetty

conventional activated carbons. The sorption kinetics of the straw and rubber tirebased carbons is likewise identical to conventional carbons based on coal and wood but appreciably faster than the coconut shell sample selected for comparison (Streat et al. 1995). Srivastava et al. (1997) attempted the removal of pollutants by carbonaceous adsorbents obtained from fertilizer waste using 2,4-DNP as the adsorbate and reported a constant adsorption for the pH range of 2.0–4.0 and a reducing adsorption trend with a pH more than 4.0, as supported by the pKa value of the adsorbate (i.e., 3.96), indicating the probable preference of the nondissociated species of the adsorbate for the negatively charged surface of the carbonaceous adsorbent. Rengaraj et al. (2000) also examined the suitability of palm seed coat for the adsorption of ortho-cresol (ocresol) and found it to have an adsorption capacity of 19.58 mg/g with film diffusion as the rate-limiting step. Tseng et al. (2003) prepared pine wood-activated carbons at different activation times (0.5, 1.5, 2.7, and 4.0 h) in steam at 900  C and batch studies were conducted to determine the adsorption capacity and rate of reaction on to three dyes and three phenols (phenol, 3-chlorophenol, and o-cresol) from aqueous media. For all six adsorbates, the carbons prepared with a longer activation time had a larger adsorption capacity. Eucalyptus grandis sawdust, a major waste from the wood industry, was used to prepare granular activated carbon (GAC) by mixing Powdered activated carbon (PAC), carboxymethyl cellulose (CMC) as a binder, and kaolin as reinforcer (Tancredi et al. 2004). In another study, beet pulp, an agricultural low-cost by-product in the sugar industry, was used to study adsorption phenol by preparing activated carbon. The BET surface area of the beet pulp carbon was measured as 47.5 mg/g and the maximum adsorption uptake was found to be 89.5 mg/g at the temperature of 60  C at pH = 6.0. The Langmuir and Freundlich isotherm were used to fit the experimental data and equilibrium data were well-described by Freundlich isotherm (Dursun and Kalayci 2005).

Industrial Solid Waste (By-Products/Waste Materials) Industrial solid wastes (by-products and waste materials) are cheap and locally available waste materials. The disposal problem relating to these materials is solved if they are recycled as adsorbents for removing pollutants from the water and waste waters. Industrial solid wastes such as blast furnace flue dust, slag (generated in chrome alloy plants), fly ash, and red mud can be used as adsorbents for the removal of pollutants from the water and waste waters. Singh et al. (1994) studied adsorption of phenol at different pH values, using fly ash as an adsorbent. The results reveal that the adsorption capacity of phenol was nearly the same at different pH values. However, the decreases in pH result in a reduction of the negative charges at the surface of fly ash, thus enhancing the adsorption of the negatively charged phenolic ions. In general, the adsorption capacity of phenolic compounds from water and waste water increases with decreasing pH. The adsorption capacity of phenol increases with increasing particle size of the sorbate and decreases with increasing concentration of the sorbate and at lower temperature.

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Das and Patnaik (2005) studied phenol removal using industrial waste in the form of flue dust (from a steel plant) and slag (chrome alloy plants) as adsorbents, with a removal efficiency of phenol of 90% and 75%, respectively, for these two adsorbents. The adsorption process is thermodynamically favorable, spontaneous, and endothermic in nature. The experimental data are best fitted to the Langmuir isotherm model. The adsorption of phenol on carbon-rich bagasse flyash (BFA) was studied by Srivastava et al. (2006). BFA is a solid waste obtained from the particulate collection equipment attached to the flue gas line of the bagasse-fired boilers in cane sugar mills. A pH ~ 6.5, adsorbent dose ~ 10 g/l of solution, and equilibrium time ~ 5 h were found to be optimum conditions for phenol removal from the solution. The phenol adsorption equilibrium data were fitted to different isotherm models – Freundlich, Langmuir, Temkin, Redlich–Peterson, Radke–Prausnitz, and Toth models – using a non-linear regression technique. Among these isotherms, the Redlich–Peterson isotherm was found to best represent the data. The change in entropy (ΔS0) and heat of adsorption (ΔH0) for phenol adsorption on BFA were noted to be 1.8 MJ/kg K and 0.5 MJ/kg, respectively. Adsorption of phenol on to BFA was spontaneous because of the high negative value of change in Gibbs free energy (ΔG0).

Agro-Residue as Adsorbent There have been extensive studies on use of agro-residue as adsorbent. Nawar and Doma (1989) studied the adsorptive capacities of rice hulls for two industrial textile dyes and found encouraging results as well successfully applying the Freundlich model. Kenaf (lignocellulosic fibers) was found to be effective for adsorption of several toxic heavy metals (namely nickel, copper, zinc, and cadmium) from storm water. The adsorption potential of these Kenaf plants was found to be related to their sugar content, extractives composition, lignin content, and physical properties. It was also found that a decrease in the lignin and cellulose content in Kenaf contributed to a lower density and easy accessibility of ions to the reactive sites on Kenaf’s surface, thus increasing the adsorptive capacity (Han 1999). Brown et al. (2000) demonstrated the relatively lower capacity of raw peanut hulls (and peanut hull pellets) than that of the ion-exchange resin; yet the substantially lower cost of production of the former than the latter would probably compensate for their deficiency more than adequately to justify their use in the treatment of low-strength metals-contaminated waste streams. Figueiredo et al. (2000) explored some natural adsorbents comprising chitin, namely squid (Loligo vulgaris) and sepia (Sepia officinalis) pens and Anodonta (Anodonta cygnea) shells, which were used to remove color from textile wastewaters and exhibited promising adsorption capacities. In addition, removal of Cr(VI) from aqueous media was batch studied using the husk of the Bengal gram (Cicer arietinum). At low concentrations (i.e., 10 mg/l) of chromium the removal efficiency was found to be 99.9% (Ahalya et al. 2005). The adsorption capacity of castor seed shell (CSS), a solid waste agricultural by-product, was used to remove methylene blue from wastewater; the adsorption capacity was

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S. Busetty

found to be 158 mg/g (Oladoja et al. 2008). Xue et al. (2008) studied removal of two basic dyes, namely malachite green and methylene blue, which were removed using rice bran and wheat bran by-products as adsorbents. Recently, there has been increasing interest in evaluating the use of practically every possible natural adsorbent for adsorption, resulting in various degrees of success. Besides naturally available soils and agro-residue (biomass), the use of several chemically synthetic and non-biodegradable wastes as adsorbents has been attempted by several researchers. Although non-biodegradable wastes are naturally of lower adsorption capacity than the naturally available soils and agro-reside-based adsorbents, due to their increased magnitude and resultant solid-waste disposal problems, attempts have been made to find at least some use for them, including in the field of adsorption. Luchesi and Maschio (1983) used carbon (with a surface area of 320 m2/g), produced from stripped-tire rubber, as an adsorbent for removal of Orange II and Acid Black 24 dyes from the aqueous phase. The Freundlich isotherm adsorption model was used to express the sorption phenomenon of the sorbate. Rengaraj et al. (2002) further investigated activated carbon prepared from rubber seed coat (RSCC), an agricultural waste by-product, for removing phenols from aqueous solution. They found that RSCC is 2.25 times more efficient than Commercial Activated Carbon (CAC). Tanthapanichakoon et al. (2005) investigated the liquid-phase adsorption–desorption characteristics of reactive dyes. The ethanol regeneration efficiency of the prepared AC saturated with either tested dye was found to be higher than that of the commercial Activated Carbon (AC).

Correlation and Estimation of Performance of Adsorption Systems A variety of different isotherm equations have been proposed, some of which have a theoretical foundation and some being of a more empirical nature. Many of these equations are valid over small relative pressure ranges but do not fit the experimental data when tested over the full range of relative pressures. Only those that are generally used for the surface of porous adsorbents are outlined here. Large numbers of researchers in the field of environmental engineering have used Freundlich and Langmuir isotherm equations to represent equilibrium adsorption data using activated carbon–organic contaminants systems. To determine the adsorption potential of the soil and soil agro-based adsorbents for the removal of phenol, study of adsorption isotherms has been carried out and tested against multi-parameter isotherm models (up to five parameters): two-parameter models – Langmuir, Freundlich, Modified Langmuir–1, Modified Langmuir–2, Dubinin–Radushkevich, and Temkin; three-parameter models – Langmuir–Freundlich, Redlich–Peterson, Sip, Fritz–Schlunder model, Radke–Prausnitz model, Toth model, and Jossens model; four-parameter models – Fritz–Schlunder model, and Baudu; and fiveparameter model – Fritz–Schlunder. In an adsorption study, describing the sorption process and evaluation of the bestfitting isotherm model is a key analysis to explore the theoretical hypothesis.

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Hence, several statistical analyses have been widely used to estimate the validity of the experimental equilibrium adsorption values with the predicted equilibrium values. Several error analyses have been carried out (average relative error deviation [ARED], Marquardt’s percentage standard error deviation [MPSED], hybrid fractional error function [HYBRID], sum of the squares of the errors [SSE], correlation coefficient, and residuals). All of these isotherm models are derived based on theoretical assumption, and several error deviation functions have been used to adequately measure the goodness of fit of the models, such as the correlation coefficient (r2), SSE, ARED, MPSED, HYBRID, and residual analysis (RESID). However, the very approach of linearization of the non-linear models necessarily yields rationalization of specific variables, which may have significant bearing on the adsorption process itself. The generalized isotherm studies are being carried out to identify the suitability of selected raw/treated adsorbents. The study of adsorption of the pollutants at equilibrium concentration (at constant temperature) is referred to as adsorption isotherm.

Isotherm Modeling Bi-Parameter Model Langmuir Isotherm Model The two-parameter Langmuir’s isotherm equation is as follows: qe ¼

x q bC e ¼ m m 1 þ bC e

(2)

where x is the amount of adsorbate adsorbed (mg), m is the mass of adsorbent (g), Ce is the equilibrium concentration of the solute in the bulk solution (mg/l), qe is the amount of solute adsorbed per unit weight of adsorbent at equilibrium (mg/g), qm is the maximum adsorption capacity (mg/g), and b is the constant related to the free energy of adsorption (l/mg). The Langmuir model is generally a better model for the adsorption of gases onto solids, whereas the Freundlich model is a better model for the adsorption of liquid solutions (Cooney 1999). The simplified form of the Langmuir equation is given as follows: Ce Ce 1 ¼ þ qe qm bC e

(3)

Freundlich Isotherm Model The Freundlich model is by far the most widely used isotherm model to determine maximum adsorption capacity (Cooney 1999). The two-parameter Freundlich model is as follows:

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S. Busetty

x ¼ qe ¼ k f C 1=n e m

(4)

where x is the mass of solute adsorbed (mg), m is the mass of adsorbent (g), Ce is the final aqueous phase concentration (mg/l), kf and n are empirical constants, and qe is the amount of adsorbate adsorbed per unit mass of adsorbent. The value for the constant kf is typically reported at a water-phase equilibrium concentration of 1 mg/l (i.e., Ce = 1), when the equation is transformed to the form mx ¼ k f , and, thus, k (adsorptive capacity) has the units of w/w or mg/g for the example stated. The Freundlich model assumes that the energy distribution for the adsorption sites is exponential in nature (Cooney 1999). At high concentrations, the equation would fail to fit experimentally (Cooney 1999). In such cases, the logarithmic form of the Freundlich equation can be expressed as follows: log

x 1 ¼ log qe ¼ log k f þ log C e m n

(5)

The linear line obtained gives a slope, which is the value of 1/n, and the Y-intercept is log k. The lower the slope, the better the adsorption, over the entire range of concentrations, whereas the steeper the slope, the better the adsorption at high concentrations. A greater value of ‘kf ’ indicates a higher capacity of adsorption. Modified Langmuir–1 The modified Langmuir–1 is given as follows: qe ¼

bT n C e 1 þ bC e

(6)

where qe is the amount of adsorbate adsorbed per unit mass of adsorbent, Ce is the final aqueous phase concentration (mg/l), T is the temperature (in Kelvin), and b is the constant related to the free energy of adsorption (expressed as l/mg). Modified Langmuir–2 The modified Langmuir–2 is given as follows: bC e qe ¼ 1 þ bC e

! 1 þ σ 2 ð1  bC e Þ 2ð1 þ bC e Þ2

(7)

where qe is the amount of adsorbate adsorbed per unit mass of adsorbent, Ce is the final aqueous phase concentration (mg/l), and σ is the isotherm constant related to the degree of sorption. Dubinin–Radushkevich Isotherm Dubinin–Radushkevich isotherm is often used to estimate the characteristic porosity in addition to the apparent free energy of adsorption. Thus, for evaluating these parameters the Dubinin–Radushkevich isotherm is used in the following form:

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Environmental Treatment Technologies: Adsorption

"

2 # 1 qe ¼ K DR exp b RTln 1 þ Ce

1383

(8)

where qe is the equilibrium adsorption capacity (mg/g), Ce is the adsorbed equilibrium concentration (mg/l), KDR relates to the free energy of sorption, and b is the Dubinin–Radushkevich isotherm constant related to the degree of sorption.

Temkin Model The Temkin isotherm is typically expressed in the following form: Qe ¼

RT ðlnAC e Þ b

(9)

where Qe is adsorption equilibrium capacity (mg/g), Ce is the adsorbed equilibrium concentration (mg/l), and A and b are the isotherm constants.

Three-Parameter Models Langmuir–Freundlich Isotherm The Langmuir–Freundlich isotherm is given as follows: qe ¼

qmLF ðK LF C e ÞmLF 1 þ ðK LF C e ÞmLF

(10)

where qe is the adsorbed amount at equilibrium (mg/g), qmLF is the Langmuir–Freundlich maximum adsorption capacity (mg/g), Ce is the adsorbed equilibrium concentration (mg/L), KLF is the equilibrium constant for a heterogeneous solid, and mLF is the Langmuir–Freundlich heterogeneity parameter which lies between 0 and 1. The Langmuir–Freundlich isotherm model at low sorbate concentrations effectively gets reduced to the Freundlich isotherm and thus does not follow Henry’s law. At high sorbate concentrations, it behaves as monolayer sorption and shows the Langmuir isotherm characteristics.

Fritz–Schlunder Equation The Fritz–Schlunder model equation is given as follows: qe ¼

qmFS K FS C e 1 þ qmFS C e /

(11)

where qe is the adsorption equilibrium capacity (mg/g), Ce is the equilibrium concentration of the adsorbate (mg/L), qmFS is the Fritz–Schlunder maximum adsorption capacity (mg/g), KFS is the Fritz–Schlunder equilibrium constant (L/mg), and α is the Fritz–Schlunder model exponent.

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S. Busetty

Radke–Prausnitz Models The Radke–Prausnitz model equation is given as follows (Vijayaraghavan et al. 2006): qe ¼

qmRP K RP C e ð1 þ K RP C e ÞmRP

(12)

where qe is the equilibrium adsorption capacity (mg/g), qmRP re the Radke–Prausnitz maximum adsorption capacities (mg/g), Ce is the adsorbate equilibrium concentration (mg/l), KRP are the Radke–Prausnitz equilibrium constants, and mRP are the Radke–Prausnitz model exponents. Toth Model To reduce the error between experimental and theoretical values, Toth has modified the existing Langmuir isotherm. The Toth model equation is given as follows: qe ¼ 

C e qmT 1 KT

þ C mT e

mT1

(13)

where qe is the adsorbed amount at equilibrium (mg g1), Ce is the equilibrium concentration of the adsorbate (mg l1), qmT is the Toth maximum adsorption capacity (mg g1), KT is the Toth equilibrium constant, and mT is the Toth model exponent. Jossens Model The Jossens model is as follows: Ce ¼

  qe exp Fqpe H

(14)

where qe is the adsorbed amount at equilibrium (mg/g), Ce is the equilibrium concentration of the adsorbate (mg/L), and H, F, and p are Jossens constants. H and F depend only on temperature. This equation can reduce to Henry’s law at low capacities. Redlich–Peterson Model The Redlich–Peterson (1959) model is a three-parameter model given by: qe ¼

x AC e ¼ m 1 þ BC βe

(15)

The linearized form is: ln

AC e  1 ¼ lnB þ β lnC e qe

(16)

where x is the moles of adsorbate adsorbed (mg), m is the mass of the adsorbent (g), and Ce is the final adsorbate concentration (mg/l).

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The three constants are A, B, and β. The parameter A is measured in units of mg/g, B is measured in the h units ofi (l/g), and β lies between 0 and 1 (Redlich and Peterson 1959). e Plotting ln AC q  1 against lnCe yields a straight line of slope β and intercept e

lnB. However, plotting of this equation is not applicable because of the three unknown parameters contained within the equation. Therefore, a minimization procedure is adopted (for maximum R2 correlation coefficient) between the theoretical data for qe predicted from Eq. 15 and the experimental data. qe ¼

RT RT lnðK T Þ þ ln ðC e Þ bT bT

(17)

where 1/bT indicates the adsorption potential of the adsorbent and KT is the Temkin isotherm constant. Temkin isotherm contains a factor that explicitly takes into account the interactions between adsorbing species and the adsorbate. This isotherm assumes that (i) the heat of adsorption of all the molecules in the layer decreases linearly; and (ii) adsorption is characterized by a uniform distribution of binding energies, up to a maximum binding energy (Kim et al. 1997). Sip Model The Sip isotherm model equation is expressed as follows: ðK S C e Þβ i qe ¼ h 1 þ ðαS C e Þβ

(18)

where KS (l/g) and α (l/mg)β are the Sip isotherm constants and β is the exponent that lies between 1 and 0.

Four-Parameter Models Fritz–Schlunder Model The Fritz–Schlunder isotherm model equation is expressed as follows: qe ¼

AC / e 1 þ BC βe

(19)

where A and B are the Fritz–Schlunder parameters and α and β are the Fritz–Schlunder equation exponents. Baudu The Baudu equation is expressed as follows: qe ¼

qm b0 C ðe1þxþyÞ 1 þ b0 C ðe1þxÞ

(20)

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where qm is the Baudu maximum adsorption capacity (mg/g), b0 is the equilibrium constant, and x and y are the Baudu parameters.

Five-Parameter Model Fritz–Schlunder Model The Fritz–Schlunder model equation is expressed as follows: qe ¼

1 qmFSS K 1 C m e m2 1 þ K 2Ce

(21)

where qmFSS is the Fritz–Schlunder maximum adsorption capacity (mg/g) and K1, K2, m1, and m2 are the Fritz–Schlunder parameters.

Estimation of Best-Fitting Isotherm Model Error Functions ARED is used to minimize the fractional error distribution across the entire concentration range:   1 X Qe,cal  Qe,exp  ARED ¼   100    N Qe,exp

(22)

The SSE, the most widely used error function, has a major drawback. The theoretical isotherm parameters calculated from the sum of the squares of the error function will result in a better fit at the higher end of the sorbate concentration range. This is due the magnitude of the errors, and hence the error function will increase as the concentration increases. SSE ¼

X

Qe,cal  Qe,exp

2 (23)

HYBRID is an error function that was developed to improve the SSE at low concentration values. In this task, each sum of the squares of the error values was divided by the theoretical adsorbent-phase concentration value.   1 X Qe,exp  Qe,cal  HYBRID ¼    100   N P Qe,exp

(24)

The MPSED is an error function distribution that follows the geometric mean error, which allows for the number of degrees of freedom of the system.

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MPSED ¼

vffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi 2 uP  u ðQexp Qcal Þ = t Qexp N P

1387

 100

(25)

The sum of the absolute errors (EABS) is similar to the SSE function and provides a better fit at higher concentrations for the isotherm parameters: EABS ¼

p   X   Qe,exp  Qe,cal 

(26)

i¼1

Statistical Functions Pearson’s correlation coefficient (r) is a sampling index that shows the degree linearity between two dependent data series. The degree of linearity varies from 1 to 1. P P P N ð XY Þ  ð X Þð Y Þ r ¼ rhffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiiffi h i P P P P N X 2  ð X Þ2 N Y 2  ð Y Þ2

(27)

The coefficient of determination (r2) explains the regression line with the percentage of variability in the dependent data series variable. The percentage degree varies from 0 to 1. r2 ¼

S2 S ðXY Þ S ðYY Þ

(28)

where SXY is the sum of squares of X and Y, and SYY is the sum of squares of Y. In addition to above mentioned error and statistical functions, the Chi-square test was also examined to predict the best-fitting isotherm models.

Kinetic Study Adsorption Kinetics Adsorption kinetics are the most important to evaluate and analyze process efficiency. Hence, the equilibrium time and adsorption capacity plays a major role in the design. Sorption equilibrium is recognized once the concentration of sorbate in the bulk solution is in dynamic stability with that of the boundary. The steady analysis is the most vital evidence required to evaluate the affinity. It is consequently important to define in what way the rate of reaction depends on the concentrations of sorbate

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S. Busetty

and also how adsorbent characteristics affect the rate of reaction. From the kinetic analysis, the equilibrium time can be determined, which in turn indicates adsorption process completion. Various kinetic models have been used by researchers, using the pseudo-first-order (Vasanth et al. 2007; Ho and McKay 1999) and pseudo-secondorder models (Ho et al. 2000).

Pseudo-First-Order Kinetics The interaction between adsorbate and adsorbent can be explained by reversible equilibrium kinetics (Fogler 1998). The Lagergren rate equation is the most widely used kinetic first-order equation (Tseng et al. 2003; Ho and McKay 2004): dqt ¼ k t ð qe  q t Þ dt

(29)

where kt (l/min) is the pseudo-first-order rate constant, and qe and qt are the sorption capacity at equilibrium and time t (mg/g). After integration by applying the conditions qt = 0 at t = 0 and at t = t, qt = qt, Eq. 31 becomes: logðqe  qt Þ ¼ logqe 

kt t 2:303

(30)

The value of kt can be calculated from the plots of log(qe  qt) versus t for different concentrations of phenol.

Pseudo-Second-Order Kinetics The pseudo-second-order chemisorption kinetic rate equation (McKay and Ho 1999) is expressed as follows: dqt ¼ k 2 ðqe  qt Þ2 dt

(31)

where k2 is the second-order rate constant (g/mg min). After integration by appalling the conditions qt = 0 at t = 0 and at t = t, qt = qt, Eq. 31 becomes: t 1 t ¼ þ 2 qt k 2 qe qe

(32)

Equation 32 can be written as follows: t 1 t ¼ þ qt h qe

(33)

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where h = k 2 q2e , by plotting of qt versus t has a linear relationship. The values of k2 t and qe can be calculated from the intercept and slope of the plots of qt versus t. t

Elovich Equation The Elovich equation is a useful model; it can be expressed as follows (Aharoni and Tompkins 1970): dqt ¼ αexpðβqt Þ dt

(34)

where α is the initial adsorption rate (mg g1 min1) and β is the desorption constant (g mg1) during any one experiment. To simply the Elovich equation (34) assume αβt >> t, and by applying boundary conditions qt = 0 at t = 0 and qt = qt at t = t, Eq. 34 becomes: 1 1 qt ¼ lnðαβÞ þ lnðt Þ β β

(35)

The plot between qt versus ln(t) yields a linear relationship with a slope of β1 and an intercept of β1 lnðαβÞ . The term β1 indicates the number of sites available for adsorption.

Thermodynamic Studies Study of adsorption involves mass transfer operations, chemical reactions, and energy variations. During these operations, thermodynamic principles are involved. To study the spontaneous nature of adsorption, the thermodynamic parameters should be essentially determined.

Fundamentals of Thermodynamics This section explains the changes in the physical and chemical energy that takes during the thermodynamics effect. Thermodynamics can be roughly encapsulated with the following three parameters: enthalpy (ΔH0), entropy (ΔS0), and free energy (ΔG0) (Mehrian et al. 1991). The work done by the system can be explained by change (endothermic or exothermic) in the enthalpy of a system, which is equal to its heat output at a constant pressure. A negative value of ΔH0 represents energy coming out of the system and indicates the process as exothermic and the sorption behavior as physio-sorptive in nature. It also indicates the free diffusion of molecules through bulk solution. The positive value of ΔH0 indicates the process as

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endothermic and the sorption behavior as being due to chemical bonding or chemisorption (Piatt et al. 1996). Entropy (ΔS0) is a measure of the disorderliness of a system. With the increase in entropy significance a spontaneous change occurs in a system. The positive values of ΔS0 show the increased uncertainty at the solid/solution boundary during the adsorption process. An extemporaneous reaction is one that happens deprived of any outside interference. The negative value of ΔS0 is due to loss of freedom as the molecules adsorb to the surface. The reversibility of the system was explained by a value of ΔS0 less than 1 (Brucher and Bergstrom 1997; Catena and Bright 1989). Developments that are impulsive in one course are non-spontaneous in the reverse direction. The free energy (ΔG0) of the system is defined as the work done by the system at a constant temperature and pressure. Negative values of ΔG0 would anticipate a favorable response and specifies the impulsive nature of the adsorption process. The increasing values of ΔG0 with an increase in temperature show that the impulsive nature of adsorption is inversely proportional to temperature (McCloskey and Bayer 1987).

Column Study Fixed-Bed Adsorption Column Adsorption of pollutants from industrial wastewaters or gaseous pollutants can be achieved by allowing fluid flowing through the fixed granular bed. To evaluate the feasibility and economics of the adsorbent, a laboratory study is usually conducted. The batch study gives the removal capacity of specific waste contaminants, but the continuous flow provides a practical application of the process. For industrial applications, a fixed-bed adsorption column study is an important step. In order to stop the adsorption period in a process before saturation, a thorough understanding of adsorption characteristics is necessary (Bautista et al. 2003). The column study for adsorption starts with prototype testing to create the breakthrough curve. The amount of adsorbate adsorbed within a bed depends both on space and time. The sorbate passes into the bed, and it immediately contacts the first few strata of absorbent. As the top strata of adsorbent become saturated with adsorbate, the mass transfer zone (MTZ) will move down in the bed until breakthrough occurs. The depth of the MTZ depends on the hydraulic loading rate and the characteristics of the adsorbent. The point at which the outlet concentration of the fluid starts to rise is known as the “breakpoint” and the point at which the effluent concentration almost equals the influent concentration is known as the “exhaust point.” For a narrow MTZ, the breakthrough curve is very steep. If the mass transfer rate is faster and no axial dispersion is present, then the MTZ width would be zero and the breakthrough curve would be a vertical line. The overall dynamics of the system determine the efficiency of the system. Adsorption ability is calculated by plotting a graph of concentrations versus time. By integrating the following equation, adsorption capacity can be measured in terms of time units:

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1 ð

C 1 dt co

tt ¼

(36)

0

where tt is the time to total capacity. In operation, the process needs to be stopped to get the “usable” capacity and the curve can be integrated as follows: t break ð

t Usable ¼



c 1 dt co

(37)

0

Further, most of the time the breakthrough time is very close to the time elapsed at usable capacity. The capacity times are directly related to bed depth: H Used ¼ H Total

t usable t Total

(38)

Therefore:

t usable H unused ¼ 1  H Total t Total

(39)

HUnused represents the MTZ. It depends on the fluid velocity and is independent of the total length of the column. HUsed is directly proportional to tbreak. The unused depth can also be readily measured in experiments. The total design depth of a bed is determined by adding the required usable capacity to the unused height: H Total ¼ H Used þ H unused

(40)

Bed Depth Service Time (BDST) In a column study a continuous mass transfer can be determined between two phases, i.e., the sorbate and the adsorbent as a fixed bed. It is a function of the depth of the zone and contact time. The aim of the column study is to study the effect of several experimental trails on the bed depth service time (BDST) model. The BDST model describes a correlation between the service time of the column and the fixed-bed column. The original work on the BDST model was carried out by Bohart and Adams (1920). The BDST equation is as follows: h ln

Co

   i =Ce  1 ¼ ln eKN 0 H=V  1  K C o t

(41)

where Co is the influent concentration (mg/l), Ce is the allowable effluent concentration (mg/l), K is the rate constant [m3/kg of adsorbent (h)], No is the adsorptive

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capacity (kg of contaminant/m3 of adsorbent), H is the depth of adsorbent bed (m), V is the linear flow rate (m/h), and t is the time in hours. Since eKN 0 H=V >> 1, Eq. 41 can be rewritten as:

No V Co H ln 1 KN o CoV Ce

No H 1 Co  ln 1 t¼ C o V KN o Ce

No 1 Co ½H  ln 1 t¼ KC o CoV Ce

t¼

(42)

(43)

(44)

Biodegradation Bioremediation is a low-cost technology for attacking sites polluted with organic pollutants (Yeom and Ghosh 1998). The governing and methodological challenge is to use low-cost technologies that can Identify complex organic mixtures in polluted soils and thus reduce the threat to human health and the environment. The challenge is also to decrease the toxicity of the organics and the movement of potential hazardous elements being treated. Thus, bioremediation is a long-term risk-protective process because it acts as a source elimination, pollution-lowering, and risklowering technique to prevent ground water contamination and therefore decreases the necessity for additional expenses. Although several bioremediation procedures are in use, selecting the process suitable for a specific site is a function of the remediation objectives to be attained, the physical and chemical characteristics of the substrate to be treated, the environmental circumstances that are formed, the materials handling and equipment necessary, and the availability of funds. Several studies have been conducted in the past to test aerobic degradation of phenol. Most of the laboratory studies reported had involved pure strains of microorganisms isolated from naturally occurring systems (Erik et al. 1989). Acclimation of the microorganisms to the organic compound in question has been an important step in the biodegradation process. It could also differ depending on concentration, environment, temperature, aeration status, and other often undefined factors (Alexander 1994). Buitron et al. (1993) have used organisms obtained from activated sludge to degrade 50 ppm of phenol in less than 2 h. D’Aquino et al. (1988) performed experiments to try and degrade high concentrations of phenol as seen in industrial effluents within a 24 h period. They successfully isolated two strains of bacteria (Pseudomonas and Acinetobacter). Both strains were able to completely biodegrade phenol and benzoate at concentrations above 200 ppm within 12–16 h. A heterogenous mixture of microbial population from swine waste was tested for its ability to degrade phenol. This consortium readily oxidized phenol and p-cresol

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under aerobic conditions (Ishaque et al. 1985). Biodegradation of phenol using yeast and fungi has been studied by Katayama et al. (1991), Kennes and Lema (1994), and Anselmo and Novais (1992). Most of the fungi degrade xenobiotics by co-metabolic reactions, only a few being able to utilize phenol as their sole source of carbon and energy. Cordova-Rosa et al. (2009) have studied remediation of phenol-contaminated soil by a mixed bacterial culture and isolated Acinetobacter calcoaceticus obtained from a coal wastewater treatment plant containing high phenolic compounds.

Cross-References ▶ Advanced Treatment Technologies ▶ Decentralized Integrated Approach of Water and Wastewater Management in Rural West Bengal ▶ Hazardous Waste Management with Special Reference to Biological Treatment ▶ Soil Pollution and Remediation

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Making Artificial Beachrock Through Bio-cementation: A Novel Technology to Inhibition of Coastal Erosion

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Md. Nakibul Hasan Khan and Satoru Kawasaki

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Beachrock . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . World Distribution of Beachrock . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Properties of Beachrock . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Age of Beachrock Formation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Origin of Beachrock . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Artificial Beachrock . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Formation of Artificial Beachrock . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Bio-cementation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Bio-cementation Technique . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Microbially Induced Carbonate Precipitation (MICP) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . The Role of Urease Activity in Carbonate Precipitation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cementation Process Through Bacterial Action . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Applications of Carbonate Precipitation Through Bacteria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sand Cementation Through Bacterial Carbonate Precipitation . . . . . . . . . . . . . . . . . . . . . . . . . . . . Inhibition of Coastal Erosion Through Artificial Beachrock . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

1400 1402 1402 1404 1404 1407 1408 1408 1409 1409 1412 1412 1413 1415 1417 1418 1419 1420

Abstract

Coastal erosion is a significant problem throughout the world. In order to prevent or minimize damage from erosion, combinations of various structures have been used traditionally. The maintenance and management for repairing and rebuilding the coast are expensive. As a hint for an alternative material in order to reduce M. N. H. Khan (*) Department of Environmental Science and Engineering, Jatiya Kabi Kazi Nazrul Islam University, Mymensingh, Bangladesh S. Kawasaki Faculty of Engineering, Hokkaido University, Sapporo, Japan © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_38

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life-cycle costs, it is focused on the beachrock. A beachrock is composed of coastal sediments that have been cemented mainly by CaCO3 within the intertidal zone, and its formation period is much shorter than that of other sedimentary rocks. The cement component of beachrocks mainly consists of either calcium carbonate or silica. Compared to the concrete structure, coral sand solidification would be considered to minimize cost. A variety of factors in the formation of beachrock has been considered. Among them are the possibility of solidification promotion by microbial action, the urea decomposition action of microorganisms, and the microbially induced carbonate precipitation (MICP) method; as principles, carbon dioxide generates as well as precipitates CaCO3 by microbial metabolism. Bio-cementation technology is used to make artificial beachrock. Bio-cementation is a sand solidification technology, in which ureolytic bacteria release carbonate from urea hydrolysis in the presence of an excess of calcium ions to form calcite (CaCO3). Bio-cementation is to enhance the strength and stiffness properties of soil and rocks through microbial activity or products. Bacterial CaCO3 precipitation under appropriate conditions is a general phenomenon where the ureolytic bacteria uses urea as an energy source and produces ammonia which increases the pH in the environment and generates carbonate, causing Ca2+ and CO32 to be precipitated as CaCO3. This CaCO3 joins sand particles and forms rocklike materials that auto-repairs by means of sunlight, seawater, and bacteria as microbially induced carbonate precipitation method. These rock particles produced artificially are called artificial rock, and this artificial rock has the potentiality to protect coastlines from erosion. Keywords

Beachrock · Artificial beachrock · Bio-cementation · Coastal erosion · Calcite precipitation · Ureolytic bacteria · Urease · Sand cementation · Microbially induced · Carbonate precipitation · Erosion · Coastal sediment · Calcium carbonate · Cement · Microorganism

Introduction Some islands in the world are in danger of being submerged due to erosion and sea level rise as impacts of climate change. Erosion of the sandy shore is often used to refer to changes in the coastline due to the collapse of the sediment balance. Coastal erosion has been a significant problem globally due to anthropogenic changes along the coastline. In order to prevent, or at least minimize damage from erosion, a combination of various structures and processes has been traditionally used, including embankments, revetments, jetties, artificial reefs, offshore breakwaters, and sand bypassing (Danjo et al. 2014). But installment and management of these technologies are expensive. The use of inexpensive alternative materials should be considered, to cope with the increase in costs related to the maintenance and management of concrete structures. In this regard, Danjo and Kawasaki (2013a) focused on alternative

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materials to replace beachrock to reduce life-cycle costs associated with the currently employed methods. Beachrock forms much more quickly than other sedimentary rocks within the intertidal zone and is composed of coastal sediments that have been cemented together mainly by CaCO3 (Danjo and Kawasaki 2013a). By using the processes that solidify beach sand, we hypothesized that we could create a highly durable artificial beachrock that would be comparable in efficacy to the existing concrete structures. Artificial beachrocks (manufacture artificially in the natural condition) have the potential to inhibit coastal erosion (Danjo et al. 2013b). Beachrocks are coastal deposits cemented mainly by calcium carbonate cement; these deposits are found in the tidal and intertidal zone of sandy beaches in tropical and subtropical regions (Danjo and Kawasaki 2013a). For the formation of beachrock, various factors have been considered. Among them, Danjo and Kawasaki (2014), focused on the possibility of promoting solidification by microbial processes, specifically urea decomposition by microorganisms. In this study, the microbial induced calcium carbonate precipitation (MICP) method was utilized, which relies on the microbial metabolism of urea that generates carbon dioxide and precipitates CaCO3 (Whiffin et al. 2007). This low environmental impact method was assessed to determine its efficiency as an alternative means to alleviate coastal erosion. Considering the use of artificial rock in order to preserve such submerged-looking islands above sea level, Danjo and Kawasaki (2012) and Danjo et al. (2012) conducted several studies in Okinawa and Ishikawa, Japan. They found sufficient information to build artificial beachrock. Bio-cementation is to enhance the strength and stiffness properties of soil and rocks through microbial activity or products. It could be used to prevent soil avalanching, reduce the swelling potential of clayey soil, mitigate the liquefaction potential of sand, and compact soil on reclaimed land sites (Ivanov and Chu 2008). Traditional grouting methods for ground improvement employ particulate (cement/ bentonite) or chemical grouts that can be rather expensive and environmentally unfriendly (Ivanov and Chu 2008). Recently, novel grouting techniques have been developed to treat unsaturated coarse soils by stimulating natural processes (DeJong et al. 2006; Whiffin et al. 2007). One of these methods, termed biogrouting, has shown some promise in soil cementation via microbially induced carbonate precipitation (MICP). This approach mimics natural processes by depositing calcite (CaCO3) on the soil grains, thereby increasing the material’s stiffness/strength and reducing its erodibility. In ground treatment applications for sandy soil, the deposition of calcite over the grain surfaces and around the grain contacts creates a sandstone-like material. In principle, MICP treatment protocols can be tailored to produce a more targeted deposition of calcite around the grain contacts, with the porosity decreasing by less than 10% (Shahrokhi-Shahraki et al. 2015). Hence, significant improvements in the geomechanical properties can be achieved while maintaining permeability. In this chapter beachrock, its global distribution, origin, artificial beachrock, formation of process of artificial beachrock, and its potentiality to protect coastal erosion are discussed.

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Beachrock Beachrock is a type of sedimentary deposit that generally occurs on tropical and subtropical beaches as a result of intertidal lithification of loose beach sands and gravels by carbonate cementation (Ginsburg 1953). The Association for the Geological Collaboration in Japan (2000) defined beachrock as follows: An extremely recent, consolidated, calcareous rock occurring in the intertidal zone on sandy beaches. It comprises multiple layers, each around 1–60 cm thick, and runs roughly in the same direction as the beach, at an oblique angle of 5–7 to the sea. It forms a cuesta shape, with the landward side having an acutely angled face. Its overall thickness is about 1.0 m, and may be separated into two or more bands. The cement material may be calcareous or may contain iron, and the non-cement substances may be any material, regardless of the grain size or substance. Beachrocks are found on beaches in tropical and sub-tropical zones. In the seas around Japan they can be found in the Nansei Island group, which is southwest of the mainland.

Around the world beachrocks have been reported to form over several thousand years owing to interactions among sand supply, cement precipitation from seawater, and coastal erosion by ocean waves (Danjo and Kawasaki 2013a). An example of beachrock of Okinawa, Japan, is shown in Fig. 1.

World Distribution of Beachrock The geographic distribution of beachrocks around the world is shown in Fig. 2. As can be seen in the figure, over 90% of beachrocks are distributed between the proximity of 40 N and the Tropic of Capricorn. This proportion, however, does not represent the total size of the distribution area but rather the number of significant distributions. A large number of beachrocks are located in the Aegean coast (Area A), southern India (Area B), the Great Barrier Reef (Area C), the West Indies (Area D), and eastern Brazil (Area E). The northernmost point for the beachrock distribution was in Scotland (40 57/N), while the southernmost point was in South Africa (34 60/S). The geographic distribution of beachrocks was not concentrated near the equator. In particular, high concentrations of beachrocks were found at relatively high latitudes such as Area A. This precludes the notion that beachrocks are concentrated only in areas where the temperature of the air or sea is high. Though, it is difficult to determine why beachrocks are distributed the way they are. For example, in Area B beachrocks are distributed in the south and northwest of India but not in the southwest region. Similarly, there are no beachrock occurrences in the eastern area of Africa, except in areas close to the equator. Beachrocks throughout the world are found in a consolidated state in close proximity to non-consolidated sediments, as is shown in Fig. 1, which shows

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Fig. 1 Beachrock in Motobu, Kunigami, Okinawa, Japan (Danjo and Kawasaki 2013b)

Fig. 2 Beachrock distribution in the world (Danjo and Kawasaki 2013b)

beachrock in Okinawa, Japan. This suggests that the conditions that led to their formation are complex (Danjo and Kawasaki 2013b). There may be a tendency for beachrocks to form in places where the ocean currents are calmer, such as in Area A, which is surrounded by land, and in Area C, which is surrounded by coral reefs. However, there is a need for further research to identify why beachrocks form in some places and not others.

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Fig. 3 Beachrock compositions around the world (Khan et al. 2015)

India (Vattakottai) India (Kanyakumari) Japan (Ishikawa)

India (Tiruchendru) Japan (Okinawa) Bangladesh (Satkhira)

100 90 80 70

%

60 50 40 30 20 10 0 CaO

SiO2

Al2O3

TiO2

FeO

Properties of Beachrock More than 90% of beachrocks are distributed between 40 N latitude and the Tropic of Capricorn and that their formative periods range from 26,000 years to just a few decades ago (Danjo et al. 2013a). But there is no relation between the formative age of beachrock and latitude (Danjo and Kawasaki 2012). Chemically, beachrock around the world differs in composition (Fig. 3). Beachrock at Tiruchendur, India, mainly consists of Ca. In contrast, the beachrock in Vattakottai, India, is mainly rich in Si and Al, and that in Kanyakumari, India, is mainly rich in Fe and Ti (Danjo and Kawasaki 2012) where Bangladesh is rich in Si (Fig. 3). The main components of beachrocks and surrounding material are calcium carbonate or silica (Danjo et al. 2013a). The proportion of CaCO3 in beachrocks was around 90%, while the other portions consisted of SiO2 and Al2O3 in the Gulf of Mannar, India (Sahayam et al. 2010). The beachrocks are mainly composed of calcium carbonate or silica, similar in Japan and the world (Table 1) and (Fig. 4).

Age of Beachrock Formation The formative periods for beachrocks in the world are shown in Fig. 5. The periods were determined using four methods: carbon-14 dating, optically stimulated luminescence (OSL), content analysis, and the relationship between present-day sea levels and changes in sea level (Danjo and Kawasaki 2013b). The units used to describe the periods were yrBP or yBP for carbon-14 dating, which indicate the numbers of years prior to 1950 (Danjo and Kawasaki 2013b). For OSL and content analysis, years prior to the time of measurement for OSL and content analysis were used.

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Table 1 Compositions of some beachrocks in the world (Danjo and Kawasaki 2013b)

Location Sumuide, Nago, Okinawa, Japan Sosogi, Wajima, Ishikawa, Japan Suzu, Ishikawa, Japan Grand Cayman Heron Island Vattakottai, India Bozcaada Island, Turkey Siesta Key, Florida North Uist, Scotland

Component materials of the beachrock

Main minerals of the beachrock Aragonite, Mg calcite

Sand, gravel, plastic piece, glass piece, iron chain

Shell sand, foraminifer, sand, gravel

Component materials of the cement Mg calcite

Si, Al

Quartz, feldspar, CaCO3

CaCO3

Algae, mollusk, organic material, coral grains Carbonate grains

Coarse sand, small gravels of sandstone, basalt, limestone, and andesite Shell fragments 81%, quartz grains 19%, lithoclasts less than 1% Micritic cement 40%, Fragmented shells 30%, Lithoclasts (mainly quartz, Feldspar, glauconite) 30%

Calcite, aragonite, quartz, heavy minerals High Mg calcite

Low Mg calcite

Calcite, aragonite, Quartz, Glauconite

Danjo and Kawasaki (2013) compared seven areas with respect to the formative periods of beachrock: (1) the Aegean coast (Area A); (2) southern India (Area B); (3) the Great Barrier Reef (Area C); (4) the West Indies (Area D); (5) eastern Brazil (Area E); (6) the northernmost point, which was Scotland; and (7) the southernmost point, which was South Africa. They found that the formative periods in South Africa ranged from several decades ago in Durban to approximately 26,000 years ago around False Bay. There was no distinct concentration of formations in any one period, and the number of available papers from which data could be drawn varied significantly depending on the distribution area. Turning next to the formative period for each distribution area, they also found that the difference between the length of the formative periods within Area A and Area C could be quite great, even at the same latitude and longitude. The widest spread in formative periods for Areas A and C were between approximately 300 and 5,400 years ago and 200 and 6,500 years ago, respectively. These differences in formation times, even within the same latitude and longitude range, may be attributable to differences in the

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Ar and LMC, 1

AI and Si, 1

Silica, 1

LMC, 3 MC and Ar, 9 Aragonite (Ar), 20 Mg calcite (MC), 9

HMC, Ar and Low Mg calcite (LMC), 2 HMC and Ar, 6

C, 14

Ar and Calcite (C), 3

High Mg calcite (HMC), 8

Fig. 4 Beachrock cements around worldwide (Danjo et al. 2013b) N 50 40 30

Latitude (°)

20 10 0 10 20 30 S 40

0

2000

4000

6000

8000

10000

Formative age (years ago) Fig. 5 Formative periods of beachrocks along with latitudes distributed around the world (Danjo and Kawasaki 2013b)

altitude at which the beachrock was sampled, and there were differences in the samples. Data were sparse for Area B, and there were only three samples available. However, the data indicate that formative periods at Area B were concentrated between 3,100 and 3,700 years ago (Danjo and Kawasaki 2013b). Although there

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were few data available for Areas D and E, the data showed that there were large differences in the formative periods for beachrocks in these areas. In Areas D and E, the formative periods ranged between approximately 800 and 4,400 years ago and approximately 200 and 6,900 years ago, respectively (Danjo and Kawasaki 2013b). For the northernmost area, Scotland, no information was reported on the formative periods. In the southernmost area, South Africa, the difference in formative periods was the greatest and ranged from several decades to approximately 26,000 years ago (Danjo and Kawasaki 2013b).

Origin of Beachrock Two contrasting theories were proposed regarding the origin of beachrocks: one cited seawater as the source for beachrocks (Tanaka 1983), and the other claimed that land-based water such as groundwater or spring water was the source (Yonetani 1963). Therefore opinions were divided between these two theories, but more support goes to the theory of seawater as the source for beachrock formation (Tanaka 1983). According to Tanaka (1983), a proponent of the seawater source theory, water analyses of seawater and groundwater around beachrocks on Otsu Beach and Yakomo Beach on Okinoerabujima Island showed that the seawater had a higher content of Ca2+ and a stronger concentration of CO32, while both the seawater and groundwater had a negative ion charge of 2, which indicates that they had a strong electrical component in common. Based on these findings, Tanaka claims that the rising temperature of the water during the day could easily cause the ions in the seawater to precipitate and form CaCO3. On the other hand, the theory that groundwater is the source for beachrocks claims that calcium carbonate precipitates in the surface portion of groundwater form a cement that enables the surrounding sand particles to consolidate when the following conditions occur together (Yonetani 1963): (1) when the original groundwater has a high Ca2+ content, (2) when the groundwater becomes warm due to the effects of sunlight on the ground surface, and (3) when the pH of the water becomes high due to the influence of seawater. Recently, more scientists have expressed the opinion that either of these theories is possible. The different stable isotopes of seawater and inland water allow the source of calcium carbonate to be estimated through measurements of stable isotope ratios (Omoto 2009). After using this method on test samples, Omoto (2009) claims to have found beachrocks that were formed from both seawater sources and inland water sources. Meanwhile, Ogasawara et al. (2004) proposed that the presence of spring water on the mountainous side of Wajima City in the Ishikawa Prefecture in Japan is strongly connected to the formation of some of the beachrock distributions in the area. Theories have also been proposed that assert that living organisms cohabiting in an area, such as coral, seaweed, sea urchins, starfish, and shellfish, also play a role in the consolidation of beachrocks (Tachibana 1964; Yonetani 1963). Since living

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organisms generate ammonia, inorganic CaCO3 precipitates are formed when seawater is added. Furthermore, it has been suggested that landforms, the mixing of marine and meteoric water, and CO2 degassing can affect beachrock formation (Danjo and Kawasaki 2013b). According of Tanaka (1983) and Erginal et al. (2010), the formation process of beachrocks could be summarized as follows: Seawater or inland water seeps into the interstitial space between particles of sand and coral in the intertidal zone and rises to the point where the particles are fully immersed. Then, the water retreats. With the ensuing exposure to sunlight, calcium carbonate or silica precipitates consolidate particles together, and as these develop into aggregates, beachrock is formed (Danjo and Kawasaki 2013b). This scenario suggests that during the creation of artificial rocks, it will be important to understand the quality of both seawater and freshwater. However, since the present water quality at the site of beachrocks that were formed long ago may have changed since the time of their formation, the investigation of water quality should target recently formed beachrocks or those that are currently in formation. The details of such a study will be determined as the present project goes forward.

Artificial Beachrock Natural beachrocks, which are formed naturally on beaches, have attracted attention as a model for artificial rocks. Danjo and Kawasaki (2013a) proposed a new method to protect coastlines from erosion – the use of artificial rock that auto-repairs by means of sunlight, seawater, and bacteria. Their model of artificial rock is beachrock. Artificial beachrock is the rock material which is made artificially in the sandy beaches. Sand particles are particles of coagulated filling voids by precipitation of carbonate (CO32) of calcium and magnesium where CO32 ion is produced by bacterial metabolism. By this process cemented sand makes a rocklike material which is strong and durable like natural rock. As it is made artificially in beach sand, it is called artificial beachrock.

Formation of Artificial Beachrock Regarding the origin of beachrocks, two contrasting theories were proposed: one cited seawater as the source for beachrocks (Tanaka 1983), and the other claimed that groundwater or spring water was the source (Yonetani 1963). Opinions were divided between these two theories; and at the time, there was more support for the theory of seawater as the source for beachrock formation (Tanaka 1983). In this regard Danjo and Kawasaki (2013a) examined a formation mechanism of beachrock in Okinawa, Japan; understanding this natural formation mechanism of beachrock is an important step to making artificial beachrock. They focused on the cement formation mechanism of beachrock, which occurs in the intertidal zone. Cement type and content have the potential to influence the strength of the material; hence, detailed

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Bacteria

B

Org.

Organic matter (including urea)

Bacterial ureolysis Buried beachrock

High tide line

Seawater evaporation B B

B 2+

Ca

B

Ca2+

Mg2+

B

Low tide line

B

Coral sand

Seawater

Ca2+

Org.

Org.

B

Org. B

Ca2+

Exposed beachrock

Fig. 6 Proposed formation mechanism of beachrock at the study site (Danjo and Kawasaki 2013a)

knowledge of beachrock cements would be valuable for producing an artificial equivalent (Danjo and Kawasaki 2013a). Beachrocks are cemented by high Mg calcite (HMC) (Fig. 7) around the world (Erginal et al. 2010). Danjo and Kawasaki (2013a) focused their investigation into the formation mechanisms of the beachrock cements on the influence of precipitation from seawater and/or seawater evaporation (PSW) and on surface microorganisms. Based on formation methods observed for Okinawa, Japan, artificial beachrock is cemented by HMC using microorganisms with urease activity, organic matter such as citrate and malate, nutrient sources, CO (NH2)2, artificial seawater, and sand (Fig. 6) (Danjo and Kawasaki 2014). With respect to PSW, Raz et al. (2000) reported that to better understand the depositional process of high-magnesian calcitic skeletons, they studied the CaCO3 precipitates formed from solutions with Mg/Ca ratios 4. In addition, sodium citrate and sodium malate favor the precipitation of MC, whereby an increase in concentration of magnesium ion and these organic materials causes the formation of Mg-rich calcite (Khan and Kawasaki 2015).

Bio-cementation Bio-cementation Technique Chemical cementation (or chemical grouting) process is used to fill the sand voids with fluid chemical grouts to produce sandstone-like masses to carry loads. This chemical cementation method is widely used in geotechnical engineering. The chemicals that are used to bind soil particles include sodium silicate, calcium

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Biological process (BIOL) and Physico/ chemical factors (P/C-F), 1

Precipitation from fresh water (PFW), 1

Listedness, 3

Precipitation from seawater and/or marine water evaporation (PSW), 10

Sea - fresh water mixing (S-FWM), 1

Fig. 7 Formation mechanisms of beachrock whose cements contain high Mg calcite (Danjo et al. 2013b)

chloride, calcium hydroxide (lime), cement, acrylates, acrylamides, and polyurethanes (Karol 2003). Bio-cementation (or microbial cementation) is a cementation process in which sand particles are coagulated using the metabolic activities of microorganisms. Microbial cementation is to form soil particle-binding material after the introduction of microbes and specific additives into soil. In this method precipitation of carbonate (CO32) of calcium, magnesium, etc., is used to fill the sand voids where CO32 ion is produced by bacterial metabolism. Urea-producing bacteria are used in this process mainly. It is different from biobinding, which is the formation of the particle-binding cellular chains (Ivanov and Chu 2008). Biobinding can be performed by mycelial fungi, actinomycetes, and filamentous phototrophic and heterotrophic bacteria (Ivanov and Chu 2008). In some experiments, the added biomass of some fungal strains binds the sand grains and increases the shear strength of soil (Meadows et al. 1994). However, biobinding does not seem to be suitable for large-scale operations such as enhancing the liquefaction resistance of land reclamation sites due to its instability of biological bindings, and it can be degraded by other microorganisms (Ivanov and Chu 2008). Chemical cementation of soil in nature is due to the precipitation of material in spaces between soil particles and binding of these particles together into a hard rock. Microorganisms are often associated with the cemented sediments, containing calcium, magnesium, iron, manganese, and aluminum, which are crystallized as carbonates, silicates, phosphates, sulphides, and hydroxides, especially iron hydroxides (DeJong et al. 2006). Chemical transformations of metals

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and ions in soil or sand are mediated by soil microorganisms. An example of sand cementation in nature is the formation of ferrihydrite in pores (Ross et al. 1989). Iron hydroxides, depending on its crystallization, can be also an important cementing agent in soils (Duiker et al. 2003). Drying of soil samples containing iron hydroxide can produce irreversible soil hardening and cementation (Ivanov and Chu 2008). An example of natural cementation is the precipitation of silica dioxide, which fills in the pores and glues the soil particles together (Ivanov and Chu 2008). It is also known as natural soil calcification due to the deposition of calcium carbonate from upward flow of groundwater, enhanced evapotranspiration from soil, or formation of calcium carbonate within zones of elevated carbonate alkalinity formed by microbial decay of organic matter (Mozley and Davis 2005). In a patent, Kucharski et al. (2005) applied microbial bio-cementation for the formation of high strength cement in a permeable material using the combination of this material with biomass of urease producing microorganism, urea, and soluble calcium salts. Microorganisms provide fast urea hydrolysis, increase the pH during hydrolysis of urea to ammonia, and form calcite in soil or rocks. The cement produced has a compressive strength up to 5 MPa. The materials, treated by bio-cementation, may be conglomerate, breccia, sandstone, siltstone, shale, limestone, gypsum, peat, lignite, sand, soil, clay, sediments, and sawdust. The urease-producing microorganisms are from genera Bacillus, Sporosarcina, Sporolactobacillus, Clostridium, and Desulfotomaculum. This method of microbial cementation could be used for the following civil and environmental engineering applications (Kucharski et al. 2005): • Controlling erosion in coastal area and rivers • Enhancing stability for retaining walls, embankments, and dams • Reinforcing or stabilizing soil to facilitate the stability of tunnels or underground constructions • Constructing a permeable reactive barrier in mining and environmental engineering • Increasing the bearing capacity of piled or non-piled foundations • Reducing the liquefaction potential of soil • Treating pavement surface • Strengthening tailing dams to prevent erosion and slope failure • Binding of the dust particles on exposed surfaces to reduce dust levels • Increasing the resistance to petroleum borehole degradation during drilling and extraction • Increasing the resistance of offshore structures to erosion of sediment within or beneath gravity foundations and pipelines • Stabilizing pollutants from soil by the binding • Creating water filters and borehole filters • Immobilizing bacterial cells into a cemented active biofilter

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Microbially Induced Carbonate Precipitation (MICP) In bio-cementation process calcium carbonate (CaCO3) or other carbonates is precipitated using microorganism to join sand particles. As carbonates are precipitated using microorganism, it is known as microbially induced carbonate precipitation (MICP). Like other biomineralization processes, calcium carbonate (CaCO3) precipitation can occur by two different mechanisms: biologically controlled or induced (Lowenstan and Weiner 1988). In biologically controlled mineralization, the organism controls the process, i.e., nucleation and growth of the mineral particles, to a high degree. The organism synthesizes minerals in a form that is unique to that species, independently of environmental conditions. Examples of controlled mineralization are magnetite formation in magnetotactic bacteria (Bazylinski et al. 2007) and silica deposition in the unicellular algae coccolithophores and diatoms, respectively (Barabesi et al. 2007). However, calcium carbonate production by bacteria is generally regarded as “induced,” as the type of mineral produced is largely dependent on the environmental conditions (Rivadeneyra et al. 1994), and no specialized structures or specific molecular mechanism are thought to be involved (Barabesi et al. 2007). Different types of bacteria, as well as abiotic factors (salinity and composition of the medium), seem to contribute in a variety of ways to calcium carbonate precipitation in a wide range of different environments. This approach mimics natural processes (Eqs. 1 and 2) by depositing calcite (CaCO3) on the soil grains, thereby increasing the material’s stiffness/strength and reducing its erodibility. The microbiological process relies on ureolytic (nonpathogenic) bacteria such as Sporosarcina pasteurii or Bacillus pasteurii to hydrolyzed urea in the presence of calcium ions, resulting in the precipitation of calcite crystals. Urease

COðNH2 Þ2 þ 2H2 O ! 2NH4 þ CO3 2 yields

Ca2þ þ CO3 2 ! CaCO3 ðsÞ

(1) (2)

The actual role of the bacterial precipitation remains, however, a matter of debate. Some authors believe this precipitation to be an unwanted and accidental by-product of the metabolism (Knorre and Krumbein 2000), while others think that it is a specific process with ecological benefits for the precipitating organisms (Ehrlich 1996).

The Role of Urease Activity in Carbonate Precipitation Urease Urease was the first enzyme to be isolated in its crystalline form from Canavalia ensiformis (jack bean) (Sumner 1926). The first three-dimensional structural model of urease was observed by X-ray of a bacterial source, Klebsiella aerogenes (Jabri et al. 1995). Kl. aerogenes urease consists of four domains, one of which contains an

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active site with a bimetallic nickel center. These primary structures of Kl. aerogenes and Sporosarcina pasteurii (formerly known as Bacillus pasteurii) urease are identical (Mulrooney and Hausinger 1990). Urease is synthesized under the condition of nitrogen starvation (Mobley et al. 1995). There are contradictory statements regarding the location of urease in the bacterial cells. The urease is located in the membrane and periplasm of Staphylococcus sp. and Proteus mirabilis (McLean et al. 1986).

Mechanism of Urease Reaction Urea is released into the environment due to biological action, for example, all mammals excrete urea as a detoxification product. Urease (urea amidohydrolase; EC 3.5.1.5) is widely distributed in soil and aquatic environments. Biotic urease activity is widespread in the environment and includes the action of bacteria, yeasts, filamentous fungi (Mulrooney and Hausinger 1990), algae (Yates and Robbins 1999), and a number of higher plants including jack beans (Canvalia ensiformis), soybean leaf and seed (Glycine max), pigweed (Chenopodium album), and mulberry leaf (Morus alba) (Al-Thawadi 2011). Urease hydrolyzes the substrate urea generating ammonia and carbamate (Eq. 3). Carbamate spontaneously decomposes to produce another molecule of ammonia and carbonic acid (Eq. 4) (Mobley and Hausinger 1989). The two ammonia molecules and carbonic acid subsequently equilibrate in water with their deprotonated and protonated forms, resulting in an increase in the pH (Eqs. 5 and 6) (Mobley and Hausinger 1989). Urease

COðNH2 Þ2 þ H2 O ! NH3 þ COðNH2 ÞOH

(3)

COðNH2 ÞOH þ H2 O ! NH3 þ H2 CO3

(4)

H2 CO3 $ HCO3 þ Hþ

(5)

2NH3 þ 2H2 O ! 2NH4 þ þ 2OH

(6)

Cementation Process Through Bacterial Action Bacterial CaCO3 precipitation under appropriate conditions is a general phenomenon (Boquet et al. 1973). There are a number of species of CaCO3 minerals associated with bacteria, for example, vaterite formation by Acinetobacter sp. (Sanchez-Moral et al. 2003), aragonitic spherulites by Deleyahlophila (Rivadeneyra et al. 1996), calcite by E. coli (Bachmeier et al. 2002), and magnesium calcite spherulites and dumbbells by the slime-producing bacteria, Myxococcus xanthus (González-Muñoz and Chekroun 2000). One of the most robust ureolytic bacteria is S. pasteurii. It is an aerobic, spore-forming, rod-shaped bacterium. It uses urea as an energy source and produces ammonia which increases the pH in the environment and generates carbonate, causing Ca2+ and CO32 to be precipitated as

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Fig. 8 Schematic model for cement precipitation (Al-Thawadi 2011)

(1) Organic substrate (Urea)

Cell Ca2+

Low pH (6.8)

(3)

EPS HCO3–

Micro-environment NH4+ (2)

High pH (9.2)

CO3– (4) CaCO3

CaCO3 (Eqs. 7, 8, and 9) (Stocks-Fischer et al. 1999). Alkaline pH is the primary means by which microbes promote calcite precipitation (Fujita et al. 2000). Based on various studies, a schematic model describing the role of ureolytic bacteria on calcium carbonate precipitation is illustrated in Fig. 8 (Al-Thawadi 2011). Ca2þ þ Cell ! Cell  Ca2þ

(7)

Cl þ HCO3  þ NH3 $ NH4 Cl þ CO3 2

(8)

Cell  Ca2þ þ CO3 2 ! Cell  CaCO3 #

(9)

The schematic model (Fig. 8) is summarizing the role of ureolytic bacteria in CaCO3 precipitation in the presence of Ca2+ ions. The processes involved in CaCO3 precipitation are (1) hydrolysis of urea (Eqs. 3, 4, and 5), (2) increasing the alkalinity of the microenvironment (Eq. 6), (3) surface adsorption of Ca2+ions (Eq. 7), and (4) nucleation and crystal growth (Eqs. 8 and 9). EPS stands for extra polysaccharide in the case of the presence of EPS surrounding the ureolytic cells. There are two metabolic pathways for bacterial carbonate formation. These pathways are autotrophic and heterotrophic pathways (Al-Thawadi 2011). Regarding autotrophic pathway, CO2 is used as a carbon source causing its depletion in the bacterial environment. In the presence of Ca2+ ions, such depletion enhances the production of CaCO3. Regarding heterotrophic pathways, bacteria can precipitate CaCO3 through active or passive precipitation. In active precipitation, the production of CO32 is due to ionic exchange through the cell membrane by calcium and/or magnesium ionic pump. During passive precipitation, the production of CO32 is due to ammonification of amino acids, dissimilatory reduction of nitrate, or degradation of urea or uric

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acid. In all cases, ammonia as a metabolic end product is produced which induces a pH increase. CaCO3 precipitation rate in general is a linear function of the ion concentration product of (Ca2+) and (CO32) hence obeying second-order kinetics or first-order kinetics if one of the reactants (e.g., calcium) is in excess (Al-Thawadi 2011). Different rate constants have been obtained for bacterial CaCO3 precipitation (Table 1).

Applications of Carbonate Precipitation Through Bacteria The evidence of microbial involvement in carbonate precipitation has subsequently led to the exploration of this process in a variety of fields. A first series of applications is situated in the field of bioremediation. In addition to conventional bioremediation strategies which rely on the biodegradation of organic pollutants (Chaturvedi et al. 2006), the use of MICP has been proposed for the removal of metal ions. Applications include the treatment of groundwater contaminated with heavy metals and radionucleotides, the removal of calcium from wastewater, etc. Another series of applications aims at modifying the properties of soil, i.e., for the enhancement of oil recovery from oil reservoirs and plugging and strengthening of sand columns. Moreover, microbially induced precipitation has been investigated for its potential to improve the durability of construction materials such as limestone and cementitious materials. Bacterial CaCO3 formation through urea hydrolysis is known as bacterial calcite precipitation (BCP). BCP is highly desirable because it is natural and pollutant-free. There are several applications for BCP, most of which are considered for purposes other than strength development (Al-Thawadi 2011). Some of these applications are the (1) removal of contaminants (e.g., radioactive pollutants) and calcium ions from groundwater and wastewaters, (2) protection and restoration of limestone monuments and statuary, (3) creation of sacrificial patinas on limestone and production of biological mortars or cements, (4) plugging the pores of the oil-recover reservoir rock, and (5) stone formation or sand cementation. Below are more details of these applications.

Contaminants Removals from Groundwater and Wastewater The capturing of divalent radionucleotide Strontium 90 (90Sr2+) in the groundwater was investigated after the addition of high concentration of urea and very low concentration of Ca2+ ions (Fujita et al. 2000). Strontium carbonate (90SrCO3) was precipitated, in such a way that (90Sr2+) replaces Ca2+ ions in the calcite crystal preventing the spread of radionucleotide contamination. The potential of removing Ca2+ ions from industrial wastewaters facilitated by BCP instead of chemical precipitation was studied (Hammes et al. 2003). The presence of a high concentration of calcium ions (500–1500 mg.L-1) in the wastewater causes severe scaling in the pipelines and reactors due to calcium formation as carbonate, phosphate, and/or gypsum. By the addition of a low concentration of urea

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(0–16 g.L-1), up to 90% of the calcium ions were removed from the examined wastewater.

Protection and Restoration of Monuments and Creation of Biological Grouts Physical, biological, and chemical factors may cause the weathering of monumental stones. Construction materials, for example, are exposed to Cl, SO42, CO2, and atmospheric moisture which react with the surface layer or penetrate inside the materials (Chunxiang et al. 2009). Consequently, a loss of cohesion of stone material, progressive mineral matrix dissolution, and micro-cracks formation will be enhanced (Chunxiang et al. 2009). In the case of calcareous stones, the porosity will increase due to CaCO3 leaching and weakening of the superficial structure of the stone (Tiano et al. 1999). The attempt which was done by Tiano (1999) in stopping or slowing down the deterioration of monumental statuary by ureolytic bacteria was successful in surface coating but not strength production, as no significant difference in strength was recorded after CaCO3 precipitation. In a recent study, a new layer of CaCO3 was precipitated on the surface of an old concrete layer by S. pasteurii (Al-Thawadi 2011). It was concluded that cracks remediation may enhance the strength and the durability of the concrete. Plugging the Pores of the Rock Bacterial cells were used to plug the highly permeable rocks of the oil reservoir. Between 8% and 30% of the total oil present in oil reservoir was recovered from ordinary oil production method (Al-Thawadi 2011). Oil recovery depends on primary and secondary treatments (water flood) to recover the crude oil in the pores of the reservoir rock (Bryant 1987). To improve the recovery method after primary and secondary treatments, conventional methods depending on chemical or thermal energy are used. These conventional methods are considered inefficient as they led to 67% retention of the total oil within the pores of the reservoir rock (Bryant 1987). Therefore, there was an interest in the use of microbes to enhance the oil recovery. This use of microbes can be through microbial production of bio-surfactants and biopolymers at the surface; microbial growth in the pores of the oil reservoir rocks producing gasses, surfactants, and other chemicals; or microbial plugging of the pores in the oil reservoir channels, which may result in increasing the sweeping effectiveness of the recovery process. The rocks of the oil reservoir contain high permeability zones. When the water is injected to displace oil, it will move through the pores of the highest permeable zone, bypassing much of the oil. Because of the small size of the bacteria, they will move to highly permeable areas, plugging the pores, and as a consequence, the sweep efficiency and oil recovery will be enhanced up to 100% (Bryant 1987). In the ordinary method of bacterial enhancement of oil recovery, plugging of the pores was due to bacterial multiplication, production of gasses that increases the pressure, and production of organic acids, surfactants, and polymer (Al-Thawadi 2011). Much attention is devoted to the plugging of highly permeable zone via bacterial urea hydrolysis. This type of plugging probably offers a feasible alternative

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to block the rock pores, improving the residual oil recovery. Complete plugging within days was achieved by mixing bacteria with sand before packing into cores followed by the application of calcium, urea, and carbonate (Gollapudi et al. 1995). Moreover, it was found that the bacteria plug the sand granules by providing a nucleation site at which CaCO3 was precipitated through alkaline environment (Stocks-Fischer et al. 1999).

Sand Cementation Through Bacterial Carbonate Precipitation Bio-cementation or biogrout is a sand consolidation technology, in which ureolytic bacteria release carbonate from urea hydrolysis in the presence of an excess of calcium ions to form calcite (CaCO3) in situ. Under the proper conditions, this calcite can result in soil solidification and has found significant commercial interest (Al-Thawadi 2008). In geotechnical engineering, the sandstone is produced by filling the sand voids with chemical grouts in a process called chemical cementation or chemical grouting (Al-Thawadi 2011). Chemical cementation depends on chemicals such as sodium silicate, calcium chloride, calcium hydroxide (lime), cement, acrylates, acrylamides, and polyurethanes to bind the sand granules (Al-Thawadi 2011). Construction materials cemented chemically are subjected to weathering which leads to increase the porosity changing the mechanical features of the cemented materials (Tiano et al. 1999). Bio-cementation could be greatly enhanced by using microorganisms with high urease enzyme activities indirectly involved in CaCO3 consolidation (Stocks-Fischer et al. 1999). Besides the high urease activity, a high tolerance to urea, calcium, ammonium, and either nitrate or chloride (depending on the calcium salt used) enhances the bio-cementation (Whiffin et al. 2007). There is a lack of knowledge regarding the high strength production of the bio-cemented products (sandstone formation). Most of the CaCO3 precipitation studies achieved a consolidation or patching treatments for existing material as described previously. Sandstone production depends on how strong the binding between the sand particles, which affects the cementation quality of the precipitated calcite. The growth of CaCO3 crystals for the purposes of artificially cementing sediments proved difficult because of the low yields obtained from a number of different reactions at room temperature. However, the successful bonding of calcareous sediments with derivatives of aluminum alkoide indicates that CaCO3 is a promising route to stabilize loose particles (Al-Thawadi 2011). Superior to this sediment cementation attempt, the loose particles were well cemented by chemical precipitation of CaCO3 through calcite in situ precipitation system (CIPS), producing high degrees of calcite cementation similar to the natural sediments within less than 24 h. This CIPS technology (a non-microbial cementation process) is similar to the natural process that forms the rocks (Joer et al. 2002). This successful rock formation by chemical CaCO3 precipitation indicates a great deal of scope for further work on the strength of microbiological CaCO3 precipitation in a porous medium.

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Nemati and Voordouw (2003) have described the use of urease to cement a porous medium. In this study, reducing the permeability of a porous medium by enzymatic CaCO3 precipitation through Canavalia ensiformis was achieved. Nemati and Voordouw used between 0.1 and 1.0 M (>33 g.L-1) calcite together with high urease activity for a successful plugging of the sand core. Unfortunately, the strength buildup was not monitored. The study of Whiffin (2004) was the first published study in bacterial plugging of loose sandy material through urea hydrolysis (biogrout) for the purpose of strength production. She used P. vulgaris and S. pasteurii during her study. This study was successful in producing strength of 1.8 MPa which was achieved through three applications of bacterial cells and cementation solution. Whiffin (2004) has established a bio-cementation method which depends on a fast flow rate to inject the bio-cementation mix (bacteria, calcium, and urea). The fast flow rate is not recommended to consolidate fine sands. The used cementation solution (calcium/ urea) ranged between 0.75 and 1.5 M. In some cases Whiffin (2004) has recorded a full precipitation of calcite within 18 h. Due to the calcite precipitation, which blocks the pores, a low penetration depth (maximum of 170 mm, i.e., penetration depth is the distance along a packed sand column which can be penetrated by bacterial cells and cementation solution to cement the sand granules) was achieved. Beside this low penetration depth, the inconsistency of urease production by P. vulgaris and S. pasteurii is considered to be another problem which has arisen throughout her study. Al-Thawadi (2008) developed successfully the use of biological cementation to produce high strength comparable to that of the traditional cemented construction materials such as sandstone and concrete with high penetration depth. This study developed a method of producing high strength cemented sandy materials (up to 30 MPa, equivalent to construction cement) through bacterial hydrolysis. This high strength production was achieved without a significant decrease in the permeability. Moreover, high penetration depth up to 1 m was achieved. Al-Thawadi (2008) suggested that the mechanical strength enhancement of cemented sandy materials was caused mostly due to the point-to-point contacts of rhombohedral CaCO3 crystals and adjacent sand grains (Fig. 9). These bridges will bind the sand granules together and increase strength and stiffness (Al-Thawadi 2008).

Inhibition of Coastal Erosion Through Artificial Beachrock Coastal erosion is the removal of beach or dune sediments naturally such as wave action, tidal currents, wave currents, drainage or high winds, and/or man-made activities as tourism, civil constructions, etc. Waves, generated by storms, wind, or fast-moving motor craft, can cause coastal erosion, which may take the form of longterm losses of sediment and rocks or merely the temporary redistribution of coastal sediments; erosion in one location may result in accretion nearby. To preserve coastlines, various countermeasures are used to combat coastal erosion. These include construction of artificial reefs, headlands, detached

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Fig. 9 Light microscopic image for calcite crystals produced by ureolytic bacteria binding two sand particles (Al-Thawadi 2008)

breakwaters, and hard shore protection, all of which control the amount of drift sand and/or beach nourishment and sand bypassing, thus overcoming shortages of drift sand. However, these solutions are expensive and require long-time periods for implementation, as well as the engineering of large amounts of materials, especially for heavyweight sand coasts. The maintenance and management for repair and rebuild the coast are expensive. As a hint for an alternative material in order to reduce life-cycle costs, it is focused on the beachrock. The cement component of beachrocks mainly consists of either calcium carbonate or silica. By utilizing the process of forming the beachrock than to solidify the beach sand, it is considered to be possible to create a highly durable with self-healing capabilities artificial beachrock comparable with existing concrete structures. Compared to the concrete structure, coral sand solidification would be considered to minimize cost. Bio-cementation technology is used to making artificial beachrock. Therefore, it may be possible to slow down the erosion of coasts by making man-made beachrock from coastal sands. Because this artificial rock is made of local materials, it has the potential to be an eco-friendly product (Danjo and Kawasaki 2013a).

Conclusion In the future, it may be possible to manufacture artificial rocks similar to beachrocks for erosion control purposes. Manufacturing of artificial beachrocks using compositional substances are possible as similar to those in beachrocks, and that follow the same formation process. Although bio-cementation technology is a promising technology due to its suitability to field application, it also may results in an environmental problem due to its high production of high concentration of ammonium during urease activity.

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Potential of Biogas Technology in Achieving the Sustainable Developmental Goals: A Review Through Case Study in Rural South Africa

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Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Objectives . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Case Study Area Presentation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Materials and Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Use of Existing Case Studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Questionnaire . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Questionnaire Design . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Results . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

More than half of the Sustainable Development Goals’ (SDGs) target group are rural poor dwellers from developing countries, particularly in Asia and Africa, who normally live on subsistence agriculture. Thus, securing a productive and profitable agricultural sector for the rural areas becomes a fundamental component in accelerating the achievement of the SDGs by 2030. One in eight people worldwide still remain hungry and under poverty. So this paper reviews progress in the achievement of the SDGs in rural South Africa and the potential role appropriate technologies played and/or could play, in particular biogas technology, in the achievement of these SDGs. Questionnaires, interviews, field trips, and community engagement T. E. Rasimphi (*) · W. M. Gitari Department of Ecology and Resource Management, University of Venda, Thohoyandou, South Africa e-mail: [email protected]; [email protected] D. Tinarwo Department of Physics, University of Venda, Thohoyandou, South Africa e-mail: [email protected] © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_40

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were used to establish both the baseline and the impact of biogas use in the improvement of livelihood. Through the use of the biogas technology, agricultural production was improved, and all the chemical fertilizer was replaced by bio-slurry from the digestion process. More economic activities were witnessed through enlargement of the crop farming to make use of the produced slurry raising the need of additional manpower and thus enhancing economic opportunity by creating jobs in the sector. A fair reduction of pressure on local forests was witnessed in the case studied. Thus the use of biogas has impacted positively to the accelerated achievement of some of the SDGs in the communities studied, and this technology presents real solution to the sustainable development of poor communities. Keywords

Sustainable rural development · Renewable energy technologies · Sustainable development · Sustainable development goals · Biogas

Introduction In the absence of new energy policies, the number of people depending on biomass will rise to over 2.6 billion by 2015 and to 2.7 billion by 2030 because of population growth (World Energy Outlook 2016). That is, one-third of the world’s population will still be relying on these fuels. There is evidence that, in areas where local prices have adjusted to recent high-energy prices, the shift to cleaner, more efficient use of energy for cooking has actually slowed and even reversed. According to the report on energy for cooking in developing countries, it was stated that the UN Millennium Project adopted a target of reducing by 50% the number of households using biomass as their primary cooking fuel by 2015 (UN Millennium Project 2005). The appropriate choice of energy source will vary by country, by region, and over time, for example, ethanol gel, plant oils, and biogas. Some communities will prefer the cleaner, more efficient use of biomass energy (World Energy outlook 2016). Large-scale substitution of traditional biomass by alternative fuels will need to take place as well. Meeting the 2015 target would mean 1.3 billion people switching to cleaner alternative energy sources as their primary fuel, while universal access in 2030 would call for 2.7 billion people to switch (IEA 2015). Biogas technology is growing steadily even in countries like South Africa, where there is energy poverty (Amigun and Von Blottnitz 2009). Biogas potential as a renewable energy is magnificent for cooking in rural areas, but currently it is not used as expected. Greben and Oelofse 2009, suggested that as biogas is considered a renewable energy, it can contribute to the South African government’s 10-year goal of promoting renewable energy biogas technology installations which are taking place yet at a very minimal level. There is positive response to the renewable energy sector especially for those who know the benefits of the technology. Poor access to recent energy supplies and unsustainable practices in rural areas of South Africa have been causing environmental destruction and health risk to users (Gender and Sustainable Development Report 2016).

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Over the past quarter century, there has been impressive progress on many fronts in human development, with people living longer, more people rising out of extreme poverty, and fewer people being malnourished (Human Development Report 2016). Human development has enriched human lives – but unfortunately not all to the same extent and even worse, not every life. It is thus not by chance but by choice that world leaders in 2015 committed to a development journey that leaves no one out – a central premise of the 2030 agenda. According to the Human Development Report (2016), there is already tremendous progress in the achievement of the SDGs, and this has been considered the most successful antipoverty push in the history with already the poverty reduction target reached through the Millennium Development Goals, though many areas are still lagging behind. Environmental sustainability is under severe threat, demanding a new level of global cooperation (Ensuring Environmental Sustainability 2005). By large, the poorest people and the poorest countries are the most affected by environmental degradation. Indoor air pollution, caused by smoke from stoves and fires, causes around 1.6 million deaths per year in developing countries (World Health Organization 2002). The problem requires urgent strategies to protect the environment, to achieve economic development, and to increase social transformation, for example, the living environment is improved by eliminating dirty sources of energy, and by-product like ash is eliminated. The smoke is no longer a problem to the biogas user and the living conditions are now improved (Hornberg and Pauli 2007). It is impossible to imagine major progress in reducing poverty without expanding access to modern energy by the poor, who still rely on dangerous and highly polluting traditional fuels, such as cow dung and wood residues. The International Energy Agency estimates that investments of about USD 17 billion per year over 12 years will be needed to provide an additional 500 million people with access to electricity by 2015 (UNDP et al. 2005). Economically, biogas technology comes with financial benefits, for instance, the money which was supposed to be used to buy electricity and other sources of energy is saved (Smith et al. 2013). The economic status of the household is improved. During construction there is skills development when the local residents are taught how to build, install, and use the biodigester. It is also simple to use because less time is spent on using the digester compared to using traditional methods of energy. Thus, a solid and current research based on primary data is needed in order to properly qualify and quantify the potential benefits technically, socially, economically, and environmentally of this technology at the rural household level. For successful implementation of biodigester, there must be availability of organic wastes. Rural are known to generate a lot of organic waste through household activity and farming (UNDP et al. 2005). However, rural areas of developing countries are dependent on primary sources of energy such as firewood and dried dung as sources of energy (O’Neill et al. 2009). Using primary energy sources is often coupled with many environmental problems such as deforestation, land degradation, and various health and social problems as well as emissions of greenhouse gases (Karekezi 2002). Thus the use of biomass as feedstocks to biodigester would reduce fuelwood collection and save money which are some of the benefits of using biogas technology.

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Organic waste-to-energy conversion brings much needed benefits for rural people such as clean heat, smokeless environment in the kitchen, and in some cases even electricity (Amigun and Von Blottnitz 2007). This effectively reduces reliance on the grid electricity whose cost is unsustainable for rural communities. On an economic level, biogas technology makes a contribution toward national economic development in that some employment will be created during construction, installation, and maintenance of digesters. Moreover income generated improves the people’s standards of living, hence the eradication of historical economic imbalances.

Objectives To demonstrate the potential of biogas technology in achieving Sustainable Development Goals. Main indicators that have been used include crop production, women involvement, and environmental sustainability.

Case Study Area Presentation The study focused on assessing the potential of biogas in achieving sustainable development goals in Vhembe District.

Materials and Methods Use of Existing Case Studies Although the use of extrapolation and benefit transfer models, especially with regard to stated preference results, is generally not favored (Pearce and Ozdemiroglu 2002), desk research is easy and quick. In the case of this research, where time and resources are not without limits, it appears to be acceptable to make provision for extrapolation and use of aggregate regional data when site-specific details are not available (European Commission 2002). Information was gathered from journals, research studies, library database, and the study sites. This involved case studies on biogas technology utilization in Vleifontein. Census 2011 rated Limpopo Province as the poorest in the country with a head count poverty level of about 63.8% (Lehlola 2014) and yet with one of the best agricultural climate and vast natural resources.

Questionnaire In order to conduct a comprehensive feasibility study of the potential benefits of rural household biodigester use (Smith et al. 2013), it was necessary to first bring information about the current characteristic and energy demands of rural household in the study area. For this purpose a questionnaire was designed to gather detailed

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information about the household, their energy requirement, their livestock keeping practice, water usage, and the production of crops and vegetables at their homesteads. The questionnaire design, survey, and analysis process is discussed in the preceding section.

Questionnaire Design The questionnaire used in this research was based on and adapted from previously used questionnaire: a questionnaire used in a biomass energy audit conducted by Rhodes University and a biogas perception and behavior questionnaire conducted by Agama Energy and Jabezi in the development of the South African National Biogas feasibility study (Austin and Blignaunt 2008; Smith 2013). The questionnaire was used as a basis for this research project as they had been tried and tested in the field and had contributed to save meaningful studies (Smith 2013).

Results Theoretically any type of biomass can be degraded to biogas; the growth of biogas technology has caused the use of cow dung, chicken manure, and organic waste to be used. Cow dung is especially suitable because of the methanogens in the stomachs of the ruminants (Bond and Templeton 2011). The dung produced by cattle per day per animal is about 10 kg, and 1 kg of cattle is not pastured at the cooperative; it was assumed that collectable waste per kg per day is about 10 kg, and 1 kg of cattle produces 0.0036 m3 of biogas (Nijaguma 2002). Since the cattle are not pastured at the cooperative, it was assumed that collectable waste per kg per day was 5 kg. In addition, 1 kg of chicken manure wastes produce 0.062 m3 (Mukumba et al. 2013). The installed digester reduced the dependence of the cooperative from fuelwood and LPG which they used for cooking. Marubini Women’s Multi-purpose Co-operative is a women’s’ cooperative in Maila Village, Makhado in the northernmost province of Limpopo in South Africa. Census 2011 rated Limpopo Province as the poorest in the country with a head count poverty level of about 63.8% (Lehlola 2014); it has the best climate for agriculture and vast natural resources. The cooperative was established in 1995 as a nonprofit organization by unemployed women and was engaged in mushroom production. The group was then encouraged to register as a cooperative, a development organization registered with the Department of Health and Social Development as a nonprofit organization. The cooperative has a membership of 12 (permanent cooperative members) and 10 volunteers. Its main activities are early childhood development with 250 children at crèche level. Besides running this crèche, the cooperative also do poultry production at a capacity of around 3200 egg layers at a time throughout the year (Tinarwo et al. 2014). The cooperative is also in cash crop production. The place is now a well-known site with a lot of visitors coming for different reasons ranging from business, buying the produce, to even learning purposes.

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The cooperative uses a lot of energy in the form of heat to cook food for children. They get firewood supplies from local entrepreneurs who charge exorbitant prizes. On the other hand, they produce a lot of waste from the egg layers which can be used to feed the biodigesters, and the waste poses disposed challenge in addition to the health issues. The cooperative is very satisfied with the benefits demonstrated by including the bioenergy system in their farming activities. Although it is still too early to make definite assumptions on all benefits, the cooperative already is saving R925 (has been realized from cooking energy only). On the agricultural part, the yields have improved significantly in both quantity and in quality as a result of the use of bio-slurry as fertilizer. The unpleasant odor from chicken wastes has significantly been reduced with the use of biodigester. After realizing the great potential of production of energy from biogas, with more waste use, the cooperative already is working toward expanding the number of egg layers quantity and increasing the crop farming sector now that fertilizer is available (Tinarwo et al. 2014). This cooperative is now used by the university project team to disseminate information to other stakeholders and to others users. Field days are regular at the project site with day-long workshops carried out on site giving other great opportunities of catering services to participants. The place has turned to be a center of attraction, receiving all kinds and levels of visitors with even local and provincial government leadership visiting to see a working self-sustaining integrated agricultural bioenergy system. Biogas technology played an important role in improving the quality of life for the cooperative since it was introduced. In addition to being a renewable source of energy that is locally available, biogas has an extensive range of additional benefits to people. Some of these benefits include: (a) More hygienic conditions, thanks to the elimination of indoor air pollution from fuelwood, which is particularly harmful to women who are responsible for cooking. (b) Preventing deforestation by reducing the detrimental harvesting of trees as fuelwood. Maila cooperative used fuelwood for cooking food for the children, and the introduction of biogas has lessened the reliance on wood as energy source. (c) Burning biogas is cleaner than burning fuelwood. Apart from being smokeless, the only emissions are carbon dioxide and water; thus carbon emission to the atmosphere is reduced. (d) Macroeconomic benefits such as job creation thanks to the staffing needs of biogas plant construction. The cooperative enjoys the benefits from their sales of agricultural practices like cabbage, pumpkin, and spinach (see Fig. 1).

Biogas technology offers an exceptional set of benefits (Fig. 1). It improves the health of users, it is a sustainable source of energy, it benefits the environment, and it provides a way to treat and reuse various wastes: animal and agricultural (Tinarwo et al. 2014).

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Fig. 1 Potential of biogas technology to improve household energy users lives (Tinarwo et al. 2014)

Whilst there is a common agreement that rural communities urgently need to increase their access to current energy sources (World Bank 2000), fuelwood use has an impact on the social and economic status of rural community as it is considered the primary energy source, and all cooking activities depend on fuelwood. Majority of rural households in South Africa are still dependent on firewood and paraffin (Shackleton et al. 2007). This is evident from the questionnaire survey conducted in the study area that large percentages still depend on fuelwood for cooking and heating. The acquisition of fuelwood is a major cost in a resource-poor household, be it in terms of time and energy. Although this study did undertake an analysis on time constraint for collection of fuelwood from the nearby forests. Other studies have been adopted to highlight time collection for fuelwood in rural areas and the benefits of biogas technology to drastically reduce time for fuelwood collection. A study by Smith et al. (2013) at Okhombe community, a typical rural community to the rural Maila, quantified the total benefits of the value of time, wherein the users of biodigester at household level require the users to spend time feeding the system, but it also considerably reduces the need for rural people to collect wood and saves time. The total time saving due to cooking with biogas is a combination of time saved on wood activities, cooking with traditional solid fuels, and cleaning cooking utensils (Smith et al. 2013). Considering the total workload of women, biogas as a new technology does not affect traditional working patterns. Cooking with solid fuels on open fires or with traditional stoves results in high levels of air pollution exposure due to pollutants such as small particles and carbon monoxide, and this affects the health of users.

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Many health benefits can result from moving from traditional energy sources to cleaner fuels. According to the World Health Organization (WHO), over three billion people worldwide continue to use solid fuels, including wood, dung, agricultural residues, and coal, to supply their energy needs (WHO 2011). Indoor air pollution, a significant proportion generated from traditional cooking stoves, is thought to be responsible for 2.7% of the total global burden of disease (WHO 2011). It was assumed that only households that cooked with solid fuels would benefit from a reduction in indoor air pollution (IAP), and thus the calculation of IAP-related deaths was based on the number of households that used fuelwood, cow dung, or coal as their primary and/or secondary cooking fuel. It was assumed that the use of biogas will reduce by 65% the IAP, and consequently IAP-related deaths will be reduced by the use of biogas in place of traditional solid fuels. The benefit of a biodigester (and the output of bio-slurry) to farming practice was quantified as the avoided cost of purchased fertilizers. A well-designed biodigester system, with the inclusion of rainwater harvesting systems and fodder cultivation, has the ability to provide benefits well beyond the mere replacement of purchased fertilizers. A project in which biodigester household members are trained to use harvested water and bio-slurry to grow food and fodder for their cattle has significant potential to improve the health of both cattle and people as well as instigating improved cattle management practice which will, in turn, contribute to the preservation of natural resources. Further study on this topic is required, but it is believed that a suitable fodder feeding program will benefit the health of cattle greatly and reduce grazing pressures on the land – resulting in reduced erosion. These outcomes have great potential in terms of food supply and resource management for sustainable development. The increased economic benefit inherent in these outcomes is likely to further increase the societal desirability of the technology. The bio-slurry from the digesters at Maila cooperative is used to water the feeds for cattle. Two types of grasses are being grown at the site, and they are watered by the bio-slurry which is applied in order to demonstrate the usefulness’ of the bio-slurry to replace the fertilizers. Evidence from the monitoring process and expression by the biogas users at the cooperative for cooking suggest that biogas if used to its full capacity can provide enormous benefits. The cooperative members revealed that their lives have dramatically changed ever since the introduction of the biogas digesters. They also have reported that their monthly expenses on cooking energy sources have been reduced dramatically since they no longer have to depend on fuelwood and LPG but rather biogas. In addition to saving on fuelwood purchases, they identified that they no longer needed to constantly manage cooking fires and that they were saving time as they were simply able to turn the gas on, light it, and begin cooking without the need to watch over the process. Timesaving, as a result of more efficient cooking practices and reduced fuelwood purchases and collection, reveals that a biodigester could contribute to promoting gender equality and empowerment of women as women are relieved of time-consuming domestic duties, while efficient cooking fuel and reduced wood harvesting also have the potential to ensure environmental sustainability.

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Biogas technology provides social and economic benefits. It can improve the health of users, is a sustainable source of energy, benefits the environment, and provides a way to treat and reuse various wastes – human, animal, agricultural, and municipal. The findings of the research strongly point out the significance of biogas technology in this time to achieve energy sustainability and to solve the problem of indoor air pollution in rural household, and to reduce the household energy cost spent on energy sources for cooking and heating thus achieving the 2030 UN Sustainable Development Goals.

Conclusions In conclusion, the use of biogas technology as a clean and renewable energy to solve the escalating energy problems in rural areas has been demonstrated in the rural area wherein many benefits where experienced. However, there are alternatives to the use of traditional biomass such as biogas technology, which can provide clean and reliable energy services. The result is a better quality of life socioeconomically, better health, and improved academic performance by children being able to study for longer hours.

References Amigun B. and Von Blottnitz H (2007) Investigation of scale economies for African biogas installations Amigun B, Von Blottnitz H (2009) Capital cost prediction for biogas installation in Africa. Environ Program Sustain Energ 28:134–142 Austin G, Blignaunt J (2008) South African national rural domestic biogas feasibility assessment: prepared for ministry for development co-operation, The State of Netherlands. Agama Energy, Cape Town Bond T, Templeton M (2011) History and future of domestic biogas plants in the developing world. Energ Sustain Dev 15:347–354 Ensuring Environmental Sustainability: Measuring Progress (2005) Toward the 7th millennium development goal. World Bank European Commission (2002) Regional foresight in the cohesion countries: experiences, concepts and lessons learned. European Commision Research Director General, Brussels Gender and Sustainable Development (2008) Maximising the economic, social and environmental role of women. OECD REPORT Gender and Sustainable Development Report (2016) Maximising the economic, social and environmental role of women. OECD Greben H, Oelofse S (2009) Unlocking the resource potential of organic waste: a South African perspective. CSIR- Natural Resource and the Environment, Waste Management Research, Pretoria Hornberg C, Pauli A (2007) Child poverty and environmental justice. Int J Hyg Environ Health 210:571–580 Human Development Report (2016) Human development for everyone, The 2016 Human Development Report is the latest in the series of global human development reports published by the United Nations Development Programme (UNDP) By the United Nations Development Programme 1 UN Plaza, New York

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IEA (2015) Energy for Cooking in developing Countries. OECD/IEA. www.iea.org Karekezi S (2002) Renewables in Africa: meeting the energy needs for the poor. Energ Policy 30:1059–1069 Lehlola P (2014) Poverty level of South Africa. Statistic South Africa Mukumba P, Makaka G, Mamphweli S (2013) A possible design and justification for a biogas plant, Journal of Energy in Southern Africa, Vol 24:4 Nijaguna BT (2002) Biogas Technology, 1st edition, New Age International. ISBN 10: 8122413803, 46–48 O’Neill M, Kinney PL, Cohen AJ (2009) Environmental equity in air quality management: local and international implications for human health and climate change. J Toxic Environ Health A Curr Issues 71:570–577 Pearce D, Ozdemiroglu E (2002) Economic valuation with stated preference techniques. Renwick, Department of Tranport, Local Government and the Regions, London Shackleton CM, Buiten E, Annecke W, Banks DA, Bester J, Everson T, Fabricius C, Ham C, Kees M, Modise M, Phago M, Prasad G, Smit W, Twine W, Underwood M, Von Maltitz G, Wenzel P (2007) Exploring the options for fuelwood policies to support poverty alleviation policies: evolving policy dimensions in South Africa. Forests, Trees & Livelihoods 17: 269–292 Smith M (2013) Financial and economic feasiblity of biodigester at Okhombe community. Masters Research Smith M, Goebel J, Blignaut J (2013) The economic and financial feasibility of rural household biodigesters for poor communities in South Africa. Waste Manag 34(2):352–362 Tinarwo D, Mugera WM, Rasimphi TE, Nekhubvi V, Machaba H (2014) Assessment of biogas technology, a case study of a rural village in Vhembe District, Limpopo, South Africa, Paper presented at a conference in Nairobi, Kenya UN Millennium Project (2005) Investing in development: a practical plan to achieve the millennium development goals World Bank (2000) Energy and Development Report 2000: Energy services for the World’s poor, World Bank, Washington, DC WHO (World Health Organization) (2002) The World Health Report 2002: Reducing risks, promoting healthy life. Geneva, Switzerland World Health Organization (2011) Progress on the Health-Related Mellinium Development Goals. www.who.int/whosis/whostat/2011/en World Energy Outlook (2016) Secure sustainable together. OECD/IEA International Energy Agency. www.iea.org

A Review on Treatment of Pharmaceuticals and Personal Care Products (PPCPs) in Water and Wastewater

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Mukesh Goel and Ashutosh Das

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Pharmaceuticals and Personal Care Products (PPCPs) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Toxicity and Evaluation of Risks Associated to PPCPs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Treatment of PPCPs Present in the Water and Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cross-References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

Water is the central element of all vital social and economic processes. Because of the development of consumer society, harmful chemicals are being generated in huge quantities throughout the world. The problems derived from the toxicological effects of these organic compounds must be resolved for the benefits of the entire society. The problem is certainly complex, and it is imperative that novel procedures are required to deal with this extensive range of tribulations. Though there are plenty of water treatment methods, a significant number of chemicals are still present in “clean” water. These chemicals are referred to as emerging contaminants (ECs). Some emerging contaminants have been used for a long time but have only recently been discovered in lakes, rivers, and groundwater – our drinking water sources. Emerging contaminants are of many types and have many sources. One large group consists of the pharmaceuticals and personal care products (PPCPs) that we purchase and use regularly. Other categories include certain pesticides, nanomaterials, flame retardants, and plasticizers. Major sources of these chemicals include residential, agricultural, and industrial activities. This chapter reviews the various treatment technologies presently available for PPCP removal. M. Goel (*) · A. Das Centre for Environmental Engg., PRIST Deemed University, Thanjavur, Tamil Nadu, India © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_41

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Keywords

Emerging contaminant (ECs) · Pharmaceuticals and personal care products (PPCPs) · Treatment · Water · Wastewater

Introduction Water is the central element of all vital social and economic processes. Because of the development of consumer society, harmful chemicals are being generated in huge quantities throughout the world. The problems derived from the toxicological effects of these organic compounds must be resolved for the benefits of entire society. The problem is certainly complex, and it is imperative that novel procedures are required to deal with this extensive range of tribulations. Though there are plenty of water treatment methods, a significant number of chemicals are still present in “clean” water. These chemicals are referred to as emerging contaminants (ECs) (Barcelo and Petrovic 2007; Kumerer 2009; Kyle and Murray 2010). Some emerging contaminants have been used for a long time but have only recently been discovered in lakes, rivers, and groundwater – our drinking water sources. In the last decades, rapid progress in detection methods has made it conceivable to measure even a small concentration of ECs. Drinking water samples in various places have reported the presence of these contaminants (Parrot and Bennie 2009). Emerging contaminants are of many types and have many sources. One large group consists of the pharmaceuticals and personal care products (PPCPs) that we purchase and use regularly. Other categories include certain pesticides, nanomaterials, flame retardants, and plasticizers. Major sources of these chemicals include residential, agricultural, and industrial activities (Parrot and Bennie 2009; Phillips et al. 2010). Primarily, they also enter through wastewater treatment plants. These are hospital wastes, unaffected by conventional treatment processes.

Pharmaceuticals and Personal Care Products (PPCPs) The greatest danger from the presence of pharmaceuticals in the water is not so much the single type of pharmaceuticals, but it is when different chemicals are consumed together and produce a magnified effect which is known as synergism. On top of this, scientists are not aware of the effects of these combined pharmaceuticals. PPCPs include prescription and non-prescription medications, nutritional supplements, diagnostic agents, as well as other consumer products such as disinfectants, fragrances, sunscreen, and cosmetics. Even though most of them are polar, have a short half-lifetime in water, and are found in trace concentrations, they are considered pseudo-persistent pollutants. This is because their universal, frequent usage by multitude of individual and animals causes a continuous discharge into the environment oftentimes sourcing from either untreated or insufficient wastewater treatment.

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Some PPCPs are easily broken down and processed by the human body or degraded quickly in the environment, but others are not easily broken down and processed, so they enter domestic sewers. Because they dissolve easily and do not evaporate at normal temperatures or pressure, PPCPs make their way into the soil and into aquatic environments via sewage, treated sewage sludge (biosolids), and irrigation with reclaimed water (Xu and Wu 2009; Zhou et al. 2009). An important source of PPCPs in groundwater is from municipal landfills through leachate. There are thousands of chemicals that are constituents of PPCPs. These are diverse and are used as active ingredients or preservatives in cosmetics, skin care, dental care, hair care products, soaps, cleansers, insect repellents, sunscreen agents, fragrances, and flame retardants. Pharmaceuticals and personal care products were first called “PPCPs” only a few years ago, but these bioactive chemicals (substances that have an effect on living tissue) have been around for decades. Their effect on the environment is now recognized as an important area of research (Lindberg et al. 2007; Scheurer et al. 2011).

Toxicity and Evaluation of Risks Associated to PPCPs The toxicological problems posed by these PPCPs have created a great concern among scientists. Several studies are being conducted to ascertain the incumbent damage in relation to existence of PPCPs and other emerging organic contaminants on aquatic organisms and other components of the ecosystem. Though their concentrations are very low (ng/L to μg/L) and may not cause an immediate risk to human health, their presence seems to create an environment of anxiety. It is still debatable whether they are actually harming the environment or not, in the present intake. Most probably, the recycle and accumulation in the wastewater in time to come will become serious environmental issues. It has been observed that aquatic organisms such as fish would be exposed to PPCPs during the course of their life cycle. The contact of fish with these chemicals resulted in changes in sperm density, gonad size, male sex reversal, and other health issues (Kolpin et al. 2002). Similarly, Brain et al. (2004) reported the effect of atorvastatin, a lipid regulator on aquatic macrophytes at 26 ppb. The synthetic hormone 17-ethinylestradiol has been shown to cause feminizing effects in fish at very low concentrations (Parrott and Blunt 2005). The transformation products of PPCPs have not been explored adequately, though there are a few reports. Gomez et al. (2012) noted transformation products of PPCPs at the Henares river basin in Spain at a very high concentration. Interestingly the concentration of these intermediates was higher than those of parent compound discharged from wastewater treatment plant. Likewise, Quintana et al. (2005) reported several alteration products of ibuprofen and other pharmaceuticals. There are some alarming incidents such as reduction of the population of vultures in India and Pakistan due to the use of diclofenac on cattle in these countries. There are many other dangerous omens found all over the world (Ong et al. 2009).

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Treatment of PPCPs Present in the Water and Wastewater The methods that are normally employed to treat the water are aimed at removing and reducing overall groups of contaminants to a safe level. However, PPCP treatments are not encouraging, as even the so-called clean water seems to have appreciable concentration of these chemicals. Thus, it is important to develop effective technologies to remove these compounds from drinking water. Various treatment methods presently used for removing PPCPs are described below: Advanced oxidation processes (AOPs): Conventional oxidation processes use oxidants like Cl2 ozone (O3), H2O2, and so on. Some of the oxidants are common in drinking water treatment, but they suffer from various disadvantages, such as production of trihalomethanes (THMs), haloacetic acids (HAAs), and other harmful by-products. The advanced oxidation processes (AOPs) generate free radicals, which act as strong oxidant to destroy the organic pollutants. AOPs are defined as the process which entails the generation of highly reactive radicals (especially hydroxyl radicals) in ample quantity to affect water purification. Rate constants are usually in the range of 108–1011 L.mol1 s1. Hydroxyl radicals (OH) react with almost all organics and are stable over a wide range of pH, which makes them highly attractive for wastewater treatment. AOP includes several methods such as H2O2/UV, O3/UV, and UV/TiO2. Being a destructive technology, it has significant benefits for the wastewater treatment. AOPs have the distinction of being stand-alone, or it can also be joined with other physicochemical or biological processes. Process coupling is conceptually beneficial usually leading to improved treatment efficiencies (Renugoat et al. 2012; Mohapotra et al. 2013). Advanced oxidation techniques like ultrasonication (Nie et al. 2014), advanced Fenton, UV/H2O2 (Deng et al. 2013), advanced photocatalytic oxidation (Bae et al. 2013), and TiO2 photocatalytic ozonation (Márquez et al. 2014) have been reported. Advantage of AOPs can be summed up as follows: • It does not transfer the pollutants from one phase to another. It completely mineralizes them. • Operation is done at mild conditions of temperature and pressure. • Reaction rate is very high. However, it has certain disadvantages, primary among them are: • Lack of sound knowledge of process design and operation of large-scale reactors. • These processes are comparatively costly. The cost includes the cost of equipment and the intense energy required to run the process. Coagulation–flocculation: Coagulation is used for removing particulate matter, colloids, as well as some dissolved substances. Coagulation–flocculation processes have so far yielded ineffective elimination of most micropollutants.

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Matamoros and Salvadó (2013) reported the treatment of micropollutant in a coagulation–flocculation chamber for treating secondary effluent. Fifty percent removal was observed for galaxolide, tonalide, and octylphenol. Suárez et al. (2009) reported significant reduction (around 80%) of musks during coagulation–flocculation treatment of hospital wastewater. Carbella et al. (2010) conducted lab-based experiments to observe the effect of coagulant loading, coagulant identity, time of mixing, and temperature on PPCP degradation. Only 20% removal was observed. Ozonation: Ozonation, though part of AOPs, stands out because of its versatile use in wastewater treatments. Ozone has been used for removing micropollutant load of full-scale WWTPs (Synder et al. 2006). Similarly Zimmemann et al. (2011) examined the efficacy of ozonation for removing micropollutants. The removal was on higher side. Studies have shown that ozone can reduce the concentrations of many PPCPs in wastewater (Lee et al. 2012; Rosal et al. 2010). Ozonation of PPCPs resulted in complete degradation of macrolides, estrogens, and sulfonamides. Oher chemicals, such as diclofenac and naproxen, were also transformed (Huber et al. 2005). Removal of sulfonamides using ozone is extensively studied by Garoma et al. (2010). Sulfonamides are highly used in the USA and can be excreted by the body at high rates. They tested the sulfonamide degradation for varying pH, ozone concentration, and bicarbonate ion concentration. Higher pH and higher ozone concentration facilitated the faster degradation of sulfonamides and intermediates. Similarly Von Gunten (2003) observed the effect of nitrite and ammonia on ozonation of EE2 (17α-ethinylestradiol). Bromate formation through ozonation of bromide-containing waters is one of the shortcomings in ozonation. Bromate is a potential hazard. Similarly, Andreozzi et al. (2001) report ozonation of carbamazepine (CBZ), an antiepileptic drug. Nitrite and nitrate positively affected the ozonation of CBZ and increased the speed of CBZ transformation. On the other hand, humic acid was detrimental in CBZ degradation and reduced the efficacy of the process. Once again, the expensive ozone and its lack of application in treatment plants handicap versatile ozonation. Photocatalysis: Heterogeneous photocatalysis with titanium dioxide is reported to be effective and useful for the management of non-biodegradable wastewaters containing recalcitrant compounds (Gouvea et al. 2000; Yeber et al. 2000). Heterogeneous photocatalysis can take place in all the fluid media such as gas phase, pure organic liquid phases, or aqueous solutions. The classical heterogeneous catalysis comprises five steps. Frist step is the transfer of reactants from fluid phase, which is followed by adsorption of a minimum of one reactants, and then reaction in the adsorbed phase. The reaction is succeeded by desorption of the product and finally removal of the products from the interfacial region (Herrmann 1999). The photocatalytic reaction takes place in the adsorbed phase. It is similar to typical catalysis process, difference being photonic activation replaces the thermal activation of catalyst. The adsorption of photons results from the illumination of semiconductor catalyst with photons, provided energy of photons is larger than the conductors’ bandgap energy. This process leads to the formation of electron-hole pairs, which further divides into free photoelectrons in the conduction band and holes in the valence band.

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The hole can either oxidize a compound directly or react with electron donors like water to form OH radicals, which in turn react with pollutants such as chlorophenols and dyes. TiO2 is generally preferred as the catalyst because of its stability, good performance, and low cost (Andreozzi et al. 2000). The disadvantage is of the catalyst separation from solution, as well as the fouling of the catalyst by the organic matter. Apart from TiO2, various oxides and sulfides have been used as catalysts: ZnO, CeO2, CdS, ZnS, etc. (Herrmann 1999). The following reactions take place during the photocatalytic degradation of organics using TiO2: TiO2 þ hv ! e þ hþ

(1)

O 2 þ e ! O  2

(2)

H2 O ! OH þ Hþ

(3)

OH þ hþ ! OH •

(4)

þ O 2 þ H ! HOO •

(5)

2HOO • ! O2 þ H2 O2

(6)

H2 O2 ! 2OH •

(7)

R  H þ • OH ! R • þ H2 O

(8)

R • þ • OH ! R  OH

(9)

R  OH þ • OH ! Intermediates ! CO2

(10)

Heterogeneous photocatalysis has been studied in detail. PPCP degradation can take place by direct photolysis (Klavarioti et al. 2009). Gomes et al. (2017) reported the photocatalytic ozonation of paraben using TiO2 doped with noble metals such as Ag, Au, Pt, and Pd. They observed that combining photocatalysis with ozone led to reduced usage of ozone thus making the process economical and more feasible. 0.5% Ag-TiO2 was found to be the best catalyst resulting in total paraben removal. However, it was also noted that toxic by-products generated from photocatalytic ozonation of paraben were not completely mineralized. Constructed wetland: Constructed wetland (CWs) is a wastewater treatment system that has the potential of reducing pollutants to levels that will meet existing standards for water quality before discharging to the environment. CWs are probable contenders to remove different kinds of pollutants such as organics, suspended solids, pathogens, nutrients, and heavy metals from wastewater (Antoniasdis et al. 2007). There are two basic kinds of constructed wetlands, horizontal flow (HF) and vertical flow (VF). CW was found to be effective for reservoir effluent, and removal efficiencies were reported to be 90.4% for TSS and 62.4% for COD (Avsar et al. 2007). Huett et al. (2005) described the application of HF-constructed wetland for

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nitrogen and phosphorus removal. They found more than 96% removal with Phragmites australis, but unplanted wetlands were unproductive and produced only 16% nitrogen removal. Calheiros et al. (2007) reported similar results. They accomplished effluents with nitrogen concentration of less than 1 ppm and phosphorous with 0.05 ppm. Aslam et al. (2007) studied the performance of VF-constructed wetland for the treatment of refinery wastewater. They used Phragmites karka and were successful in the reduction of TSS, COD, and BOD. They also compared the compost- and gravel-based wetland and observed the superior performance of compost-filled wetland. Matamoros et al. (2005) reported the removal of ibuprofen (IB) using HF and VF. Matamoros et al. (2007) found a prevalence of OH-IB over Ca-IB metabolite in a HF wetland, which would be explained by its predominant aerobic biodegradation. A study by Heberer (2002) showed that persistent chemical like diclofenac was successfully degraded using constructed wetlands. Matamoros and Bayona (2006) stated lower removal of diclofenac, while Avila et al. (2010) observed approximately 99% degradation using HF. IB was also successfully tested by Hijosa-Valsero et al. (2011), where they observed lower degradation in winter but more than 90% removal in summer season. There are also reports on the removal of ethynilestradiol, musk fragrance and tonalide, triclosan, etc. using constructed wetland (Park et al. 2009; Matamoros et al. 2007). Song et al. (2009) did a comparative study on saturated and unsaturated VF for ethynilestradiol degradation and observed that unsaturated VF produced high efficiency of removal. Membrane filtration: Membrane technology, particularly nanofiltration (NF) and reverse osmosis (RO), is another option for augmenting PPCP removal from wastewaters. The choice of an appropriate membrane process depends on the scope of removed effluents and admixtures present in water. They can be used to remove effluents as independent processes or be combined with unit complementary processes, forming a treatment process line. Increasingly, NF/RO membranes have been successfully applied in removing micropollutants from the aquatic environment (Yang et al. 2011; Serrano et al. 2011; Sahar et al. 2011; Chon et al. 2012). Adsorption: In the last decades, several studies were carried out to remove emerging pollutants by adsorption process. Commercial adsorbents such as natural clays, minerals, and activated carbons have been used (Yoon et al. 2005; Snyder et al. 2007). Redding et al. reported the treatment of a lake water spiked with 29 EDCs and PPCPs with concentration values of 100–200 ng/L. There are other successful studies by Sotelo et al. (2012). Modified adsorbents have also been successfully employed in the treatment of PPCPs containing water. Zhuo et al. (2017) described the adsorption of PPCPs (benzoic acid (BEN), ibuprofen (IBU), and ketoprofen (KET)) onto the two composite beads of porous metal–organic frameworks (MOFs) containing chitosan and sodium alginate. It was observed that MOFs containing chitosan reported higher adsorption capacity. There are other reports enumerating the adsorption capacity of chitosan, especially nano chitosan. It has drawn particular attention as effective nanosorbent due to its low cost and its high contents of amino and hydroxyl functional groups showing high adsorption

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potential for various aquatic pollutants (Crini and Badot 2008; Chatterjee et al. 2009; Guibal et al. 2006). Apart from that, there is another category of adsorbents, magnetic nanoparticles, which are reported to have high affinity toward organics. Besides, these can be easily separated from the process stream owing to their magnetic nature (Nassar 2010; Ruzmanova et al. 2013; Ochando-Pulido et al. 2013). Magnetic resins were also found to be effective adsorbents for water treatment and PPCP removal (Zhu et al. 2016). They compared the adsorption behavior of this resin with two activated carbons, namely, Norit and F400D. Norit was faster in removing PPCPs compared to magnetic resin. However, humic acid could not affect the adsorption performance of resin, while both the activated carbons were affected. Besides, resin displayed enhanced regeneration potential.

Conclusions We are living in the world of water crisis, thanks to our unchecked industrial pollution and population explosion leading to massive diversification of freshwater for agriculture and urban needs. Moreover, global climate change variation, which has resulted in repeated tragedies in India, will certainly impact water issues. The dry regions of North Africa, Asia, Europe, etc. have to deal with water shortages in the coming decades. The industrialized countries like the USA have its own issues with the water, thanks to massive consumption in the industry as well as in domestic sector. The USA has faced the most severe drought in the last decades. Similarly, China is confronted with the fact that their biggest rivers are exorbitantly polluted, which cannot be even used for irrigation. All these countries now have to deal with additional sets of problems unheard of in the last decades. The policy makers in several countries are in great anxiety and are devising various strategies to treat these malign chemicals. The search for water treatment, however, should not lead to uneconomical and impractical solutions causing more problems in the future. This chapter presented a detailed review on PPCP removal from water and wastewater. Novel technologies are required to deal with this menace, conventional technologies will fail, and they can produce more harmful by-products. Nanocomposites are likely nano-adsorbents that can remove PPCPs from waters. They have diversified applications in different areas such as biological sciences, drug delivery systems, and wastewater treatment. Modified nanocomposites such as magnetic nanocomposites are much more efficient class of nanocomposites in which magnetic nanoparticles are dispersed in a magnetic or nonmagnetic insulating matrix. The shape, size, and distribution of the magnetic particles play an important role in determining the properties of such materials. Magnetic nanocomposites are proved to be better than nanocomposites as they have the advantages of both magnetic separation techniques and nano-sized materials, which can be easily recovered or manipulated with an external magnetic field. They are also very effective for the removal of both organic and inorganic pollutants from the pollutant water.

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Cross-References ▶ Advancement of photocatalytic water treatment technology for environmental control ▶ Biostimulation and Bioaugmentation: An alternative Strategy for Bioremediation of ground water contaminated mixed landfill leachate and sea water in low income ASEAN countries ▶ Biotechnological Approach for Mitigation Studies of Effluents of Automobile Industries ▶ Constructed Wetland: A Green Approach to Handle Wastewater ▶ Micro-remediation of Metals: A New Frontier in Bioremediation ▶ Nanomembranes for Environment ▶ Techniques for Remediation of Paper and Pulp Mill Effluents: Processes and Constraints

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Constructed Wetland: A Green Approach to Handle Wastewater

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Ashutosh Das and Mukesh Goel

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Types of Constructed Wetlands and Their Suitability . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Types of Wetland Plants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Transformation Mechanisms of Major Nutrients and Metals in Wetlands . . . . . . . . . . . . . . . . . . . . Nitrogen Transformations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Phosphorus Cycling . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Nutrient Uptake by Wetland Vegetation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sulfur Cycling . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Pathogen Reduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Water Temperature . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Standard Wetland Systems for Treatment of Specific Wastewater Types . . . . . . . . . . . . . . . . . . . . . Dairy Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Landfill Leachate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Oil and Grease-Rich Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Pulp and Paper Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acid Mine Drainage . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Winery Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Limitations of Wetland Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Operation and Maintenance of Constructed Wetlands . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Operation and Maintenance of FWS Wetlands . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Operation and Maintenance of VSB Wetlands . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cost of Wetland Treatment Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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A. Das · M. Goel (*) Centre for Environmental Engg., PRIST Deemed University, Thanjavur, Tamil Nadu, India © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_42

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Abstract

Natural wetlands are able to manage all local contamination by naturally occurring agents such as sunlight, aeration, biotic interaction, microbial action, precipitation, and gravitational attraction. Contaminants are removed from the system through volatilization, sedimentation, infiltration, dilution, and oxidation, among others. Constructed wetlands aim to simulate a natural wetland system through optimal configurations of wetland plant types and density, waste characteristics, flow rate, soil types, and stratification. In this way, wastewater can be effectively treated in a natural way, without heavy machinery and additional chemicals that may increase pollution. This chapter discusses the different types of constructed wetlands, the types of wetland plants, and the transformation mechanisms of major nutrients and metals in wetlands, along with suitability and limitations. Finally, the operational requirements, maintenance, and cost of a wetland system are described. Keywords

Constructed wetland · Free water surface (FWS) · Vegetated submerged bed (VSB) · Nutrient transformation

Introduction Although natural wetlands have always been used as sinks for wastewater, engineered wetland systems were only conceptualized in the past century to simulate a natural wetland’s characteristics (Wallace and Robert 2006). Phenomenal growth in constructed wetland technology has occurred all over the world especially, because of their ability to recycle, transfer, and immobilize a wide range of potential contaminants (Vymazal 2005, 2010; Greenway 2007; Kadlec and Wallace 2009). In fact, constructed wetlands are especially suitable for small communities in developing countries due to their adaptability to locally available materials, lack of dependency on electric power, lack of requirements for skilled manpower in operation and maintenance, and non-dependency on centralized waste-treatment facilities (Denny 1997; Cooper and Findlater 2013).

Types of Constructed Wetlands and Their Suitability Three constructed wetland systems are discussed in the literature: hydrobotanical systems (Pawęska and Kuczewski 2013), aquaculture systems (Brix 1999), and soil systems (Blazejewski and Murat-Blazejewska 1997). Of these, soil systems are most studied because of their abundance and simplicity (Cooper and Findlater 2013). Small-scale constructed wetlands essentially comprise four types: free water surface (FWS), vegetated submerged bed (VSB), vertical flow (VF), and sludge dewatering reed bed (SDR) (Wallace and Robert 2006). FWS wetlands have areas of open water, floating vegetation, and emergent vegetation, thus providing a habitat for a variety of wildlife, similar to natural

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wetlands (Knight 1993; Kadlec and Knight 1996). These wetlands are most suited to warmer climates because of their higher biological decomposition rates. FWS wetlands are generally used for treating effluent from secondary treatment processes (e.g., trickling filters, oxidation ponds/lagoons, activated sludge units) rather than directly as a secondary treatment process because of the potential for human exposure to pathogens. Based on soil conditions, berms, dikes, and liners may be used to control flow and infiltration in such wetlands. Essentially, basic wetland treatment includes the processes of sedimentation, filtration, oxidation, reduction, adsorption, and precipitation (U.S. EPA 2000). In VSB wetlands (also known as subsurface-flow wetlands), the wastewater stays below the ground’s surface. Thus, unlike in an FWS wetland, the risk of human exposure to pathogens is minimal, as is the breeding of mosquitoes and other vector organisms. Generally, VSB wetlands consist of a lined gravel or soil-based bed that is planted with emergent vegetation. The wastewater is treated as it travels through the gravel media and root zones (especially rhizomes). VSB wetlands are low-cost and low-maintenance systems; they are most suitable for single-family homes or small cluster systems. VF wetlands, as the name indicates, allow effluents to be filtered vertically, similar to sand filters. Wastewater travels across the surface of a sand or gravel bed (which is planted with wetland vegetation) through the plant root zone. This process yields nitrified effluent through the oxidation of ammonia—something that is not possible in VSB wetlands. These systems are used in combination with VSB wetlands in many European countries, resulting in a nitrification-denitrification treatment process (Cooper et al. 1999). SDRs (reed beds) were developed by Seidel in the early 1960s. This system is essentially an enclosed basin with a sand layer underlain by drainage pipes. The evapotranspiration capacity of emergent wetland plants (typically Phragmites) dewater sewage sludge. These wetlands are less common than the other three types.

Types of Wetland Plants All wetlands depend on plants to shade the water column, act as substratum for biofilm growth, and cycle nutrients and organic carbon. However, FWS wetlands are more dependent on plants than are VSB wetlands. Although several species of plants can be sustained in wetlands, only a few species are widely used because of their tolerance to high nutrient levels, saturated soil, and anaerobic conditions, as well as their ability to propagate through rhizome spread and their local availability (Wallace and Robert 2006). When selecting plant types for a wetland system, the following parameters should be evaluated: • Hydrology: The water level at peak and normal flow can be used to determine variability in thickness, frequency and duration of an unsaturated environment, and inundation.

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• Climate and latitude: It is preferable to choose plant types that are able to survive in the local climate. • Cost and availability of plant stock: It is preferable to choose native species rather than exotic species. • Plant size: A larger root biomass is preferable for a higher rate of survival and prevention of weeds. • Rate of colonization: The spreading rate of the plant species dictates the requirements for the initial planting stock. Balanced spreading is desired; fast spreading reduces weed invasion, whereas slow spreading reduces species diversity. • Water quality: In particular, variations in salinity, pH, alkalinity, and oxygen demand should be within the tolerance levels of selected plants. • Objectives: A wildlife habitat would require a predetermined threshold diversity level, whereas wastewater treatment is better accomplished by a simple planting scheme with less diversity. • Maintenance: Especially during the first year, maintenance should include removing invasive vegetation, maintaining water levels, removing plant detritus deposits, ensuring good oxygen transfer, and preventing mosquitoes (Wallace and Robert 2006). For FWS Wetlands (Fig. 1), emergent wetland plants are preferred. Sedimentation can be increased by reducing wind-induced mixing and resuspension, The additional surface area can increase biofilm growth and the uptake of soluble pollutants for a range of treatment mechanisms. Particle interception is also increased. Shade is provided from the plant canopy over the water column, thus reducing algae growth.

Fig. 1 Schematic of FSW and VSB wetlands

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Furthermore, flocculation of smaller colloidal particles into larger, settleable particles is induced (Sinclair Knight Merz 2000). Ideal plants include perennial and fastgrowing species with high lignin content to provide a year-round aboveground structure, including cattails (Typha spp.), bulrushes (Scirpus spp.), and common reeds (Phragmitesaustralis). VSB wetlands (Fig. 1) are less dependent on plants. However, they still require careful planting, preferably with plants that have significant root penetration into the bed media. This provides more surface area for biofilm growth, chemical exudates that detoxify the root environment, and the introduction of fungi species and symbiotic bacteria. Studies on VSB wetlands have not provided significant evidence on the relationship between plant types and treatments (Gersberg et al. 1984; DeBusk et al. 1989; Van Oostrom and Cooper 1990; Batchelor et al. 1990; Knight 1993). However, plants that are easy to propagate and are able to survive in relatively hostile environments are suitable for a VSB wetland, including mostly Phragmites (common reed) and Phalarisarundinacea (reed canary grass; Vymazal 1998). Plant selection for VF wetlands and sludge dewatering beds (reed beds) have not been explored much in the literature because they are used less often than FWS and VSB wetlands. Wildlife grazing may need to be considered for the protection of wetland plants. Whitetail deer (Odocoileusvirginianus), for example, threaten winter plantation in VSB wetlands in North America, which is often countered by spring plantation. Similarly, Canadian geese (Brantacanadensis) can affect FWS wetlands by obliterating the planting stocks. Proactive methods of animal control (e.g., screening, deterrents) are often employed to deal with these situations.

Transformation Mechanisms of Major Nutrients and Metals in Wetlands The migration and transformation of nutrients and metals in a wetland are affected by a wide range of factors, including settling (or vertical stratification), dispersion, preferential flow, loading characteristics (hydraulic and organic), and basic hydrological variants such as precipitation, evapotranspiration, infiltration, runoff, and the overall water balance. The most common transformation phenomena are discussed in this section (Wallace and Robert 2006).

Nitrogen Transformations The redox conditions in wetlands are affected by oxygen transfer and organic matter loadings (both internal and external). Nitrogen can exist in various forms in wetlands, such as ammonia, nitrite, nitrate, organic, and gaseous nitrogen. Mineralization (also known as ammonification) is the process of converting organic nitrogen into inorganic nitrogen (especially ammonia) in either aerobic or anaerobic conditions; the process occurs the quickest in oxygenated zones. Most of

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the nitrogen present in domestic wastewater ends up as organic nitrogen or ammonia. The mineralization rate in wetlands depends on the temperature (doubling with a temperature increase of 10  C), pH (optimum pH is between 6.5 and 8.5), carbon-tonitrogen ratio of the residue, available nutrients in the system, and characteristics of the wetland substrate (namely, texture and structure) (Reddy and Patrick 1984; Reddy et al. 1979). Ammonia volatilization involves ammonium-nitrogen (equilibrated between gaseous NH4+ and aqueous NH3) escaping from soils and sediments in wetlands; the levels are significantly higher when the pH is greater than 8.0 (Reddy and Patrick 1984). The most conducive environments for this process are submerged vegetation and algal photosynthesis (through depletion of carbon dioxide and bicarbonate ions from the water column). Ammonia volatilization is affected by the NH4+ concentration in water, temperature, wind velocity, solar radiation, type and number of aquatic plants present in the system, and the capacity of the system to change the pH during diurnal cycles (Wetzel 2001; DeBusk et al. 2004; Vymazal 1995). Nitrogen reduction (denitrification) refers to the reduction of oxidized nitrogen compounds (nitrate or nitrite) to nitrogen gases (N2 and N2O), which vent from the water column. This phenomenon is caused by a variety of bacteria and some fungi under both aerobic and anaerobic conditions. Denitrification is able to proceed soon after the onset of anoxic conditions (a redox potential of approximately 220 mV) (Hauck 1984). The denitrification rate depends on the biochemical oxygen demand (BOD) of wastewater (which is the carbon source required for the process), temperature, type of available carbon, and carbon-to-nitrogen ratio in the system (Ingersoll and Baker 1998; Baker 1998; Hume et al. 2002a,b). Nitrogen transformation in FWS wetlands is essentially guided by the amount of dissolved oxygen available in the wetland system, where the decay of plant detritus often serves as the main source of this organic carbon. The relative magnitude of these mechanisms changes both daily and seasonally (Liehr et al. 2000; Gerke et al. 2001; Shipin et al. 2004). Here, in the root rhizosphere, bacteria compete for organic oxidation, nitrogenous complexes, and sulfides. Most VSB wetlands are primarily used for treatment of high BOD and anaerobic effluents. Thus, mineralization that leads to the formation of ammonia occurs here more often than in FWS wetlands due to their lower oxygen availability. Hence, the influent pH of the water and the bed media is the factor affecting nitrogen transformation (Johns et al. 1998).

Phosphorus Cycling Because of the high phosphorus and nitrogen loadings in wastewater, constructed wetlands are often significantly richer in nutrients than are natural wetlands. In FWS wetlands, phosphorus is mostly cycled sequentially: net plant uptake, sorption, and sedimentation/degradation (accretion through direct settling and trapping) of plant biomass through growth, death, and decay (Kadlec and Knight 1996; Craft and Richardson 1993; Kadlec 1994). The rate of phosphorus removal is determined primarily by the phosphorus loading rate, size of the wetland, prevailing climatic

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conditions, and type of vegetation (Vymazal 2001; Kadlec 1999). In VSB wetlands, the plant components quickly reach an equilibrium with the influent phosphorus, leading to poor removal of phosphorous unless special bed media are used.

Nutrient Uptake by Wetland Vegetation The available nutrient content in the tissues of wetland plants and their growth rates determine their nutrient uptake and storage potential, which is greater in constructed wetlands than natural wetlands (Greenway 2003). Interestingly, most of the nutrients incorporated into plant tissue during growth can return to the water column through decomposition if the wetland is not harvested (Reddy and DeBusk 1987). The amount of nitrogen and phosphorus that can be removed by harvesting is generally insignificant compared to applied wastewater loads; it does not exceed 10% of the total removed nutrients (both nitrogen and phosphorus), even in well-developed constructed wetlands with emergent macrophytes (Brix 1994; Vymazal 2001; Gersberg 1986; Herskowitz 1986; Vymazal et al. 1999).

Sulfur Cycling Sulfate is probably the most common anion in surface waters. Its formation through oxidation indicates aerobic conditions, whereas its reduction to sulfide indicates anaerobic conditions (similar to ammonia). In VSB wetlands, sulfide oxidation to sulfate occurs in parallel to and competes with nitrification, which is also associated with heavy metal precipitation (especially copper and nickel) (Eger 1992; Eger and Lapakko 1989; Frostman 1993).

Pathogen Reduction The four common pathogens found in domestic wastewater are bacteria, viruses, protozoa, and helminths, which often sustain the regular treatment train. Several pathogens can be removed through wetland treatment, including fecal streptococci, salmonellae, Yersinia, Pseudomonas, and Clostridium bacteria species, through physical (e.g., filtration, ultraviolet disinfection, sedimentation), chemical (e.g., oxidation, wetland generated-biocidal disinfection, adsorption into organosorbents), and biological (e.g., predation, natural death) pathways (Herskowitz 1986; Gersberg et al. 1989). In FWS wetlands, all of these pathogen removal processes have been demonstrated with ultraviolet radiation exposure in proportion to open-water areas (Mara 1976). In Fish & Wildlife Service (FWS) wetlands with high populations of birds and warm-blooded animals, higher levels of certain pathogen occur from these non-human sources, including fecal contamination, streptococci, and salmonella (PBSJ 1989). Although VSB wetlands also offer robust pathogen removal (more than 90%), few studies have investigated the removal mechanisms. In these systems,

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sedimentation, filtration, and interception seem to be common removal processes, especially for helminth ova; adsorption and natural death are used for bacteria and viruses (Gerba et al. 1999; Thurston et al. 2001; Gersberg et al. 1989; Mandi et al. 1998; Stott et al. 2002). Pathogen removal efficiency has been reported to be higher when the fineness of bed materials and retention time are increased (García et al. 2003a,b).

Water Temperature Thermal energy gains (through solar radiation and influent water temperature) and losses (through evapotranspiration and effluent flow temperature) are associated with diurnal and seasonal cycles, which cause short- and long-term water temperature variations, respectively. Additionally, ground-heat storage transfer forms a two-way transfer mechanism, from the wetland to the ground during summer (because of high solar radiation) and from ground to the wetland during winter (because of low solar radiation). A similar two-way mechanism has also been observed for convective energy transfer, wherein the wetland receives energy from the air or releases heat energy back to air, depending on the difference between the air temperature and water temperature. When the energy inputs and outputs are balanced on seasonal basis, the water temperature is said to be a balance point temperature, which is usually observed in wetland-inlet zones. Depending on the degree of vegetative cover, climate, and solar intensity, FWS wetlands typically show temperature variations on both daily and seasonal bases, ranging from 5  C to 15  C (Kadlec and Knight 1996; Kadlec and Reddy 2001). In VSB wetland systems, the temperature response of the wetland is related to the air temperature in tropical and equatorial regions; in temperate regions, leaf litter on the top of the system has been observed to play a role in retaining and transferring heat (Kadlec and Knight 1996; Kadlec 2001a).

Standard Wetland Systems for Treatment of Specific Wastewater Types As discussed previously, two standard wetland systems are used extensively around the world: free water surface and vegetated submerged bed wetland systems. FWS wetlands are the earliest engineered wetland designs. VSB constructed wetland systems were introduced more recently; hence, the technology is continuing to develop. The main characteristics of these two wetland systems are presented in Table 1. Domestic and municipal wastewater can be efficiently treated by small-scale constructed wetland systems. They also can treat several other types of wastewater, either remedially or industrially. The following sections discuss some of the wetland treatment systems used in a variety of special applications.

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Table 1 Comparative analysis of free water surface and vegetated submerged bed wetland systems Salient features Structure of wetland

Influent wastewater type

Size and cost

Effect of treatment mechanisms (either simultaneously or sequentially)

Use in cold climates

Wildlife habitat and aesthetic appeal

Free water surface (FWS) Emergent wetland plants, open water, and a lined/unlined basin. Soil forms the rooting media, above which a water layer is maintained. This leads to constant anaerobic soil conditions, with breeding of mosquitoes and other vectors Used for treated wastewater (not usually raw domestic waste) to avoid exposure to pathogens and vectors Suitable for large size with high wastewater flow (more than 75,000 m3/d) with less chance for hydraulic failure and lower cost on per-hectare basis Physical treatment involves gravitational settling, diffusion, and volatilization, leading to reduction of dissolved and particulate contaminants. Chemical treatment involves absorption, ultraviolet radiation, and chemical precipitation, leading to removal of phosphorus and dissolved metals. Biological treatment involves transforming organic matter and nitrogenous constituents. Here, submerged plant shoots in addition to suspended-growth systems such as lagoons provide an increased surface area for microbial growth Water temperature affects the rate at which biological processes occur. Effluent water temperature closely mimics air temperature during non-ice periods, whereas ice formation limits functionality during icy periods More suited

Vegetated submerged bed (VSB) Gravel or soil-based beds are planted with wetland vegetation. Wastewater is not exposed during treatment because it is kept within the bed media and hence not provide habitat for mosquitoes or other vector organisms Most commonly used to treat primary domestic wastewater to secondary effluent discharge standards Well suited for treating very small wastewater flows (less than 75 m3/d) and higher cost on per-hectare basis Less understood with regard to physical and chemical treatment mechanisms and effects. In relation to biological treatment, VSB is better than FWP because of its significantly larger available surface area for microbial growth, thus improving efficiency of the system. The increased microbial density allows for a much smaller system footprint compared to FWS systems

Unlike FWS, the normal accumulation of plant detritus may not provide enough insulation to keep the system from freezing. Hence, insulating materials (e.g., hay, straw, wood, mulch) are used for insulation Less suited

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Dairy Wastewater Dairy wastewater is a mixture of two distinct wastewater types: wastewater from cleaning and sterilization/pasteurization and wastewater from the washing of cowshed in various proportions at different dairies. Although FWS and VSB wetlands have been used for the treatment of dairy wastewater, the former is less effective than the latter wetland system because of insufficient oxygen transfer and hence anaerobic conditions (especially because of the high effluent concentration of ammonia nitrogen) (Newman et al. 2000; Schaafsma et al. 2000). However, the milk proteins in this wastewater degrade much slower than in domestic wastewaters, which needs to be considered in the system design. Thus, designers should use this approach with caution. Vertical-flow wetland processes with high rates of oxygen transfer have been found to be successful in treating dairy wastewater (Green et al. 2002; Liénard et al. 2002).

Landfill Leachate The chemical characteristics, concentration, and volume of leachate depend on the source waste material disposed of at a particular landfill site, its stage of decomposition, the lining of landfill cells (or lack thereof), and the rate of groundwater flow through the waste material. The basic decomposition of municipal solid waste occurs in the following five phases: 1. Hydrolysis and acidification: This phase of aerobic decomposition lasts less than a month or so—after the cell is covered until the exhaustion of available oxygen. 2. Acid fermentation: This phase of anaerobic decomposition lasts for years, even decades, with the production of organic acids (namely acetic, proprionic, and butyric acids). It is characterized by a BOD-to-COD (chemical oxygen demand) ratio of more than 0.7, a pH level of 5–6, strong odors, ammonia concentration greater than 500 mg/L and dissolution of heavy metals such as iron, manganese, zinc, calcium, and magnesium. 3. Methanogenesis. This phase, in which the organic acid products of acid fermentation produce methane, takes many years or even decades. The end product has relatively low BOD values and BOD-to-COD ratios, yet it has high ammonia, continuing metal dissolution and precipitation, and circumneutral pH. 4. Progressive stabilization: This phase is characterized by a slowdown of bacterial decomposition with exhaustion of organic material. 5. Final storage: In this phase, decomposition is complete (McBean and Rovers 1999). Activated sludge treatment and aerated lagoons have been effective in treating landfill leachate. Constructed wetlands are most effectively used only as a polishing step after lagoons for improving treatment efficiency. Constructed wetlands are unable to treat the high concentration of ammonia present in the leachate due to insufficient oxygen transfer rates. Thus, aerobic pretreatment of the leachate is

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necessary prior to its entry into a wetland system, which results in a multi-stage treatment process (Schwarzenbeck et al. 2003; Birbeck et al. 1990; Mæhlum et al. 2002; DeBusk 1999). Active wetland treatment systems have been successfully used to treat leachate directly, but only with a mechanism to boost oxygen transfer in order to oxidize the high concentrations of ammonia commonly found in landfill leachates (Vasel et al. 2004). FWS wetlands can remove the high ammonia in leachate; however, they are handicapped by lower nitrate reduction. Thus, a better wetland option is an aerated VSB wetland (Kadlec 1999; Liehr et al. 2000; Urbanc-Bercic 1994; Kinsley et al. 2002).

Oil and Grease-Rich Wastewater Wetland vegetation is very effective at removing oil and grease in stormwater (domestic or industrial). In fact, studies have reported the successful removal of BP diesel-range organics (in the C-10 to C-26 range) through horizontal-flow vegetated subsurface-flow wetland cells in VSB systems; the removal was found to be in proportion to the vertical distribution of roots and rhizomes in the wetland cell (Wass and Fox 1993; Omari et al. 2003). Similarly, petroleum-containing wastewaters are rich in volatile organic compounds and are mostly released from loading facilities or containment areas (with relatively high concentrations during rain and relatively low concentrations during the pumping of contaminated groundwater). This wastewater was reported to undergo natural degradation in both FWS and VSB constructed wetlands through a microbial community associated with the plant rhizosphere (Wemple and Hendricks 2000; Schnoor et al. 1995; Pardue et al. 2000; American Petroleum Institute 1998). FWS wetlands can be effectively coupled with mechanical treatment systems when there are higher concentrations of petroleum hydrocarbons in wastewater (Lakatos 2000). VSB wetlands have been found to have greater efficiency and robustness than FWS wetlands because of their higher biological treatment per unit area (Kadlec and Knight 1996). In fact, the installation of adequate aeration has been found to enhance both volatilization and hydrocarbon degradation (Kadlec 2001b).

Pulp and Paper Wastewater The preparation of wood pulp requires a high-temperature environment to dissolve the lignin in raw wood and thus isolate the cellulose fibers, using either a strongly alkaline sodium sulfide solution (Kraft process) or a mildly acidic bisulfate salt solution (sulfite process). Furthermore, bleaching may occur using chlorine, chlorine dioxide, sodium hydroxide, oxygen, or hydrogen peroxide to manufacture a bright-white product (Thut 1993). The dark-brown color-forming compounds of pulp wastewater (called refractory organics) are mostly resistant to microbial degradation, with residual BOD (approximately 10–100 mg/L after reduction from 200–800 mg/L).

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Maryville Mill, a paper manufacturer in Victoria, Australia used aerated lagoons and activated sludge basins in pioneering research on lignosulphonate, color, Total suspended solid (TSS), BOD, and foaming tendency. The mill used wetland plants in a partially-vegetated facultative treatment pond to treat pulp and paper mill wastewaters with five emergent wetland plant species in FWS basins, resulting in significant pollutant removal (Allender 1984). In fact, increased retention time was also found to result in higher removal efficiencies of TSS, BOD, NH4-N, org-N, and phosphorous (Thut 1993). Constructed wetlands have the advantages of avoiding residual solid management, enhancing wildlife habitats, and allowing for limited public use of the facilities.

Acid Mine Drainage Acid mine drainage contains high concentrations of dissolved metals from the excavation and exposure of sulfide minerals, such as pyrite (FeS2) and Fe+2. This drainage often enters aquifers and raises the pH (to between 4 and 7) through buffering by mineral dissolution. Ferric oxy-hydroxide precipitates may form, thus lowering the pH again to less than 3 (Benner et al. 1999; Brodie et al. 1993). This low pH is the best condition for a dissolution of a wide array of metals available in the vicinity, such as nickel, lead, manganese, copper, aluminum, and zinc. FWS wetlands are commonly used to oxidize iron and manganese to insoluble precipitates, which are two major components of acid mine drainage (Kleinman and Girts 1987; Mays and Edwards 2001; Weider 1989; U.S. EPA 1993). FWS systems are more effective than VSB systems at oxidizing iron and manganese due to their higher oxygen transfer rates; furthermore, they have far greater storage capacity for accumulated iron and manganese sludge because of their larger dimensions. However, poorly designed FWS wetland hydraulics can adversely affect the performance through short-circuiting. Some heavy metals (e.g., copper, nickel, lead, zinc, and arsenic) that are often present in mine drainage waters are also soluble at the low pH of acid mine drainage; furthermore, they precipitate at high pH independently or co-precipitate with iron in constructed wetlands (Morin et al. 2003; Fitch and Burken 2003; Benner et al. 1999). During the conversion of sulfate to sulfide, which is typical in VSB wetlands with anaerobic conditions and a continuous supply of organic matter, these metals undergo precipitation. They are often collected using adsorbent beds of spent mushroom compost, manure, rice hulls, sawdust, and other similar materials (Palmer et al. 1988; Fitch and Burken 2003).

Winery Wastewater Winery wastewater is often richer than domestic wastewater, with lower pathogen levels and variable loading over time (being the highest during the grape-processing period) (Grismer et al. 2003). This wastewater possess high levels of suspended

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solids from grape stems, seeds, and skins. Hence, there is a need for pre-treatment through screens or clarifiers, as well as associated facultative ponds and sand filters (Shepherd and Grismer 1997; Grismer et al. 2003). A typical VSB wetland is best suited for treating winery wastewater (Grismer et al. 2003).

Limitations of Wetland Systems Although constructed wetlands can provide low-cost and low-maintenance biological treatment of wastewaters, the major drawback of these systems is their variability in treatment efficiency. Much of the variability can be attributed to inadequate knowledge of the system dynamics. Furthermore, the dominant physical, chemical and biological processes have not yet been optimized (Cooper and Findlater 2013). Constructed wetlands are very complex systems with variable interplay of both external factors (e.g., flow rate, wastewater composition, temperature) and internal factors (e.g., bacteria growth and development). Hence, these systems are mostly regarded as “black-box” models, which can be simulated using only input and output data and incorporate few degradation models (Rousseau et al. 2004; Wynn and Liehr 2001; Samsó and Garcia 2013). At present, no direct observational methodologies are available to assess the systems; efforts in this direction would involve sophisticated and expensive instrumentations (Mburu et al. 2013). FWS wetlands are a known source of carbon dioxide (CO2), methane (CH4), and nitrous oxide (N2O). Hence, they are potential contributors to global warming (although their global warming potential is widely variable, poorly quantified due to data scarcity, and poorly quantifiable as well) (Bartlett and Harriss 1993; Bridgham et al. 2006; Gleason et al. 2009; Sha et al. 2011; Brinson and Eckles 2011). However, compared with natural wetlands, constructed wetlands seem to have less soil organic matter, shorter hydroperiods, more aerobic soil conditions, and hence lower fluxes of greenhouse gases (Bruland and Richardson 2005; Knutsen and Euliss 2001). In fact, some authors have argued that the water quality and biodiversity benefits obtained from constructed wetlands may outweigh their potential negative impacts due to greenhouse gas emissions (Hopple and Craft 2013; Marton et al. 2014).

Operation and Maintenance of Constructed Wetlands A constructed wetland can be indigenously constructed using local materials. It involves minimal usage of mechanical equipment (as well as the accompanying risks of failure and malfunction). The constructed wetland is often scalable for smaller units and sustainable with variable flow rates and loading conditions. Thus, it can produce a consistent effluent quality, without constant tweaking and adjustments. Overall, a constructed wetland is an attractive option for decentralized wastewater management (U.S. EPA 1997; Wallace and Robert 2006). In comparison to mechanical

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treatment works, wetland systems have very few operational controls because they rely largely on passive treatment mechanisms.

Operation and Maintenance of FWS Wetlands Control of the Water Level In an FWS wetland, adjustment of the water level plays a critical role. The growth of wetland plants must be accommodated during the plant establishment phase. Furthermore, the depth should be increased during winter operations, decreased to dry out and kill accumulated mats of filamentous algae, and decreased to minimize odors. Hence, operation and maintenance of FWS wetlands primarily involves adjusting the water level in the wetland cell (up or down) and varying the loading between treatment cells (if the system has two or more parallel treatment trains) through adjustment of hydraulic control structures. Additionally, water leakage should be prevented to ensure the water level; thus, a watertight control structure is recommended, such as an adjustable weir gate system or manufactured stacking plate system for large- or small-scale systems, respectively. For plant establishment and maintenance, water level control is also essential at all three phases of the FWS wetland, as follows: • Phase 1: Low water level and exposure of the rooting media (mud flat) is required for the initial (re)planting. • Phase 2: A gradual raise of the water should be synchronized with the emergence of plants to allow the emergent shoots to be above the water surface at all times, thus providing the plants with access to sunlight and oxygen. The required degree of adjustment to the water level is inversely related to organic/nutrient loading. (The maximum should be less than 2 cm/day.) • Phase 3: The water level is maintained at the design depth upon establishment of the wetland plants.

Winter Operations The water level should be raised during winter months (especially in temperate climate) to partially enhance the season-induced low biological activities. In colder periods, the raised water may form an ice layer. In this case, the water level should be lowered below the ice layer to create an insulative air layer below the ice layer, thus providing warmth to the wetland. A rapid increase of the water level when an ice cover has already formed should be avoided to prevent the plants from becoming loose at the rooting media. An early spring rain may naturally create such a condition; hence, frequent checks are crucial until the ice layer has melted. Algae Control Filamentous algae mats can displace emergent wetland plants in FWS systems, especially with high nutrient loadings and poor establishment of emergent plants. This may reduce the dissolved oxygen and sunlight below the algal mat, with

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subsequent deterioration of the wetland plants. In such a case, the water level can be lowered to expose the rooting media and cause removal of the algal mat through oxidation. This should be followed by fresh replantation and increases to the water level along with the growth of emergent plants.

Odor Control FWS systems have an anaerobic zone at the bottom of the water column, which may release hydrogen sulfide or other odiferous compounds. The odor may increase with higher organic loading of the influent due to improper or only primary treatment, turbulence, or splashing at the inlet. Streamlining the inlet flow, improving pre-treatment before entry to the wetland, and making the wetland depth shallower (allowing better oxygen mixing) can help to reduce the odor. Mosquito Control A mosquito population decreases exponentially with increasing distance from the breeding ground (although they can travel up to 5 km from the breeding ground) (Service 1993). FWS wetlands have dense vegetation, heavy accumulation of plant detritus (as with most emergent wetland vegetation, such as cattail, bulrush, and common reed), and shallow berms, which provide ideal conditions for mosquito production; however, these are also the ideal conditions for wetland efficiency (Knight et al. 2004; Sinclair Knight Merz 2000). Efforts to control mosquitoes through deeper habitats, steeper berms and more open water adversely affect the life of the emergent plants and decrease the biofilm surface area (plant detritus)—and thus decrease wetland efficiency. Therefore, mosquito control and treatment efficiency have to be a tradeoff (Russell 1999; Martin and Eldridge 1989; Kadlec 2003; Knight et al. 2004). Because egg-laying by adult mosquitoes is difficult to stop, favorable conditions should be created for mosquito larva predators—such as certain aquatic insects and select fish (e.g., Gambusia holbrooki, top minnows)—to access the entire wetland with no depositional plugging, especially at the inlet. This is an effective method of mosquito control that does not compromise wetland efficiency. However, in cold climates and with less dissolved oxygen in wastewater, most of these predators cannot survive, although mosquito larva can (Greenway and Chapman 2002; Stowell et al. 1985). Two bacteria strains—namely, Bacillus thuringiensis israelensis and Bacillus sphaericus—can be applied to wetlands (either in pellet or granular form) for significant control of mosquito production (Williams et al. 1996; Knight et al. 2004). Vegetation Management The removal of excess plant detritus can help with weed and mosquito control; however, this disrupts wetland operations and creates additional problems with regard to their disposal and the prevention of further regeneration, thereby leading to no long-term benefits (Williams et al. 1996; Knight et al. 2004; Thullen et al. 2002). The controlled burning of excess wetland biomass (and even weeds and invading vegetation growing in the wetland or its berms) seems to be the most

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effective means of vegetation control. This approach allows for the reintroduction of nutrients (in the form of ash) into the water column, although it does temporarily affect the efficiency of the wetland (Kadlec and Knight 1996; Knight et al. 2004).

Management of Animals and People Some herbivores, such as muskrats (Ondratazibethicus) and nutria (Myocastor coypus), breed rapidly and consume emergent vegetation, such as cattails (Typha spp.). These animals may also use the vegetation as building material for feeding platforms and nest mounds. They can even destroy the entire wetland vegetation, removing any ability for reestablishment through their grazing pressure and thereby converting the wetland into a TSS and BOD-rich algal pond. Animals may also damage the berms and dykes by burrowing. Wire mesh (chicken wire) or armoring berms with rip-rap have been used widely to manage these animals (Pineywoods 2001). Similarly, beavers are attracted to flowing water for habitation. They often construct dams and thus create plugging of outlet controls or open water runs, which lead to hydraulic short-circuiting and a reduction in treatment efficiency. Trapping, chain link fencing around the wetland, and exclusionary controls such as rodent guards should be used. Common carp (Cyprinuscarpio) can also be a nuisance species. They can uproot vegetation and disturb sediments, thereby affecting treatment performance. Thus, these fish should be excluded from wetlands (Hey et al. 1994). Even seemingly innocent birds as such as waterfowl, especially the Canadian goose (Branta canadensis), and many resident bird populations may consume newly planted wetland vegetation, thereby destroying the wetland. This problem should be tackled by netting and deterring these birds, which can also contribute to elevated fecal contamination levels in the wetland effluent. FWS wetlands are often encroached upon by birdwatchers and outdoor enthusiasts who are exploring the habitats. This adds a further responsibility to provide safe and well-lit public access with adequate parking. Specific measures such as fencing, prohibiting children and domestic pets from entering the area, and the provision of signboards may also be required in providing risk-free access for people while protecting the ecology of the environment (Kadlec and Knight 1996).

Operation and Maintenance of VSB Wetlands Control of the Water Level Similar to FWS wetlands, water level adjustments in a VSB wetland (or between splits for parallel cells, if any) are crucial, especially in those using gravel as bed media. The wetlands should have no leaks, with the water level below the soil yet in contact with the root zone, with the vegetation having poor drought resistance. Because the water in this system is not visible, its operation requires more intuition and experience than does an FWS wetland. For plant establishment and maintenance, water level control is also essential in all three phases of a VSB wetland, as follows:

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• Phase 1: The water level should be maintained at a fixed level before plantation to ensure that it is in contact with the rootstock during the plant establishment; this helps to avoid exposure of the rootstock to the sun, which leads to wilting. Furthermore, it should be ensured that vegetation is not unable to access the water on its own due to the poor capillary action of gravel. As the plants grow, fluctuations in the water depth need to be managed to enhance root penetration into the bed media. • Phase 2: The water level should be lowered gradually to the design’s water level after the establishment of wetland vegetation. Lowering the water level incrementally assists with deeper root penetration into the VSB bed. • Phase 3: The water level should be kept in contact with plants, yet below the soil level, for plants that are already established.

Odor Control A VSB wetland system often involves only primary treated effluent (mainly from a septic tank). Thus, the water level needs to be minimized to prevent the odors resulting from anaerobic decomposition, except if a VSB system is covered with insulating and odor-arresting mulch layers. Winter Operations In cold but not freezing climates, no insulation is necessary. However, temporary or permanent insulation of hay, straw, or leaf litter, among others, can prevent a VSB wetland from freezing. If the insulation materials are degraded, however, they may break down and thereby adversely affect treatment. The water depth should be decreased during the winter to create an insulating air gap. The most critical point in a VSB system is the outlet water control structure, as it is most susceptible to freezing; thus, this point needs to be monitored to avoid clogging, insulated, and (in severe freezing conditions) heated through an electrically powered source (e.g., incandescent light bulb, stock tank heater) (Kadlec and Knight 1996; Wallace et al. 2001). Bed Clogging Clogging is a major constraint of a VSB system as it obstructs the flow. Thus, the operator should carry out a series of troubleshooting steps to assess the cause, as there is a lack of visual cues (unlike FWS wetlands).The troubleshooting steps should involve the following: • Checking the outlet valve for scum clogging • Checking the surge adequacy for surfacing (which may be confused with clogging) • Checking the daily flow rate and proper dimensioning to prevent flooding • Checking the adequacy of organic loading across the cross-sectional area and influent suspended solids levels of the wetland, which can lead to precipitation and clogging • Checking the influent iron levels (especially for landfill leachate, which are normally rich in iron), as this causes oxidation and precipitation

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• Checking the system hydraulics • Observing deliberate fluctuation (over a day or longer) and corresponding response of the wetland A basic strategy for addressing a VSB clogging problem includes the following: • Replacing the connecting septic tank baffles (if non-functional or missing) so that they are adequate for handling the pumping rate • Reducing surges at the inlet end of the VSB by reducing the pumping rates and dose volumes (including pretreatment of influents if they are rich in BOD, suspended solids, or iron) • Replacing clogged VSB media with new (preferably coarser) bed media • Adjusting the influent distribution header to create a higher cross-sectional area (at the header/media interface)

Undesirable Plant Species A layer of mulch and/or plant detritus on top of a VSB wetland may lead to the invasion of undesirable plants into the wetland system. To solve this problem, complete coverage of the desired plant species should be maintained, with no space for invaders. Any invaders can be controlled locally by hand weeding and spot applications of herbicides. Some common invaders of VSB systems that receive full sun include cottonwood (Populus deltoids), giant ragweed (Ambrosia trifida), Canada thistle (Cirsium arvense), and marijuana (Cannabis sativa). Stinging nettle (Urtica dioica) and jewelweed (Impatiens capensis) commonly invade VSB systems that receive partial sun.

Cost of Wetland Treatment Systems Wetland systems generally use local labor and local materials. Hence, a standardized cost estimation cannot be applied to all treatment systems. The cost of the basic components of a wetland treatment system, such as earthwork, gravel (for VSB wetlands), and plants, are distance sensitive. Labor costs are also widely variable. However, in determining the economic feasibility of a wetland treatment system, local cost figures (involving capital and operating costs) can be used to compare various treatment technologies.

Conclusions The use of constructed wetlands for domestic wastewater treatment has increased exponentially in recent years, especially for small-scale decentralized units in individual homes and small communities. Wetland systems have been constructed extensively across the globe to treat a wide range of wastewater types. However, the reliability of the effluent quality is still unclear because their dynamics are not

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well understood. However, as a passive treatment unit, a constructed wetland is the most environmentally friendly treatment option. Performance monitoring and reporting requirements vary widely spatiotemporally, especially for small-scale systems, so a centralized repository of information on wetland systems does not exist. Many small-scale wetland designs are based on intuition, best practices, or borrowed designs from other treatment technologies (e.g., facultative lagoons). Hence, a deeper exploration of the feasibility of constructed wetland systems and a better understanding of the dynamics of this green approach are required for synergistic networking and standardization of the system specifications.

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Vymazal J (2005) Horizontal sub-surface flow and hybrid constructed wetlands systems for wastewater treatment. Ecol Eng 25:478–490 Vymazal J (2010) Constructed wetlands for wastewater treatment: five decades of experience †. Environ Sci Technol 45(1):61–69 Vymazal J, Dusek J, Kvet J (1999) Nutrient uptake and storage by plants in constructed wetlands with horizontal sub-surface flow: a comparative study. In: Vymazal J (ed) Nutrient cycling and retention in natural and constructed wetlands. Backhuys Publishers, Leiden, pp 85–100 Wallace SD, Parkin GF, Cross CS (2001) Cold climate wetlands: design and performance. Water Sci Technol 44(11/12):259–266 Wallace SD, Knight RL, Water Environment Research Foundation (2006) Small-scale constructed wetland treatment systems: feasibility, design criteria, and O and M requirements. WERF/IWA Publishing, Alexandria/London Wass RD, Fox P (1993) Constructed wetlands for nonpoint source control of wastewater from a vehicle maintenance yard. In: Proceedings of the 66th annual WEFTEC conference, Anaheim. Water Environment Federation, Alexandria, pp 9–20 Weider RK (1989) A survey of constructed wetlands for acid coal mine drainage treatment in the eastern United States. Wetlands 9:299–315 Wemple C, Hendricks L (2000). Documenting the recovery of hydrocarbon-impacted wetlands: a multi-disciplinary approach. In: Wetlands and remediation: an international conference. Battelle Press, Columbus, pp 73–78 Wetzel RG (2001) Limnology – lake and river ecosystems, 3rd edn. Academic Press, Cambridge, USA Williams CR, Jones RD, Wright SA (1996) Mosquito control in constructed wetlands. In: Proceedings of WEFTEC’96, 69th annual conference and exposition in Dallas, Texas. Water Environment Federation, Alexandria, pp 333–344 Wynn TM, Liehr SK (2001) Development of a constructed subsurface-flow wetland simulation model. Ecol Eng 16(4):519–536

How to Improve Selectivity of a Material for Adsorptive Separation Applications

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Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Adsorption Technique and Adsorbent Materials . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Role of Adsorbent in Different Technologies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Industrial Wastewater Treatment Using Low-Cost Materials as Adsorbent . . . . . . . . . . . . . . . Adsorption Selectivity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Post-modification of Adsorbents to Improve Selectivity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chemically Modified Natural Adsorbents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Activated Carbon . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Improvement in Adsorption Selectivity of Natural Zeolite . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Modification in Natural Clay . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chemically Modified Chitosan as Biomaterial . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Agricultural Waste Adsorbents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Soil/Silica Adsorbents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Metal-Organic Framework . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Future Prospects of Modification Techniques . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cross-References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

This chapter analyzes how adsorption selectivity of natural and synthetic materials was improved with different approaches, particularly for their adsorptive water treatment applications. To explain this approach, it includes some of the frequent modification in natural materials like activated carbon, natural zeolite, natural clay, biopolymer (like chitosan, cellulosic waste), sand, and some of the post-synthetic modification in materials like metal oxides, mesoporous silica, and metal-organic frameworks (MOFs). Herein, different types of modification approaches applied to an individual category of material were reviewed, V. K. Saini (*) · A. Shankar School of Environment and Natural Resources, Doon University, Dehradun, Uttrakhand, India © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_43

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systematically. The effects of these modifications on surface area, porosity, adsorption capacity, and surface chemistry were discussed in terms of change in their adsorption selectivity. For instance, it comprises impacts of oxidation, sulfurization, nitrogenation, and coordinate ligands on the surface properties of activated carbon. Similarly, the role of metal ion, surfactant cations, metal oxides, and polymers on the change in ion exchange properties of zeolites was explained. The change in physicochemical properties of clay on thermal treatment and treatment with ionic and organic species was compared. Likewise, the adsorption selectivity of the clay-based composite with different polymers and effects of modification of chitosan, agricultural by-products, sand, silica, metal oxides, and MOFs were deliberated. In the end, some of the future challenges in the field of adsorption selectivity are discussed. Keywords

Adsorption technique · Adsorbent materials · Activated carbon · Natural zeolite · Chitosan · Biomaterial · Metal-organic framework

Introduction Adsorption Technique and Adsorbent Materials Adsorption is a universal process, which has long been considered as a competent loom for pollutant control. This process occurs when a gas or liquid solute makes a molecular or atomic film (adsorbate) on the surface of a solid or liquid (adsorbent) (Fig. 1). It is operational in most natural physical, biological, and chemical systems, widely in the field of industrial wastewater treatment because of its simplicity and cost-effectiveness. This process is generally classified into physical adsorption (due to weak van der Waals forces) and chemical adsorption (due to covalent bonding). Adsorption can also occur due to electrostatic attraction. The adsorption by adsorbent materials is a classical technique for water treatment. It is useful for the removal of organic as well as inorganic pollutant from wastewater. There are many other methods like coagulation, precipitation, ozonation, reverse osmosis, ion exchange, and advanced oxidation processes available for the purification of polluted water and wastewater. Adsorption processes do not produce harmful by-products. Further, it has advantage over the other methods because of its economical and simple design. In every adsorption-based process, adsorbent plays a critical role. In general, an adsorbent material is characterized by its large specific surface area, pore size distribution, and surface polarity (surface functionality). A large number of adsorbents are available for adsorption application. All these adsorbents can be categorized in a number of ways. One of the classifications is in terms of natural adsorbents and synthetic adsorbents. The natural adsorbents are those which can be extracted from the nature like clay materials, natural minerals, metal oxides, cellulose, natural fibers, and many others. These materials are eco-friendly, low cost, and rich in supply, whereas those adsorbents which are not found naturally can be prepared

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Fig. 1 Adsorption phenomenon

ADSORBATE

ADSORPTIVE

ADSORBENT

chemically, known as synthetic adsorbents. At present, several synthetic adsorbents such as synthetic resin, synthetic zeolites, synthetic carbonaceous adsorbents, and polymeric adsorbents are frequently in use. The selectivity of natural adsorbents is slighter than the synthetic ones; however the selectivity and adsorption capacity of natural adsorbent or even synthetic adsorbents can be improved by further modification in their structural properties.

Role of Adsorbent in Different Technologies The adsorbents have various applications in different areas of science and research as shown in Fig. 2. The versatile physicochemical properties of these adsorbents make them ideal choice for a number of important technologies. Some of the important technologies are discussed here below. In the gas separation application, adsorbent materials selectively capture the gas from the gases stream bought into contact to its surface (Velu et al. 2003). Adsorptive separation of many gaseous pollutants like NOx, SOx (due to the combustion of gasoline, jet fuel, and diesel fuel), natural gases, and hydrocarbons such as CH4, N2, CO2, H2S, and He has been reported in the literatures (Du and Zou 2016; Misra et al. 2016). In petroleum refinery process, deep hydrosulfurization with the conventional hydrotreating technologies (HDS. removal of sulfur; HDN, removal of nitrogen) is applied for the removal of nitrogen and sulfur, found in the gas oil, over the novel polymeric adsorbent prepared by using unsaturated polyester resins (UPR), glycidyl methacrylate (GMA), and divinylbenzene (DVB) as monomers (Al-Daous and Ali 2012). Alumina immobilized with Ni and Mo is also widely used in industry for hydrotreating the gas oil (Toni Raabe et al. 2014). However for removal of pollutants from the natural environment, the adsorption on the surface of zeolite immobilized with (Cu, Pb, Zn, Ni, Ca) ions and modified active carbon, silica, and metal oxides has been reported. Iron-based adsorbent has been used for removal of H2S and oxygen from biogas (Shimekit and Mukhtar 2012). In spite of gas separation, the adsorbent is also being used in gas purification technology. In gas purification technology, the use of cyclic thermal swing adsorption is a widely applicable process for CO2 removal from natural gas or drying (Mersmann et al. 2011). Mostly the temperature swing adsorption is applied on the molecular

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Photocatalysis

Drug Photocatalyst Nanoparticle

Waste water

Drug loaded nanoparticle

Clean water

Nanofiltration

Reverse osmosis

Ultrafiltration Microfiltration

E.coli Oil Macromolecules Colloids Suspended particles

Viruses Proteins

Ions Small compounds

Water Treatment

Gas separation & Purification

Fig. 2 Application of adsorbents in different technologies

sieves loaded with H2O, SO2, and CO2. The TSA process is also applied for the regeneration of activated carbon, zeolites, and silica gel loaded with hydrocarbons present in natural gas (Cavenati et al. 2006). Zeolite as microporous material preferentially absorbs CO2, which is mostly used in pressure swing adsorption (PSA) processes. Zeolites, due to their higher selectivity for CO2, are preferred over other adsorbent (activated carbon) for this purpose (Worch 2012). From ancient Greece times, the adsorbents are being used for purification of drinking water. Since then, this technique has been continuously used for water treatment with new improvements from time to time. Adsorption-based water purification technologies are efficient for removing specific types of impurities down to very low levels in the parts per billion (ppb) and parts per trillion (ppt) ranges. The application of activated carbon has proven it as good adsorbent for the removal of wide range of water pollutants such as phenols, chlorinated hydrocarbons, pesticides, pharmaceuticals, personal care products, corrosion inhibitors, and so on (Yang et al. 2016a). Fibrous filters and activated carbon filters are also used as a tertiary treatment process in drinking water purification. Ion exchange technique for the purification of water by porous beads as adsorbent prepared from highly crosslinked insoluble polymers with large number of strongly ionic exchange sites is also an important technique for the removal of charged ions from water. In the field of medical science, adsorbents in the form of nanoparticles are used as a drug carrier for the treatment of diseases. To transmit adequate dose to the lesion, suitable carrier of drugs is needed. Nano- and microparticles such as chitosan, drug-

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clay nano hybrid (Torres et al. 2016), carbon nanotubes (Tamayo et al. 2015), mesoporous silica (Wang et al. 2015a), and carbon nano fiber (Kumar 2000) carrier have important potential applications for the administration of therapeutic molecules. Controlled drug delivery technology offers various advantages over the conventional dosage forms, which includes improved efficacy, reduced toxicity, and improved patient compliance and convenience. Synthesis of polymeric nanosphere micelle based on biocompatible copolymer of poly (ethylene oxide)-poly (L-Lactic acid)/poly(β-benzyl-L-aspartate) has been tested as a vehicle for delivery of anti-inflammatory and antitumor drugs (Borges et al. 2016). Adsorbents are good catalytic supports due to their large surface area. In heterogeneous catalysis, the active catalysts are often supported/incorporated on the surface of adsorbents. It improves the efficiency and recovery of these catalysts during catalytic reaction. Photodegradation by the photocatalyst-loaded adsorbents for the removal of organic pollutants from the wastewater has been reported in the literature. Activated carbon developed from several sources such as natural mineral material (Etim et al. 2015), coconut shell, rice husk, and bark has been immobilized with TiO2 which has also been reported for the removal of phenol from water. Combination of UV radiation and TiO2 gave a reasonable degradation efficiency (Ijaola et al. 2013). Pure chitosan has been grafted with ZnO photocatalyst to obtain ZnO/Cht-MCM. This material can be used for photodegradation of azo dye and methyl orange from the aqueous solution.

Industrial Wastewater Treatment Using Low-Cost Materials as Adsorbent Sudden increase in population, rapid urbanization, industrial and technological expansion, energy utilization, and waste generation from domestic and industrial source have rendered many noxious and hazardous waste into the environment which are harmful to human health as well as the aquatic life. Inadequately treated wastewater or untreated discharge by the industries into the water ways is the main cause of water pollution. Industrial wastewater is often contaminated with numerous compounds, for example, heavy metals, dissolved organic compounds, suspended solids, etc. Industrial discharges resist the self-purification capabilities of the river as well as decomposition in wastewater treatment plant (Ademiluyi et al. 2009; Khan et al. 2016). Therefore, it is imperative that it should be treated to an environmentally acceptable limit. The physicochemical processes have proven useful tool for selective removal of organic and inorganic pollutants from the wastewater. Various natural occurring materials are present in large quantity having properties of adsorbents. The abundance of these materials and their low cost actually make them suitable as an adsorbent for the removal of various water pollutants. Their naturally occurring adsorbents involve clay, sand, charcoal, chitin, agricultural by-products and natural zeolites, etc. Although all these types of materials are abundant in nature, their use is often limited, due to their poor adsorption properties. Some of the natural adsorbents used for wastewater treatment are discussed below.

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Activated carbon prepared from different natural source materials (e.g., bamboo, rice husk, peanut hull, coffee husk, coal, coconut shells, lignite, wood, etc.) is the most popular and widely used adsorbent in the field of wastewater treatment throughout the world. Removal of pollutants such as dyes (Kumar and Jena 2016), phenols (Ibrahim et al. 2016), heavy metals (Saygili and Guzel 2016), pharmaceutical drug (Salman 2014), pesticides (Water Treatment Using Carbon Filters: GAC Filter Information 2004), fertilizers, detergents, oil, and greases has been reported on this type of materials. Post-modification of materials improves properties toward pollutants. A filter with granular activated carbon (GAC) is a proven option for the removal of chemicals that give objectionable odors or taste of water such as hydrogen sulfide (rotten egg odor) or chlorine. Excess release of heavy metals in aqueous waste of many industries such as metal plating, tanneries, radiator manufacturing, alloy industries, mining operation, and storage batteries has created an environmental problem worldwide. There are numerous reports available involving heavy metal removal using natural adsorbents developed from agricultural waste or industrial by-products. Various natural adsorbents such as chitosan (Turan et al. 2014), peat (Poots et al. 1976a), wood (Poots et al. 1976b), natural coal (Kong et al. 2016), natural clay (Yang et al. 2016b), sawdust (Al-Riyami et al. 2014), and zeolite (Mudasir et al. 2016) have been reported for the removal of dyes, heavy metals, and phenols (Su-Hsia Lin and Juang 2009) from the wastewater. Nowadays, the use of adsorbents prepared from low-cost agricultural waste has been increasing for the purification of water (Fig. 3). Various studies on adsorbents like rice husk, neem bark, waste tea, walnut shell, maize leaf, sago waste, etc. have been also reported in literature. Important benefits of using agricultural waste in wastewater treatment include easy technique, modest processing, reasonable adsorption ability, Adsorbed Pollutants H2O In

H2O Out

Adsorbent Material Fig. 3 Water treatment using adsorbent material as a filter

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economical, easy availability, and regeneration. On the other hand, the use of untreated agriculture waste can fetch problems of COD, BOD, and TOC in water and make them suitable for aquatic life (Tripathi and Rawat Ranjan 2015). Widespread industrial activities are producing large amount of solid waste materials. Some materials are being put to use, and some get dumped elsewhere. The industrial by-products can be found easily in free of cost. The dumping of solid waste is also an environmental issue. So, one of the best ways to reduce the solid waste is to use these wastes as an adsorbent material. With this view, several industrial wastes are being investigated with or without treatment as an adsorbent for wastewater treatment. The use of fly ash produced as a by-product from the thermal power plant has been investigated for the removal of phenol and its analogue (Sarkar and Acharya 2006). Other industrial wastes such as blast furnace slag, dust, and sludge (Chazarenc et al. 2010) from steel industry and red mud (López et al. 1998) from aluminum industry have also shown considerable effect for the removal of heavy metal pollutants from wastewater.

Adsorption Selectivity The preferential adsorption of a component on over an adsorbent surface from the mixture of components is termed as its adsorption selectivity. The adsorption selectivity is usually controlled by two factors. One of which is surface functionalization. The functional groups present on the surface of adsorbents interact differently with different molecules from the adsorbate mixture. The extent of interaction is related with the functional groups present on the adsorbate molecules. This leads to preferential adsorption of most interactive molecules on the surface of adsorbent and thereby causing separation. The other factor that controls selectivity in some cases is the pore size of adsorbents. This phenomenon is also called as size exclusion. In this process the molecule that matches the size of pores is able to enter in pores and gets adsorbed, whereas others remain in adsorbate mixture. This phenomenon is often observed in controlled pore size materials like zeolites and metal-organic framework (MOFs). The adsorption selectivity is a useful property of an adsorbent particularly for its application in separation and purification process. Separation of pure gas such as CO2 and CH4 from the gas mixture is achieved by using microporous MOFs. The MOF is prepared by mixing ligands, and the resulting material demonstrates higher selective CO2 uptake over CH4 (Misra et al. 2016). The new porous tetrahedral Zn(II)-tetrazolate framework exhibits high CO2 adsorption selectivity over N2 (Li et al. 2016). Microporous MOFs (HKUST-1) have also been used for competitive adsorption and selectivity of benzene and water vapor, and it was found that the benzene molecule is more selective than water vapor at high temperature and pressure (Zhao et al. 2015). The hyper-cross-linked polymeric adsorbent of NDA-100 is more selective for salicylic acid as compared to amine-modified NDA-99 which solves the discharge pollution problem resulted from esterification process (Liu et al. 2007). Discharge of radioactive elements into the environment through nuclear waste effluents, nuclear

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weapon testing, and nuclear power accidents contaminates agriculture soil, water body, and crops. Adsorptive selection of the radioactive elements like Cs and Sr from the environment on the several zeolite materials is also been reported in the literature (Munthali et al. 2015). Adsorption selectivity is also very important for selective removal of emerging water pollutants like pharmaceutical drugs. A study had shown the removal of nonsteroidal anti-inflammatory drugs (naproxen, ibuprofen, and diclofenac) on the surface of molecularly imprinted polymer adsorbents. It was found that the extraction efficiency of MIP was higher than those obtained when non-imprinted polymer was used as adsorbent. The synthesized MIP had proven higher selectivity to diclofenac rather than naproxen and ibuprofen (Trinh et al. 2016). Hydroquinone-modified porous adsorbents have been studied for simultaneous separation and purification of (EGCG) ()-epigallocatechin gallate, and caffeine (CAF) from crude extract of green tea was established by size-exclusion effect. The result showed that the hydroquinone-modified porous adsorbent provided the best separation power due to its porous structure and phenolic hydroxyl group (Zhang et al. 2016a).

Post-modification of Adsorbents to Improve Selectivity As there are many natural and synthetic adsorbents available for the purification, separation, and removal of pollutants from water and gases stream, the modification of these adsorbents is gaining prominence to increase their affinity for certain contaminants to facilitate their removal from the mixture stream. To increase efficiency the adsorbent materials are processed by both physical and chemical treatment. Physical treatment includes freezing, boiling, heating, lyophilization, etc., whereas chemical treatment involves cross-linking with organic solvents, washing with detergents, chemical reaction with a variety of organic and inorganic compounds, alkali or acid treatment, etc. (Farooq et al. 2011). The pretreatment modifies the surface characteristics of adsorbents by exposing more metal binding sites (Vieira and Volesky 2000). Post-modification process tailors the surface property of adsorbents such as surface area, surface charge, pore volume, pore size, and fast adsorption kinetics. Adsorbents such as zeolite, activated carbon, natural clay, low-cost materials, MOFs, MCM-41, and SBA-15 had been post-modified to improve its adsorption capacity. In the literatures, successful synthesis of carboxymethyl chitosan modified with magnetic-cored dendrimers (CCMDs) for the removal of methylene blue (MB) and methyl orange (MO) has been reported. Dendrimers are hyper-branched molecules that possess excellent properties of nanosize and facile surface modification (Calabretta et al. 2007). Because of their dendritic structures, they contain much higher functional group densities than conventional macromolecules. To enhance the application of dendrimer-based sorbents, various materials such as silica gel (Niu et al. 2013), mesoporous silica (Jiang et al. 2007), natural clay (Zhou et al. 2015), fiber (Zhang et al. 2013a), grapheme oxide (DeFever et al. 2015), etc. have been modified (Kim et al. 2016). Adsorptive

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removal of organophosphate insecticide (Azodrin) onto biopolymer-modified montmorillonite (MMT)-CuO composites, viz., MMT-Cuo-Chitosan(ch), MMT-CuOGum ghatti(Gg), and MMT-CuO poly lactic acid (PLA), has also been studied in literatures (Sahithya et al. 2016). In addition, cross-linked chitosan magnetic beads modified with cysteine-glutaraldehyde Schiff base (Chi-CG) were utilized as a possible adsorbent for the elimination of Cu (II) and Cr (VI) from aqueous solutions (El-Reash 2016). Post-modification of other natural adsorbents increases their adsorption performances and permitted an easy separation of pollutants from wastewater (Fig. 4). Zeolite with silver nanoparticle increased the removal of E.coli and heavy metals (Pb, Cd, and Zn) in a continuous flow system (Akhigbe et al. 2016). Selectivity of toxic metal ion Hg(II) has also been investigated on the natural zeolite surface immobilized with dithizone in the toluene medium and shown a high adsorption capacity (Mudasir et al. 2016). To obtain the heterogeneous Fenton-like catalyst, inexpensive adsorbent iron-modified bentonite had been prepared by wet impregnation method to rapidly remove the organic pollutant (rhodamine B) from aqueous solution. In another work, natural magnetite ore was modified with aluminum and lanthanum ions to increase the fluoride removal efficiency from waste water (GarcíaSánchez et al. 2016). Treatment of commercial activated carbon as well as activated carbons prepared from wastes and by-products to improve its surface property has also been explored in the literatures (Rivera-Utrilla et al. 2011). Surface-functionalized mesoporous materials have potential applications in various areas such as catalysis, chromatography, metal ion extraction, adsorption, nanotechnology, and imprinting for metal recognition (Alothman 2012). Nitrate anion removal from water by using an amine-functionalized MCM-41 as adsorbent is reported (Ebrahimi-Gatkash et al. 2015). Similarly, the synthesis and post-modification of another mesoporous silica material (SBA-15) with methyl-diethyl-amine (MDEA) and piperazine (PZ) were also studied for CO2 separation from biogas. The adsorbent showed good performance for separation of CO2 (Xue and Liu 2011). This review aims to summarize the modification techniques available to improve the adsorption selectivity of natural adsorbents. This review can serve as a report that

Unmodified structure

Modified structure

Fig. 4 Image showing the adsorbent after surface modification

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compares different modification techniques in view of their effects in terms of selectivity. It also provides an overview of material-specific techniques and application-specific techniques. It is believed that this article would be useful to researcher in deducing a preferred modification for their objective applications.

Chemically Modified Natural Adsorbents Natural adsorbents are abundant, low cost, and widely been used for the separation and purification processes from the past decades. Sorption on natural surface is one of the cheapest processes to reduce environmental pollution. Chemical modification of these adsorbents by different techniques has shown satisfactory results in terms of their efficiency and applications in various fields. In the surface modification process, functional groups such as amino, aldehyde, phenol, metal oxides, hydroxyl group, carboxyl, etc. are grafted over the surface of natural adsorbents to increase its surface properties like adsorption selectivity, enhanced surface area, pore volume, high adsorption capacity, and adsorption kinetics. In the following section, natural adsorbents along with its available chemical modification techniques that have been reported for modification of different natural adsorbents have been revised.

Activated Carbon Activated carbon is one of the widely used adsorbent (Su-Hsia Lin and Juang 2009). For several decades, it is used as a water industry’s standard adsorbent for the reclamation of industrial and municipal wastewater (Kaushik et al. 2008). Activated carbon developed from plant or animal waste and industrial by-products have shown commendable removal efficiency over pollutants. This material is low cost, requires little or no processing, is easily available, is biodegradable, and can be easily disposed by incineration. Modification and impregnation techniques improve its selectivity and removal capacity over pollutants in gases and water streams. Modification of activated carbon can be performed by treating with oxidizing agents, such as nitric acid, ammonium persulfate, hydrogen peroxide, etc. By this treatment, alteration in textual property of AC and generation of acidic oxides (carboxylic, phenolic, lactones) and/or basic groups (pyrone-like groups) occur. Oxidized activated carbon extracted by using HNO3 as a reflux is reported in literatures (de Mesquita et al. 2006). Likewise, activated carbon was also prepared from olive stones and almond shells with different degrees of activation oxidized with HNO3, H2O2, and (NH4)2S2O8 mentioned in literatures (Moreno-Castilla et al. 1995, 1997, 2000). One to 10 wt% aqueous solution of sodium dichloroisocyanurate (DCI) was used to introduce oxygen and chlorine surface groups on activated carbon by chemisorption (Molina-Sabio et al. 2011). Electrochemical oxidation of activated carbon fiber has also been reported (Chiang and Chen 2010). Sulfur modification is another technique to increase the efficiency of AC. Activity of carbon-sulfur complexes in adsorption processes is the most important factor due

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to their chemical nature. This determines during the acid-base character of carbon surface, in such process. Activated carbon derived from nut shell had been modified with sulfur by applying SO2 gas as sulfurizing agent (Fouladi Tajar et al. 2009). In a different approach, AC from sugarcane bagasse was made followed by single-step steam pyrolysis in the presence of SO2 and H2S as sulfurizing agent (Krishnan and Anirudhan 2002). Nitrogenation on the surface of AC increases its basicity which is a required property for adsorption and catalysis. Depending on the reagent used, the process was conducted either in gas (ammonia, hydrogen cyanide, and amines) or liquid (nitric acid, urea amines, and ammonia) phase. Sometimes, the process may consist two stages, (i) single nitrogenation stage and (ii) two successive stages in which an AC substratum was first oxidized in liquid phase and then nitrogenation in gas or liquid phase. To enhance the adsorption of phenol from aqueous solution, woodbased activated carbon had been modified with nitrogenation by flowing ammonia and nitrogen throughout the process in quartz tube (Yang et al. 2014). Similarly, activated carbon developed from polish brown coal, bituminous and subbituminous coals enriched with nitrogen by ammoxidation (simultaneous oxidation and nitrogenation of fossil coals) for removal of NO, and contaminants from liquid phase has been studied (Pietrzak et al. 2010; Pietrzak 2009). In addition, ruthenium-based catalysts supported on activated carbon (neutral and coconut carbon) modified with NO2, NH2, and N-H-N groups had been assessed for their catalytic activity for acetylene hydrochlorination, aimed to improve the activity and stability of the acetylene hydrochlorination catalysts (Xu et al. 2015). Functionalization of nitrogen onto surface of AC with nitration to improve its selectivity for CO2 capture has been mentioned in literatures (Zhang et al. 2013b; Jennifer Wilcox et al. 2014). Coordinate ligands are also being used to alter the AC and the textural (surface area, porosity) chemical properties and enhance its adsorption capacities. Modification of AC depends on the composition of legends used. Studies show the effect of water-soluble ligands (sodium borohydride, polyvinyl alcohol, glucose, and galactose) on the preparation of nano-silver-supported activated carbon prepared from German beech wood (Eltugral et al. 2015). Activated carbon modified with hybrid ligands, i.e., combined treatment of nitric acid and thionyl chloride, followed by the reaction with ethylenediamine, to introduce S-, N-, and Cl-containing functional groups for the removal of mercury in aqueous solution had been reported (Zhu et al. 2009). To improve photocatalytic properties of coordination complex-modified polyoxometalates (CC/POMs) in visible light region, its nanobelts were loaded on activated carbon fiber. The resulted material (CC/POMNBs/ACF) was studied for the degradation of rhodamine B under visible light irradiation (Lu et al. 2015).

Improvement in Adsorption Selectivity of Natural Zeolite Zeolites are naturally occurring hydrated aluminosilicate minerals with a cage-like structure. These materials come under the class of minerals known as “tectosilicates.”

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Naturally zeolites are formed by the volcanic rocks; ash layer reacts with alkaline groundwater and crystallizes in postdepositional environment over the years in shallow marine basins. The structures of zeolite consist of three-dimensional frameworks of SiO+4 and AlO+4 tetrahedral (Khachatryan 2014). The position in the center of the tetrahedron of four oxygen atoms is occupied by the aluminum ion, and the isomorphous replacement of Si+4 by Al+3 lifts up a negative charge in the lattice. The net negative charge is balanced by the exchangeable cation (potassium, sodium, or calcium). With certain cations these cations are exchangeable in solutions such as lead, zinc, manganese, and cadmium (Barrer 1978; Breck 1964). This exchangeable ion makes zeolite an effective adsorbent for the removal of heavy metals in industrial effluent water. Commonly, surface alteration of natural zeolite adsorbent is being done using a specific and sensitive ligand to enhance the capacity and selectivity. Different techniques are discussed under subsequent heading for the modification of zeolites.

Metal Cation-Modified Zeolites Zeolites are capable to exchange ions with external medium, which is one of their remarkable characteristics. The ion exchange property of natural zeolite depends on several factors, including ion shape, size, charge density of mineral network, framework structure, ionic charge, and concentration of the external electrolyte solution (Bish and Ming 2001). Natural zeolite (clinoptilolite) has been modified with Na and Ce-Fe. To obtain zeolite-rich tuff, the adsorbent was kept in contact with NaCl and CeCl3-Fecl3 solutions (Olguin and Deng 2016). Likewise, activation of natural zeolites (clinoptilolite) with Na+ cation and functionalization with barium and copper ions have also been reported (Rodrigues and Rubio 2007). The modification involves two steps of ion exchange, (i) activation of Na-Z and (ii) functionalization of Ba-Fz and Cu-Fz. For the activation of zeolites, powdered material was taken in an aqueous solution. Barium chloride and copper nitrate in aqueous solution along with activated zeolite are used for functionalization of the materials. This process increases the exchange of cations from solution to the zeolite framework. For disinfection and metal removal, silver modified with zeolite has also been investigated (Akhigbe et al. 2016). Modification of natural zeolites with surfactant cations has shown commendable effects. Cationic surfactants have positive charge head group attached with hydrocarbon moiety. To prepare cationic surfactant-modified zeolites (CSMZ), several cationic surfactants such as tetramethylammonium, tetraethylammonium bromide, hexadecyltrimethylammonium bromide or chloride, cetylpyridinium bromide, ethyl hexadecyl dimethyl ammonium, and 4-methylpyridinium are mentioned in the literatures (Apreutesei et al. 2008; Krajišnik et al. 2010). Similarly, cation surfactants (benzalkonium chloride and cetylpyridinium chloride) have also been observed for the modification of clinoptilolite (Daković et al. 2013). In addition, Pd2+- and Ni2+-modified aluminosilicates are used as a filter to decrease heavy metals in cigarette smoke (Afzali et al. 2007). Ion exchange on the surface of NaX and NaY zeolites modified with alkali metal cations such as Li+, K+, Rb+, Cs+, Ni2+, and Cr3+ is studied for the adsorption of CO2 and nitrogen (Bendenia et al. 2011; Walton et al. 2006).

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Oxide-Modified Zeolites One of the most common modifications to enhance adsorption capacity of natural zeolite is by iron oxide. According to literature, the FeOZ has a higher specific surface area. This characteristic makes it better than the unmodified zeolites (Doula and Dimirkou 2008). For example, natural zeolite modified with FeO was obtained by adding natural clinoptilolite in an iron nitrate solution under strongly basic conditions (Doula and Dimirkou 2008). Similarly, clinoptilolite modified iron oxide prepared by mixing natural clay into the freshly prepared solution of Fe(NO3)3.9H2O and KOH solution followed by drying and washing. This modification increases the surface of produced adsorbent that consists of negative charge, and this makes cations adsorb on the adsorbent under acidic condition (Dimirkou and Doula 2008). In another study, iron oxide-modified clinoptilolite was prepared by adding a mixed amount of (natural clinoptilolite sodium) NC-Na into a solution containing metal and NaOH. Afterward the sample is immersed into FeCl2 solution. This fabrication increases the metallic species on the surface of adsorbent for the adsorption of pollutants (Trinh et al. 2016). As the like, the Mexican natural zeolite-impregnated iron oxides were obtained by adding zeolite into the Fe solution prepared from FeSO4.7H2O and FeCl3 salt. NH4OH added dropwise into the suspension. This modification enhanced the pores and surface properties of the adsorbents (Mockovčiaková et al. 2006). Zeolites can also be modified with manganese oxides (Catts and Langmuir 1986). In this view, to prepare the manganese oxide-modified clinoptilolite, the zeolite was added in MnCl2 solution for ion exchange. Afterward, the ion-exchanged zeolite is washed with water, filtered and dried, and immersed in KMnO4 solution. This modification increases the selectivity of the adsorbent toward the guest molecules of Zn2+ (Irannajad et al. 2015). Similarly, clinoptilolite- and bentonite-modified MnO was formed via precipitation. For this, the samples of zeolites were stirred in KMnO4 solution. Then, HCl is added dropwise into the solution and heated. After titration, the product is stirred, washed, and dried. After modification the cation exchange capacity is increased over the surface of adsorbents (Schütz et al. 2013; Mohamadreza and Maryam 2014). Polymer-Modified Zeolite Surface modification of natural zeolites with polymer improves their efficiency and makes them more selective for adsorption process. Polymer modified with zeolite forms a bilayer structure on zeolite surfaces. A hybrid polymer has been developed with silicon alkoxide and a metal alkoxide, a co-monomer. The material is modified by contact of hybrid polymer with a suspension consisting of zeolite and a solvent. This modification increases the removal efficiency of the adsorbent (Nemeth and Xu 2014). Hydrothermal acid-treated faujasite-type zeolites (H-Y14, H-Y40, H-Y144, and H-Y770) can be modified with two units of polyvinyl alcohols (PVA) with polymerization degree of 500 (PVA 500) and 2000 (PVA 2000). The zeolite was mixed with a solution of PVA and stirred. This fabrication results the strong adsorption affinity toward the organic pollutants (Kawai 2013). The improvements in permeability and selectivity during gas/liquid separation on introduction of inorganic sieve materials into an organic polymer matrix have received a worldwide recognition.

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In this context, a silane coupling agent ((3-aminopropyl)-diethoxymethylsilane (APDEMS)) has been used to modify surface of zeolite for mixed matrix membranes (MMMs). For this, mixture of toluene, APDEMS, and zeolite is stirred under N2. Afterward the mixture is washed with toluene and methanol to remove the untreated silane and then dried the modified zeolite under vacuum condition. Other silane coupling agents are (aminopropyl)-triethoxy silane, N-(aminoethyl)-aminopropyltrimethoxy silane, (-glycidyloxypropyl)-trimethoxy silane, and (3- aminopropyl)dimethylethoxy silane (Li et al. 2006). The negatively charged surface of zeolite can be converted into positively charged ions; zeolite is modified with propylamine, and N- cetyl- N,N,N- trimethyl ammonium brome (CTAB) has also been mentioned in the literature (Frida et al. 2014).

Modification in Natural Clay Clays are hydrous aluminosilicate natural adsorbent and are nontoxic to ecosystem. Clay minerals are classified as chlorite, illite, kaolinite, montmorillonite, smectite, etc. In which, montmorillonite and kaolinite are widely used because of their low cost, chemical and mechanical stability, high surface area, and a variety of surface and structural properties. Natural clays act as a scavenger of pollutants by embracing cations and anions either through ion exchange or adsorption or both. The clay and their modified composites have grown to be the materials of increasing interest due to their nanosized structural and functional properties. Different techniques that have been used for clay modification are reviewed in this section below.

Polymer-Clay Composites Polymer-clay composites have received special attention in the field of adsorption. In minimal addition, it boosts the mechanical, thermal, dimensional, and barrier performance properties of clay material. Several methods have been developed to fabricate polymer-clay composites (Fig. 5). These are in situ polymerization, where monomer is dissolved in a solution with clay followed by in situ polymerization, induced intercalation where polymer or a prepolymer is dissolved in a solution with clay, and melt processing where polymer is melt-mixed with the clay above the softening point of the polymer (Gao 2004). Modification of montmorillonite (MMT) with poly (diallyl dimethyl ammonium chloride) (PDADMAC) and poly-4-vinylpyridine-co-styrene (PVPcoS) has been reported. The surface is modified by adding the polymer solution to dispersion of MMT (at a volumetric ratio of 2:1) and continually stirred for 1–4 days. After centrifuged and washing to remove free polymers, the polymers are fabricated over its surface. This modification increases the affinity of adsorbent to get selective for organic pollutant (Ganigar et al. 2010). In another study, polyacrylonitrile (PAN) was grafted on kaolinite by treating kaolinite with different ratios of hexadecylethyldimethylammonium (HDEDMA) for the sorption of methylene blue and Cr (VI) has also been investigated (El-Zahhar 2015).

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a Phase separated (Microcomposite)

b Intercalated (Nanocomposite)

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c Exfoliated (Nanocomposite)

Fig. 5 Different types of composites arise from interaction between layered silicate and polymers

A variety of polymers including polypropylene, polyethylene, polystyrene, poly(ethylene oxide), polycaprolactone, polyimides, polyamides, poly(ethylene terephthalate), polycarbonate, polyurethane, polyethylenimine, and epoxy resins have been grafted over different types of clay in literatures (Nguyen and Baird 2006; Alsewailem and Aljlil 2013; Bahrami and Mirzaie 2011; Chen et al. 2016; Zheng and Wilkie 2003). For instance, a natural montmorillonite modified with polypropylene was prepared by using benzyl tallow dimethyl ammonium salt composites (at different ratios) by melt mixing and pressing. In this process maleic anhydride- modified polypropylene was used as a compatibilizer. This process increases the mechanical, thermal, and dye adsorption capacity of modified clay (Bahrami and Mirzaie 2011).

Clay Modified with Phosphate Clays can also be modified with phosphate, where it can be studied for the removal of heavy metals (Pb, Zn, Cd, Cu, Co, etc.) from aqueous medium that has been reported. In this view, natural kaolinite clay can be modified by treating with sodium polyphosphate and monopotassium phosphate (KH2PO4). This modification increases the ion exchange capacity of clay and, hence, increases the adsorption of pollutant over surface (Adebowale et al. 2006; Amer et al. 2010; Unuabonah et al. 2007; Olu-Owolabi and Unuabonah 2011). Similarly, montmorillonite and bentonite modified with KH2PO4 were studied for the removal of heavy metals (Zn, Cu, Co, Sr, and Cs) (Ma et al. 2011; Unuabonah et al. 2010). Potassium dihydrogen phosphate and sodium polyphosphate have also been used for the modification of kaolin. The phosphate-modified kaolinite clay pretreated with Ca2+ is studied for the adsorption of Pb2+ from aqueous solution in batch mode. The resulting adsorbent showed high Pb2+/Ca2+ ion exchange on modified surface (Cojocariu et al. 2012).

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Chitosan-Coated Clay Chitosan-clay nanocomposites have a considerable prospective for the biomedical field and wastewater treatment. For combining the advantage of this biopolymer with clay in a drug delivery processes, the hot intercalation technique is used. It develops quaternized chitosan-montmorillonite (HTCC-MMT) nanocomposite materials. Glutaraldehyde can also be used for building cross-linking between chitosan with clay for the release of paracetamol and theophylline and xanthine and NO-donor compounds (Bothiraja et al. 2014). Chitosan biopolymer-modified montmorillonite and halloysite have also been reported in the literatures for the modulation of drugs (Saji George and Luda 2006; Liu et al. 2015a). Similarly, sepiolite and attapulgite-modified chitosan can also be prepared by acid treatment with HCl, where acid-treated clay suspension was mixed into the chitosan solution by stirring. It causes protonation of – NH2 group on chitosan that induces an expansion of the polymer network and thereby increases the adsorption (Deng et al. 2012; Elsergany et al. 2016). Eight adsorbents with different chitosan-clay (bentonite) ratios were prepared by treating chitosan with acid (acetic acid), and then the dissolved chitosan was mixed with bentonite clay suspension followed by washing and drying. This modification improved separation and physical characteristics of adsorbents and increases its selectivity for metal removal (Dey et al. 2016). Similarly, nano-chitosan kaolin clay was produced in 1 M acetic acid and stabilized after dropwise addition of mixture of kaolinite and chitosan solution in 3 M NaOH which are mentioned in literatures. Numerous methods for the modification of composite materials such as pultrusion, autoclave processing, prepreg method, wet lay-up, filament winding, template synthesis, fiber placement technology, intercalation of polymer, melt intercalation, in situ polymerization, etc. have also been reported in the literature (Mazumdar 2001; Dubois 2000; Evans and Pancoski 1989). Organic Modification of Layered Clay Organically modified clays are prepared by replacing the exchangeable inorganic cations present in the natural clay with organic cations of large size. Organic cations used in replacement process are then able to adsorb other compounds. Organically modified clay can be prepared by either wet or dry process. In a typical wet process, unmodified clay slurry is allowed to react with quaternary ammonium compound. Thereafter, the mixture was filtered, dried, and ground. In dry process, limited amount of moisture is added to the clay. The clay is then allowed to react with organic compound in a mixer or pug mill. Finally, the reacted material is dried and ground (Masooleh et al. 2010). Organically modified montmorillonite (MMT), bentonite (BN), and commercial Na-BN clay by exchanging hexadecyltrimethylammonium bromide (HDTMA) with sodium ions have been tested for the adsorption of hydrocarbons, phenol, and dyes. The organoclay was prepared by the dry process in which HDTMA was used as an ammonium source. After modification, surface properties alter from hydrophilic to hydrophobic, imparting better interaction of the clay products toward hydrophobic pollutants in the environment (Binoy et al. 2010; Bedin et al. 2013). In another study, toluene adsorption by the sodium smectite modified with (HDTMA) by the process of organophilization

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has been reported. The technique consisted of organophilization cation exchange of sodium present in commercial clay, with the quaternary ammonium salt (GarcíaLópez et al. 2010). Modification of sepiolite clay was made with a quaternary ammonium salt, trimethyl-hydrogenated tallow (3MTH) through melt-compounding approach (Sperberga et al. 2015).

Thermally Activated Clay Adsorption capacity of an adsorbent is directly related to its physicochemical properties such as moisture content, volatile matter content, specific surface area, etc. To improve the adsorption on illite clay surface, it was chemically and thermally activated by using NaOH and KOH solutions; afterward the material was thermally activated at 600  C, 700  C, or 800  C temperatures for increase in their mechanical strength (Egbuna et al. 2014). The thermal activation showed high correlation of the specific surface area and methylene blue adsorbent value with temperature (Sarikaya et al. 2000). Thermal activation of calcium bentonite clay at temperature ranges from 100  C to 1300  C has also been reported (Chen and Mah 1997). To increase the adsorption capacity and specific surface area of kaolinite, thermal treatment at temperature from 450  C to 650  C has been applied. During this treatment, its crystalline structure breaks into amorphous and forms one metakaolinite (Hall and Yalpani 1980).

Chemically Modified Chitosan as Biomaterial Chitin is the second most abundant natural biopolymer. Chitosan can be produced chemically from chitin, found in crustaceans and insects. It is a linear cationic heteropolymer of randomly distributed N-acetylglucosamine (GlcNAc) and glucosamine (GlcN) residues with b-1,4-linkage. Its surface is consisting of amino group and both binary and secondary hydroxyl groups at the C-2, C-3, and C-6 positions, respectively. Among the other polysaccharides, its chemical structure allows specific modifications, especially in C-2 position. Being an ideal support material for enzyme immobilization, chitosan offers an interesting characteristic of nontoxicity, biodegradability, biocompatibility, and bioactivity. During modification, the functional groups allow direct substitution reactions and yield useful materials for different domains of application. In the following section, different techniques of chitosan modification are discussed.

Chitosan Modified with Sugar Improvement in surface potential of chitosan can be made by modification with sugar moieties for their applications, especially in the field of biomedical science. The first report on sugar-modified chitosan derivative was published in 1980 by Hall and Yaplani (Sashiwa and Aiba 2004). They synthesized sugar-bound chitosan by reductive N-alkylation using NaCNBH3 and unmodified sugar (Sashiwa and Aiba 2004; Morimoto et al. 2001). Primarily, sugar-modified chitosan had been investigated mainly for the rheological studies. But now, sugar-modified chitosan can be used

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Adsorbent

Drug Loading

Drug loaded adsorbent aer washing

Fig. 6 Diagrammatic representation of drug loading in porous adsorbent

for drug targeting because of specific recognition of cells, viruses, and bacteria. This type of modification has usually been used to introduce cell-specific sugars into chitosan. Synthesis of sugar-modified chitosan with D- and L-fucose and their specific binding with lectin and canine polymorphonuclear leukocyte cells by reductive alkylation were studied. Ulex europaeus agglutinin I (UEA-I) and N-acetyl-D-glucosamine were used as sugar derivatives, respectively (Park et al. 2003). Chitosan derivative prepared by the covalent attachment of gluconic acid has also been studied. This sugar-bearing chitosan was further modified by N-acetylation in alcoholic aqueous solution. This modification improved the biodegradability of the chitosan (Li et al. 1999). Modification of chitosan by N-alkylation performed in aqueous methanol with various aldehydes, monosaccharides, and disaccharides (glycolaldehyde, DL-glyceraldehyde, D-ribose, D-xylose, D-arabinose, 2-deoxy-D-ribose, D-galactose, 2-deoxy-D-glucose, 3-O-Me-D-glucose, L-rhamnose, L-fucose, GlcNAc) for gene delivery, cell culture, targeted drug delivery system (Fig. 6), and tissue engineering has also been studied in the literature (Morimoto et al. 2001; Mourya and Inamdar 2008; Lee et al. 2005).

Chitosan-Dendrimer Hybrid Dendrimers are highly branched and symmetrical macromolecule which comes under polymer family (Zhao and Crooks 1999). Dendrimers are attractive because of their multifunctional properties (Dong-Lin and Aida 1997; Zimmerman et al. 1996; Shu et al. 2000). The attach-to route and the macromonomer route are the two main approaches to dendronized polymers (Grayson and Fréchet 2001). In the attach-to route, the dendrons are attached to the polymers that possess anchor group (Yin et al. 1998; Jahromi et al. 1998). In macromonomer route, desired number of dendrons can be attached to the monomer before polymerization. Most of the chitosan-dendrimer hybrids are synthesized using the attach-to route approach (Sadeghi-Kiakhani and Safapour 2016). Dendrimers and their derivatives are commonly used for the dye removal and modification of dyes and textiles

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(Sadeghi-Kiakhani et al. 2013a, b, 2015; Mahmoodi et al. 2010). Chitosanpolypropylene imine dendrimers hybrid (CS-PPI) as a biomordant can be obtained by firstly treating chitosan with acetic acid and water/methanol mixture. Then, ethyl acrylate is introduced into the mixture. After stirring at fixed temperature for 10 days, the solution is quenched and precipitated in acetate saturated with NaHCO3. After filtration, the filtrate is dispersed in H2O saturated with NaHCO3. Dialyzed against H2O and lyophilized is done to obtain N-carboxyethyl chitosan ethyl ester. NaOH solution is added in prepared compound to obtain N-carboxyethyl chitosan. This modification makes it a highly biocompatible compound (Tsubokawa and Takayama 2000). Similarly, chitosan powder grafted with hyper-branched dendritic polyamide has also been reported. This modification was achieved by repeating two processes: (i) addition of methyl acrylate to the surface amino groups and (ii) amidation of resulting ester with ethylenediamine to give polyamidoamine dendrimer-grafted chitosan powder (Jianchao and Pu 2012). Other chitosan-dendrimer hybrids that have been reported mainly include chitosan-polyamide amine, chitosan-polyethylenimine, chitosan-DOBOB acid, chitosan-DOVOB acid hybrid, chiton-polyglycerol hybrid, etc. that have also been reported (Sakti 2015). Dendrimerized magnetic chitosan for the removal of heavy metals has also been mentioned in the literatures (Buranaboripan 2014).

Cyclodextrin-Linked Chitosan The cyclomalto-oligosaccharides involve α-, β-, and γ-cyclodextrin (CD), which are important because of their ability to encapsulate hydrophobic molecules in their toroidal hydrophobic cavity whose selectivity depends on the number of glucose units. CD-modified chitosan has applications in various fields including water treatment (Alamdarnejad et al. 2013), pharmaceuticals, cosmetics, drug delivery (Wang et al. 2015b), and analytical chemistry. Thiolated carboxymethyl chitosang-cyclodextrin has been prepared by ionic gelation method. In this method, β-cyclodextrin was grafted onto carboxymethyl chitosan (CMC) in the presence of 1, 6-hexamethylene diisocyanate (HMDI) reagent, which acts as a spacer between the CD and CMC. Stannous 2-ethylhexanoate catalyst was used to promote reaction between isocyanate groups of the CMC-HMDI and β-CD. The resulted material is washed and dried (Wang et al. 2015b; Aoki et al. 2007). Similarly, three different kinds of CD-modified chitosan beads by ionic gelation method have been studied for the adsorption of nonionic surfactant 4-nonylphenol ethoxylates (NPEs) (Luzardo-Á lvarez et al. 2012). In addition, β-CD-modified chitosan has been prepared by treating with p-toluenesulfonic anhydride. After addition of NH4Cl and HCl, the resulting product was filtered and dried. This modification increases the hydrophobicity, biocompatibility, and enzymatic biodegradability of chitosan and makes adsorbent a better alternative for hydrophobic drug encapsulation (Buranaboripan et al. 2014). In another study, the preparation of α-CD and β-CD derivatives by reductive amination method using 6-deoxy-6-(4-oxobutyramido)-β-CD and 6-oxo-β-CD (Tojima et al. 1998) and mono 2-O-allyl-α-cyclodextrin has also been discussed in the literatures (Wan et al. 2002).

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Chitosan Bound with Crown Ether Crown ethers have a unique molecular structure and selective coordination ability. Cross-linking with crown ether imparts properties like tensions of multiple rings in the molecular structures, active chemical properties, and easy opening of loop under acid or alkali catalysis (Yuting et al. 2000; Ji et al. 2014). Chitosan and crown ether form stronger complexing with salt and better selectivity for metal ions because of the synergistic effect of high molecular weight (Peng et al. 2003). Two different chitosan-crown ether resins, Schiff base type chitosan-benzo-15crown-5 (CTS-B15) and chitosan-benzo-18-crown-6, are synthesized by the reaction between –NH2 group in chitosan and – CHO in 40 -formyl benzo-crown ether. This modification increases the adsorption selectivity of chitosan for Pb2+, Cu2+, and Hg2+ (Yi et al. 2007). Similarly, the chitosan-crown ether cross-link was grafted by 40 -formal benzo-15-crown-5 to obtain cross-linked chitosan with 40 -formal benzo15-crown-5(CCTS-N=CH-B-A5-C-5). After grafting it was loaded with palladium chloride to achieve the heterogeneous catalyst, which can be easily isolated from the reaction system. This process enhanced its metal ion selectivity (Jenkins and Hudson 2001). In another study, N-benzylidene chitosan (CTB) was synthesized by the reaction of benzaldehyde with chitosan (CTS). Chitosan-dibenzo-18-crown-6 crown ether (CTSD) and chitosan-dibenzo-18-crown-6 crown ether-bearing Schiff base group were prepared by the reaction with 4, 40 -dibromodibenzo-18-crown-6 crown ether, respectively. This modification improved the sorption behavior of the formed chitosan adsorbent (Yuting et al. 2000). Chemically Grafted Chitosan Grafting of chitin and chitosan is important for the functionalization and development of practically useful derivatives. Ceric ion, Fenton’s reagent, gamma irradiation, ring opening, and various radicals are the various routes of grafting (Nguyen et al. 2016). Chemically modified chitosan has wide application in the field of adsorption. For instance, for the biosorption of nickel (II) and phenol from binary mixtures on glutaraldehyde, cross-linked chitosan beads were grafted with histidine and Saccharomyces cerevisiae (SC). Glutaraldehyde cross-linked chitosan was prepared by agitating chitosan beads in glutaraldehyde solution. Immobilization of histidine (HIS) on the cross-linked chitosan beads (CCBs) resulted as HIS-CCBs was suspended in phosphate buffer and agitated with NaOH-treated SC powder. This modification improved the selectivity and adsorption capacity of the adsorbent material (Nishad et al. 2012). Ion-imprinting technique to prepare cross-linked metal complexed chitosan in which metal cation acts as a template and regulator for cross-linking has also been used for decades. Metal ion-imprinted chitosan material such as Ag+, Ca2+, Cu2+, Fe2+, Ni2+, Pb2+, Cd2+, Hg2+, Cr(VI), Co(II), etc. has been synthesized to adsorb the corresponding metal ion from aqueous solution (Shofiyani et al. 2015; Feng et al. 2013; Zhang et al. 2015a; Tang et al. 2013; Liu et al. 2013, 2015b; Ge et al. 2016). For instance, poly(acrylic acid) grafted and glutaraldehyde cross-linked chitosan nano adsorbent were synthesized by using Pb2+ as a template ion. The process starts with stirring of chitosan with acrylic acid. The mixture was heated under a nitrogen stream, and potassium persulfate was

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added. After cooling and filtration, the filtrate was again treated with glutaraldehyde. The treated product was washed several times and dried. HCl was used to remove the template from the resulted adsorbent material. This modification enhanced the adsorption properties of produced adsorbent (Sjoholm et al. 2009).

Enzymatic Modification of Chitosan Modification of chitin and chitosan by enzymatic approach is interesting owing to its specificity and environmental advantages as compared to chemical modification. With respect to health and safety, an enzyme offers the potential of eliminating the hazards associated with reactive reagents. K. H. Sjohlm et al. reported the reductive amination to hydrophobic modification of chitosan to induce a micellar structure, and the ability of the modified chitosan to immobilize enzymes has been investigated. Deacetylated chitosan was hydrophobically modified by butyraldehyde (longchain aldehyde) via reductive amination. Immobilization is performed by mixing of glucose oxidase with sodium acetate buffer with chitosan suspension dissolved in 1% acetic acid. Then, reaction mixture consisting of β-D-glucose solution and Odianiside was added to the above solution mixture. At last, peroxidase solution in cold deionized water is added in the solution. This modification increases the enzyme activity for the adsorbent material (Torres et al. 2012). Enzymatic modification of chitosan with quercetin was performed by dissolving chitosan in acetic acid. After continual stirring, pH was raised to precipitate the modified chitosan. This solid precipitate was washed by isopropanol. Finally, the modified chitosan was dried under a vacuum in a phosphorous pentoxide atmosphere. This enzymatic modification enhanced plasticity, antioxidant, and antimicrobial properties and thermal degradability of the adsorbent material (Kilinç et al. 2002). In another study, papain was immobilized by cross-linking the enzyme and chitosan by means of glutaraldehyde. Chitosan treated with acetic acid was precipitated and then washed with phosphate buffer. After stirring, resulting pellets are resuspended in Buffer I containing glutaraldehyde. The chitosan-papain conjugate was washed with Buffer I containing HCl and alone with Buffer I. This modification increases the thermal stability of the modified adsorbent (Yu et al. 2015). Modification of chitosan with hydrolytic enzymes such as cellulose (Li et al. 2012a), protease, lipase, pepsin, chitinase, chitosanase, and lysozymes is also reported in the literature (Yi et al. 2009; Peng et al. 2015; Dubeau et al. 2011; Zhang et al. 2015b; Salleh et al. 2011).

Agricultural Waste Adsorbents Agricultural wastes are by-products from the forestry and agricultural sectors. These waste materials have sufficient potential to be used as a good sorbent because of their physicochemical characteristics, inexpensive and abundance in nature, and low cost. The improper disposal of agrowastes can create a serious environmental issue; that is why the use of these wastes as an adsorbent is an appropriate way to reduce the solid waste. Agricultural wastes are generally lignocellulosic materials that consist of three main high-molecular-weight structural components that are lignin, cellulose,

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and hemicelluloses. Lignocellulosic materials also contain certain extractive structural components that have smaller molecular size (Zhu et al. 2016). Agricultural wastes are better as compared to other industrial waste for development of adsorbents because it can be used without or minimum processing (washing, drying, grinding) and thus reduce production cost. The chemical modification of these wastes offers several advantages, for instance, it carried out in a single step, performed at lower temperature, and develops a better porous structure.

Modification Through Acid and Alkali Solutions The use of agricultural waste as precursor for the production of activated carbon has increased lately. These adsorbents have commendable removal efficiency for various organic contaminants through adsorption process. Acidic modification is applied by different oxidants to increase acidity and the hydrophilic nature the adsorbent. In this view, the Nigerian waste bamboo (from a petroleum refining industry) was carbonized and chemically activated with hydrochloric acid. Carbonization was done at 400–500  C, and then the material is treated with 0.1 M HCl and kept in muffle furnace to active the sample. The acid activation increases the adsorption capacity of adsorbent (Khan et al. 2016). Walnut shells (WNS) have been chemically modified for the sorption of hydrocarbon (PAH) and fatty acids (capric acid, lauric acid, palmitic acid, and aleic acid). Surface modification is achieved via esterification, i.e., conducted in n-hexane with H2SO4 as catalyst. The – COOH group of fatty acid is esterified with the – OH group of WNS to form – OCOC group. This reaction changes the surface properties of WNS from being polar hydrophilic to organophilic. The polarity and aromaticity of the material induce considerable effect in terms of its adsorption capacity (Khan Rao and Khatoon 2016). Beside acidic activation, alkali modification could also be applied to the agrowaste material for surface modification adsorbents. This modification produces positive surface charges by which it can adsorb the negatively charged pollutant species. For instance, a plant-based adsorbent Caryota urens seeds were chemically modified with NaOH. The material was modified by treating Fenton’s reagent, sodium silicate, and NaOH (Johari et al. 2016). In the literatures, alkali-modified coconut pitch and coconut powder were reported by using NaOH. This treatment increases the surface roughness, amount of cellulose exposed on the surface, amount of amorphous cellulose at the expense of crystalline cellulose, and the removal bonding in the network structures (Afroze et al. 2016). In another study, for the removal of Zn2+, Eucalyptus sheathiana bark (agricultural by-product) has been modified with NaOH. To prepare the adsorbent, bark powder was stirred with NaOH followed by filtration and washing. The modified product increases the adsorption capacity in aqueous medium (Ramalingam Subramaniam 2015). Activated carbon from palm oil fronds and cashew nut shell have been prepared and activated by KOH which has been reported in the literatures. The procedure followed the carbonation process at 700  C and then soaked with KOH solution (Water Treatment Using Carbon Filters: GAC Filter Information 2004; Li et al. 2012b).

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Amine-Functionalized Agricultural Waste The functionalization of lignocellulosic agricultural residues with anime group provides a mild and eco-friendly alternative as it enhances the binding of functional group over surface. Raw cassava straw functionalized with epichlorohydrin is used as an adsorbent material. To prepare the adsorbent, raw material is dispersed in N, N-dimethylformamide. Aliquot of epichlorohydrin is added in the presence of pyridine and the reaction product washed with diluted ethanol. Finally, amino group is introduced into epoxypropyl by-product after reaction with dimethylamine solution and washed with ethanol. Modification increases the amino group over surface and composed an effective adsorbent for the sorption of nitrate (Kalaruban et al. 2016). Similarly, corncob and coconut copra grafted with epichlorohydrin by the abovementioned method (Kalaruban et al. 2016) have also been reported (Ezegbirika and Nnaobi 2012). Grafting of amine changed the surface charges from negative to positive, which is favorable for the nitrate adsorption by electrostatic forces (Ezegbirika and Nnaobi 2012). In another study, fabrication of groundnut husk with EDTA has been reported. For this, the powder groundnut husk was steeped in nitric acid solution and hydrolyzed with sulfuric acid. Finally the mixture was filtered and the hydrolyzed husk was heated with pyridine and EDTA. The incorporation of EDTA improved the binding metal ion capacity since it provides the amino group for exchangeable metal ions (Hena et al. 2015). Modification of sugarcane bagasse with epichlorohydrin and triethylamine in the presence of ethylenediamine and N,N-dimethylformamide has been investigated. The reaction between epichlorohydrin and cellulose was induced after the activation of hydroxyl group of cellulose molecule, producing hydroxyl cellulose ether. It is then cyclized by the catalyst to produce epoxy cellulose ether which is used as an intermediate in the reaction. The modified product was obtained after the grafting reaction between epoxy cellulose ether and triethylamine. As a result, organic monomers were embedded in the cellulose skeleton, which resulted in a broad network and increased the porous structure on the surface of the product (Bayramoglu et al. 2013). The change in adsorption potential of amine-modified plant biomass of Alyssum caricum has also been reported. The sorbent, the native biomass, is soaked in NaOH and filtered and immersed in epichlorohydrin. The resulted mass was washed with phosphate buffer and grafted with hexamethylenediamine (HMDA). The surface modification so obtained increases the adsorption of acidic dyes (reactive green 19 and reactive red 19) (Goncalves et al. 2013). Zinc Chloride-Modified Agrowaste Material Activation of agricultural waste materials with zinc chloride can be very effective in adsorbent development. In the literature, activated carbon from coffee husk waste prepared by chemical activation has been reported. Coffee husk was impregnated with activating agent ZnCl2, and the activation is carried out in tubular furnace. Further iron oxide (goethite) was aged in the solution of FeCl3 and NaOH to get precipitate on the surface of activated carbon. This process dispersed the iron oxide all over the surface and enhances the catalytic activity (Itodo et al. 2010). In another

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study, four substrates, viz., groundnut shell, shea nut shells, poultry wastes, and poultry dropping-based activated carbon, were chemically activated with ZnCl2 and H3PO4. In this process, the agrowaste samples were washed and carbonized. Afterward, these samples were separately mixed with the activating agents and kept into the ice bath. The activated carbon so generated was washed with HCl (Matichenkov and Bocharnikova 2001). Tomato-based activated carbon, modified/activated with Zncl2, has been studied for tetracycline removal. The adsorbent was prepared by aging the carbonized tomato sorbents with zinc chloride solution. It was observed that this fabrication improves the pore structure of carbon adsorbent (Salman 2014).

Soil/Silica Adsorbents Soil, as a natural adsorbent, consists a mixture of minerals, organic matter, gases, liquids, and various organisms. Soil minerals and organic matter control its physical and chemical properties (Rosales-Landeros et al. 2013). Important functions of soil are to provide a growth medium for plants, means of water storage, supply, and purification. Silica (SiO2) is a very common mineral, and free silica occurs in many crystalline forms. Its most common form is quartz. Its composition is very close to pure silicon dioxide: 46.75 w% silicon and 53.25 w% oxygen (Priya 2014). Modification onto their surface affects their adsorption properties by increasing the efficiency, sensitivity, and selectivity. Some of modification techniques are discussed here below.

Modified Soil Concurrent contamination of soil with organic matter and heavy metals has become an environmental problem. Through adsorption by soil, these pollutants can be minimized from their harmful effect by reducing their movement in soil and thus reduce their potential for plant uptake or groundwater contamination. In the literature, soil has been modified with sulfuric acid to activate the soil surface for the removal of textile effluent. This can be prepared by washing, drying, crushing, and sieving the soil sample. At last, the soil was modified by treating it with sulfuric acid and stirred for half an hour. This modification due to the presence of aluminosilicates is anticipated in the color and COD removal and also increases the adsorption efficiency of the adsorbent (Meng et al. 2009). Lou soil on both layers tillage and clay was modified with cetyltrimethylammonium bromide (CTMAB), and sodium dodecylsulphonate to study the adsorption kinetics of phenol has been reported. This modification increases the adsorption rate of phenol on the surface of adsorbent (She Xiao-yan et al. 2015). In another study, soft rock addition to sandy soil for phosphorous adsorption has also been mentioned (Sharma et al. 2013). Ligand-Modified Silica Gel Silica gel is a porous, granular form of silica tetrachloride or substituted chlorosilane/ orthosilicate. Chemical adsorption on silica surface provides immobility, mechanical stability, and water insolubility. Silica gel surface can be modified by two following

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distinct processes: organo-functionalization, in which the modifying agent is an organic group, and inorgano-functionalization, where the group anchored on the surface can be organometallic composite or a metal oxide. The most convenient way to prepare a chemically modified silica gel is achieved by simple immobilization of the group on the surface by adsorption or electrostatic interaction or other type of interactions (Priya 2014). In this view, (HBAPE) modified silica gel is prepared by, firstly, the reaction between the silylating agent (APTES) and the silanol groups on the surface of silica gel to obtain aminopropylated silica gel (APSG). After it, functionalization of APSG with HBAPE is performed. This modification increases the sorption capacity level and selectivity of the sorbent (Das 2010). In another study, silica gel modified with 4,40 -diaminodiphenylsulfone (DDS)- O-hydroxybenzaldehyde (O-HB) is achieved by the reaction of 3-chloropropylsilica gel with DDS. The product (Si-DDS) is further reacted with (O-HB) to obtain the desired material (Si-DDS-O-HB) (Radi et al. 2014). Another modification of silica gel doped with 3-aminopropyltrimethoxysilane via 1H-pyrrole-2-carbaldehyde has been reported. The adsorbent is prepared by mixing of aminopropyltrimethoxysilane into the suspension of activated silica gel to produce 3-aminopropylsilica. The product so obtained reacts with pyrrol-2-carbaldehyde to produce the final adsorbent material. This modification process enhances the sorption of metal ions Cu, Zn, and Cd onto the surface of adsorbent (Fan et al. 2007). A numerous organic molecules have been immobilized onto silica surface which include xylenol orange (Akhond et al. 2006), di(n-propyl) thiuram disulfide (Amarasekara et al. 2009), 4-acylpyrazolone (Ngeontae et al. 2007), aminothioamidoanthraquinone (Goswami and Sing 2002), and 1,8-dihydroxyanthraquinone (Ebrahim 2013).

Metal-Organic Framework Metal-organic frameworks are a class of porous crystalline material that consists of metal ion clusters and multidentate organic ligands. During the past two decades, MOFs have shown considerable applications in many areas such as gas storage, catalysis, gas separation, sensing, light harvesting, and optical luminescence due to their tunable crystalline hybrid network, high porosity, and rich functionality. Besides the advantage of their textual properties, modulation of their chemical properties is also an interesting feature of these materials. After functionalization the incorporated functional groups improve interaction of material with incoming molecules. Some of the common approaches used for the MOF modification are being discussed herein below:

Metal Doping in MOF Structures The doping of metals in the MOF framework increases the uptake of guest molecules by providing additional sites for adsorption. For instance, Ce(III)-doped zirconiumbased MOF (UiO) was developed by dissolving zirconium tetrachloride (ZrCl4) and benzene-1,4-dicarboxylic acid (BDC) in N,N-dimethylformamide (DMF). Then, CeCl3.7H2O is stirred in DMF and transferred into ZrCl4/BDC solution and stirred.

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The reagents are then added with DMF. The system was heated and the product formed is filtered, washed, and immersed in a dichloromethane solution for solvent exchange (Lee et al. 2014). In another approach, copper-based MOF doped with TiO2-carbon nanotube was prepared for sensitization by using layer-by-layer technique; hole blocking layer-coated fluorine-doped tin oxide (FTO) was dipped into trimesic acid in DMF solution. Then, acid-functionalized FTO substrate is dipped into Cu (NO3)2.3H2O in DMF solution followed by washing and drying (Anbia and Pazoki 2015). Similarly, doping of lithium (Li) in Cu-based MOF was performed by thermal decomposition in anion species method. In this process, vacuum-activated copper carboxylate (Cu-BDC) is immersed in LiNO3/ethanol solution. After filtration, LiNO3-doped Cu-BDC is heated in vacuum for the removal of nitrate anion and thus obtains Li-doped MIL-53 (Adhikari and Lin 2016). Likewise, MOF-74(Ni) and MOF-74(Co) were post-modified with Pd-incorporated activated carbon. The carbonation bridge was formed by the carbonization of sucrose (Botas et al. 2010). Doping of Co, Cu, and Mg into Zn-based MOF-74 has been prepared by using nitrate salts as metal sources with 2,5-dihydroxybenzene-1,4-dicarboxylic acid (Zhang et al. 2016b). Coumarinloaded Zn-MOF-74 was prepared by mixing activated Zn-MOF-74 in a solution containing n-hexane and coumarin. N-hexane was evaporated and the resulting material was washed by ethanol and dried under vacuum (Brozek et al. 2015). Similarly, numerous MOFs doped with other metals like Na, K, Li, Ni, Co, Mg, etc. have also been reported in the literature (Botas et al. 2010; Mulfort et al. 2009; Chen et al. 2014).

Functional Groups over MOFs Post-synthetic modifications of MOFs with functional group while maintaining the lattice structure have also proven to be a good approach for modification. In this view, amine functionalization of MOFs has been modified by alkylation with 2-dimethylaminoethyl chloride. This basic amino-functionalized material had shown significant change in catalytic performances (Gadzikwa et al. 2008). In another study, azide-functionalized Zn-cornered mixed ligand MOF was formed by reaction with ethidium bromide monoazide (Ko and Kim 2011). For the sorption of CO2, four mesoporous UMCM MOFs (1) were post-modified with acetic, butyric, benzoic, and 4-trifluoromethyl phenyl anhydrides. The materials were prepared by dissolving synthesized material into the solutions of acetic anhydride, butyric anhydride, benzoic anhydride, and 4-trifluoromethyl phenyl anhydrides with chloromethane, respectively. This decoration of functional groups over resulted product affected the pore environment of the cages and hence increases the affinity toward sorption (Deng et al. 2010). A multivariate MOF-5 which contains eight distinct functionalities (-NH 2, -Br, -Cl2, -NO2, (CH3)2, -C4H4, -(OC3H5)2, and -(OC7H7)2) in one phase has been reported in literature. Crystals of MTV-MOFs were obtained by adding Zn(NO 3)2.4H2O to a N,N-dimethylformamide solution mixture of the acid forms of the selected organic links (Abid et al. 2013). In another study, alkylamine-functionalized

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Zr-MOF and Mg-MOF are prepared by adding the synthesized MOF into the solution of DMF (Wu et al. 2015; Yang et al. 2016c).

Future Prospects of Modification Techniques In the recent years, natural adsorbents have proven their appreciable performances in various areas and industries. Different fabrication approaches reviewed via different functional groups, metal oxides, multidentate compounds, etc. increase efficiency and surface properties of the materials. These natural materials have wide application in the areas of biomedical science, building constructions, photocatalytic degradation, gas purification, separation and storage, pharmaceuticals and cosmetics, solvent vapor recovery, etc., and progressive increase in their applications is expected. In this view, for the gas transport and storage in the shale matrix, natural porous clay material has been studied. Gas molecules dynamically adsorb and desorb simultaneously on the adsorption sites (Mann et al. 2016). Radiation leakage accident at Fukushima, Japan (2011), calls attention to the awful need of studies for the durability and effectiveness of emerging materials used for radiation shielding purpose. As a result, safe storage of nuclear waste materials has become an important issue, globally. For this, bricks have been produced from the mixture of clay and borogypsum. This modification increases the comprehensive strength of brick clay (Eprikashvili et al. 2016). Worldwide agricultural practices demonstrate that among various agronomical factors, fertilizers are of supreme importance in increasing the soil fertility and crop yield. In this prospect, combination of natural zeolite (clinoptilolite) as a natural mineral fertilizer with brown coal as a soil substitute has been prepared for transferring organic substances in the brown coal into the matter adorable by plants. This natural material-based fertilizer increases the productivity of crops and improved the physical and chemical properties of soil (Zettl et al. 2015). Similarly, zeolite A4 impregnated with LiCl/MgSO4 and MgCl2/MgSO4 has been developed for open sorption storage process for space heating and hot water. The impregnation method includes wetting of the zeolite by an aqueous mineral salt followed by drying (Elsayed et al. 2017). In most of the reported studies for the adsorption, desalination of silica gel as adsorbent had been used. It suffers from limited water uptake capabilities leading to a low system performance. To overcome this problem, CPO-27(Ni) and aluminum fumarate MOFs are commercially available. In the view of their increasing applications in various fields, natural materials can be considered as a green cleaner and problem-solver.

Cross-References ▶ Environmental Nanotechnology ▶ Role of Bioremediation as a Low-Cost Adsorbent for Excessive Fluoride Removal in Groundwater

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Advanced Pretreatment Strategies for Bioenergy Production from Biomass and Biowaste

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C. Veluchamy, Ajay S. Kalamdhad, and Brandon H. Gilroyed

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . The Composition of Lignocellulosic Biomaterial . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Limiting Factors for Biodegradation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Bioenergy Production by Fermentation and Anaerobic Digestion . . . . . . . . . . . . . . . . . . . . . . . . . . . . Pretreatment Strategies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Physical Pretreatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chemical Pretreatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biological Pretreatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cross-References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

Production of biowaste is increasing day by day, due to the abundant availability of raw materials and their utilities, especially from lignocellulosic-based material that constitutes 50% of the total biomass around the world. The recalcitrant structure of lignocellulose material has become resistant for their degradation

C. Veluchamy (*) Department of Civil Engineering, Indian Institute of Technology Guwahati, Guwahati, India School of Environmental Science, University of Guelph Ridgetown Campus, Ridgetown, ON, Canada A. S. Kalamdhad Department of Civil Engineering, Indian Institute of Technology Guwahati, Guwahati, India e-mail: [email protected] B. H. Gilroyed School of Environmental Science, University of Guelph Ridgetown Campus, Ridgetown, ON, Canada © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_45

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and their recovery of valuable material. Many factors like lignin content, crystallinity of cellulose, and particle size limit the digestibility of hemicellulose and cellulose present in the lignocellulose material. To overcome the bottleneck, several pretreatment strategies have been employed to diminish the rate-limiting steps. They are generally classified as physical, chemical, and biological pretreatment. The physical pretreatments include mechanical (comminution, ultrasound, mechanical jet, high-pressure homogenizer), and thermal pretreatments (liquid hot water, autoclave, electrohydrolysis, microwave oven). The chemical pretreatments include acid, alkali, organosolv, wet oxidation, and ozonolysis pretreatment. The biological pretreatments (fungi, microbial consortia, enzyme pretreatment) or combined pretreatments (thermochemical or alkali thermochemical) are often employed nowadays. This chapter reviews the process description, mode of action, and challenges of several pretreatments. This chapter also discusses the future research needs of advanced pretreatment strategies for bioenergy production from biomaterial. Keywords

Lignocellulose · Pretreatment strategies · Bioenergy · Limiting factors · Biodegradation

Introduction Increasing global energy demands and the threat of climate change due to greenhouse gas (GHG) emissions have flashed the urgent need to develop sustainable and affordable environmental-friendly energy resources. Worldwide, energy consumption reached 524 QBtu in 2010, and it is estimated to reach 800 QBtu by 2040, with an average growth rate of 1.5% per year. A large fraction of the world’s total energy demands, more than 84%, is supported by nonrenewable fossil resources such as coal, oil, and natural gas. These are not only in limited supply but also cause the GHG emissions into the atmosphere (EIA 2013). Bioenergy produced from the renewable feedstocks such as biomass and its associated waste is considered to be one of the promising alternatives to fossil-derived energy due to several inherent and significant advantages (Sawatdeenarunat et al. 2015; Veluchamy and Kalamdhad 2017a). Many biodegradable feedstocks, such as wastewater, sewage sludge, food waste, animal manure, agri-residue, and organic fraction of municipal solid waste, are used as substrates for bioenergy production. These facilities elucidate the unique potential for bioremediation and waste stabilization with concurrent bioenergy production. In recent years, lignocellulose biomass and its associated waste such as energy crops and agri-residues have gained much attention as essential feedstocks for the production of bioenergy and bio-based products (Veluchamy and Kalamdhad 2017e). These lignocellulose feedstocks do not directly compete with the food or feed production, like conventional biorenewable feedstocks such as sugar and starch-based crops. In addition to that, high biomass yield even under low inputs

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Biogas (CH4+CO2) Anaerobic digestion

Forest & agri-residues Biomass and biowaste

Pretreatment strategies

Hydrogen (H2)

Sugars Ethanol Fermentation

Energy crops

Chemicals (carboxylicacids, alcohols, ethers)

Chemicals Lignin and residues

Combustion (syngas, heat, electricity)

Fig. 1 Overview of integrated biorefinery approach for lignocellulose materials

of energy, water, and fertilizers makes these feedstocks ideal for bioenergy production (Sawatdeenarunat et al. 2015). It is expected that the present fossil-based petrochemical industry gradually will be replaced by biorefineries which produce a number of valuable products from lignocellulose materials such as gases, alcohols (fuels), heat, and electricity (Fig. 1).

The Composition of Lignocellulosic Biomaterial Lignocellulosic biomass is an abundant organic resource with an annual yield of over 200 billion dry metric tons per year (Kumar et al. 2008). The basic structure of lignocellulose is comprised primarily of cellulose (35–60%), hemicellulose (20–35%), and lignin (10–25%) (Zheng et al. 2014) along with smaller quantities of other organic and nonorganic compounds such as proteins, lipids, and other extractives (Frigon and Guiot 2010). The carbohydrate components (cellulose and hemicellulose) are fermentable after hydrolysis, which makes lignocellulosic biomass a suitable feedstock for the bioenergy production. However, the inherent characteristics of native biomass and its biowaste, such as structural and chemical properties, make it resistant to biodegradation by enzymes and microbes. Some commonly used lignocellulosic feedstocks are summarized in Table 1. The amount of these constituents not only varies within the species but also varies due to growth conditions and maturation. Cellulose in a plant consists of parts with a crystalline (organized) structure and parts with a not well-organized (amorphous) structure. Cellulose, the major constituent of plant cell wall, consists of β-1,4-linked D-glucose units that form linear polymeric chain of about 8000–12,000 glucose units. Cellulose consists of chains that are packed together by hydrogen bonds to form highly insoluble structures called microfibrils (de Vries and Visser 2001).

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Table 1 The composition of lignocellulose content in biomass and biowaste Biomass Corn stover Corn cobs Wheat straw Switch grass Bagasse Sugarcane Rice straw Eucalyptus Giant reed stalk Giant reed leaves Sunflower stalk Sorghum Barley straw Rye straw Napier grass Coastal Bermuda grass

Cellulose (%) 37.5 45.0 38.2 31.0–45.0 38.2 25.0 32.0 38.0–45.0 33.1 20.9 31.0 22.2 37.5 38.0 45.7 25.0

Hemicellulose (%) 22.4 35.0 21.2 20.0–31.0 27.1 17.0 24.0 12.0–13.0 18.5 17.7 15.6 19.4 25.3 36.9 33.7 35.7

Lignin (%) 17.6 15.0 23.4 12.0–18.0 12.2 12.0 13.0 25.0–37.0 24.5 25.4 29.2 21.4 26.1 17.6 20.6 6.4

Source: Adapted from Sawatdeenarunat et al. (2015)

Hemicellulose, the second most abundant heterogeneous polysaccharides, consisted by the different units of sugars in the plant cell wall. These are classified according to the main residue of sugars present in the backbone of the structural polymer. Xylan, the most abundant hemicellulose polymer, is composed by β-1,4-linked D-xylose units in the main backbone, which can be substituted by different side groups such as D-galactose, L-arabinose, glucuronic acid, acetyl, feruloyl, and p-coumaroyl residues (de Vries and Visser 2001). The other two major hemicelluloses in plant cell wall are galacto(gluco) mannans, which consist of a backbone of β-1,4-linked D-mannose and D-glucose residues with D-galactose side chains, and xyloglucans that consist of a β-1,4-linked D-glucose backbone substituted by D-xylose. The short and branched chains of hemicelluloses help built a network with cellulose microfibrils and interact with lignin, rendering the cellulose-hemicellulose-lignin matrix extremely rigid. Lignin is a large and complex aromatic phenolic polymer that confers strength to the plant cell wall. Lignin is a highly insoluble complex polymer of phenylpropane units that are joined together by ether and carbon-carbon linkages, forming an extensive cross-linked network within the plant cell wall. The cross-linking between the different polymers confers the complexity and rigidity, which is responsible for the protection of plant cell wall as a whole. In addition to that, it offers protection against mechanical stress and osmotic lysis, an effective barrier against pathogens. These properties of lignin make it a more recalcitrant component of the plant cell wall. The higher the lignin content, the greater the resistance of biomass to chemical and biological degradation.

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Limiting Factors for Biodegradation Lignocellulose biomass represents a rather unused source for the bioenergy production. This is because of many limiting factors such as lignin content, crystallinity of cellulose, degree of polymerization, moisture content, available surface area, and the particle size that limit the digestibility of the hemicellulose and cellulose present in the material (Veluchamy et al. 2017). Decreased particle size with increase in available surface rather than crystallinity affects the rate and extent of the hydrolysis stage in biodegradation. Few researchers conclude that the pore size of the substrate in relation to the particle size of the enzymes is the main limiting factor in the enzymatic hydrolysis of lignocellulose biomass. Lignin acts as a shield that limits the rate and extent of enzymatic hydrolysis, preventing the digestible portions of the substrate to be hydrolyzed (Chang and Holtzapple 2000).

Bioenergy Production by Fermentation and Anaerobic Digestion The production of bioenergy from lignocellulose material consists of few phases, namely, pretreatment, anaerobic digestion and fermentation, product separation, and posttreatment (Fig. 1). A product separation step is not needed for the anaerobic digestion, because biogas (methane and carbon dioxide) could itself separate from the digestate liquid fraction under normal condition. The pretreatment is necessary to improve the rate of production and the total yield of monomeric sugars in hydrolysis step. Anaerobic digestion is among the most promising bioconversion technologies for production of renewable bioenergy. However, the production of methane alone may not sufficiently justify the capital and operating cost associated with a commercial biogas facility. Hence, anaerobic digestion could be integrated into a biorefinery that facilitates the subsequent breakdown of lignocellulosic biomass into its constituents: sugars (glucose, galactose, xylose, arabinose, and mannose) and/or volatile fatty acids (acetic, propionic, butyric, and valeric acids). The produced monomeric hexoses can be fermented to ethanol easily, while the fermentation of pentoses is only done by few strains. Volatile products are also not fermented to ethanol. The solubilized components could be used as precursors for the production of diverse products ranging from bioenergy (methane, hydrogen), biofuels (ethanol, methanol, and butanol) (Agler et al. 2011; Rabelo et al. 2011), and organic acids such as succinic acid (Cherubini 2010).

Pretreatment Strategies Due to the complexity and variability of lignocellulose biomass chemical structure, the optimal pretreatment method and its associated conditions depend on the types of lignocellulose content. Several structural and compositional properties were found to have impacts on biodegradability. In general, different pretreatment methods such as physical, chemical, and biological techniques have been proposed as shown in

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Fig. 2. However, all methods had a major effect on the accessible area of lignocellulose biomass. In order to improve bioenergy production, an efficient pretreatment is necessary to disrupt the naturally recalcitrant carbohydrate-lignin shields that impair the accessibility of enzymes and microbes to cellulose and hemicellulose (Yang and Wyman 2008; Veluchamy and Kalamdhad 2017d). The goal of the pretreatment process is to remove lignin and hemicellulose, reduce the crystallinity of cellulose, and increase the porosity of the lignocellulosic

Fig. 2 The pretreatment strategies applied for the lignocellulose materials

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biomass. Pretreatment must meet the following requirements: improve the formation of sugars or the ability to subsequently form sugars by hydrolysis, avoid the degradation or loss of carbohydrate, avoid the formation of bioproducts that are inhibitory to the subsequent hydrolysis and fermentation processes, and be costeffective.

Physical Pretreatment Physical pretreatment does not use the chemicals or microorganisms during the pretreatment processes. Physical pretreatment is further grouped into mechanical and thermal pretreatment.

Mechanical Pretreatment In mechanical pretreatment, the lignocellulose biomass could be comminuted into smaller piece by numerous mechanisms such as chipping, grinding, and milling. Mechanical pretreatment is done primarily to disrupt the crystallinity of cellulose (not applicable for biosludge) by breaking into smaller components resulting in higher specific surface area and reduced degree of polymerization, thus rendering the lignocellulose biomass more amenable to succeeding enzymatic hydrolysis. On the other hand, mechanical pretreatment is energy-intensive and expensive, moreover time-consuming. Mechanical pretreatment is lesser effective than chemical pretreatment because it does not remove the lignin content (Zheng et al. 2014). Moreover, it has been observed that significant restriction in the cellulose accessibility also inhibits the cellulose enzyme. Therefore, mechanical pretreatment method was used occasionally.

Comminution (Mechanical Ball Mills) Comminution is a mechanical pretreatment of lignocellulose biomass that can be applied to reduce the cellulose crystallinity (cutting the lignocellulose biomass into smaller pieces through a combination of processes such as chipping, grinding, and/or milling). The comminution of lignocellulose biomass can be accomplished by using the ball, vibro, hammer, knife, two-rolled, colloid, and attrition mills. The size of the materials is usually 10–30 mm after chipping and 0.2–2 mm after milling or grinding (Sun and Cheng 2002). The selection of a comminution machine depends on the feedstock moisture content. Comminuting dry biomass with moisture contents ranging from 10% to 15% (wet basis) can be done by two-roll, hammer, attrition, and knife mills, while colloid mills and extruders are suitable for comminuting wet materials with moisture content more than 15–20% (wet basis), whereas the ball and vibro mills can be used for either dry or wet materials (Kratky and Jirout 2011; Eisenlauer and Teipel 2017). Reduction in particle size can alter the inherent ultrastructure of lignocellulosic biomass, increase the accessible surface area, and reduce the degree of polymerization for improved digestibility (Eisenlauer and

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Teipel 2017). It was found that vibratory ball milling is more effective than ordinary ball milling in the case of reducing the crystallinity of cellulose of lignocellulose materials and improving their digestibility after pretreatment. High-Pressure Homogenizer There are a number of pretreatment technologies that incorporate the use of highpressure gradients to rupture the cell wall of lignocellulose biomass. As the sample approaches the impacting rings or collision plates, there is a rapid increase in the velocity with a decrement in pressure, resulting intense energy release outcomes in severe turbulence and localized pressure differences that lead to the formation of cavitation bubbles. As the sample approaches the impacting rings or collision plates, there is a rapid increase in the velocity with a decrement in pressure. Hence, these developed forces tear the cell wall and release the cellular cytoplasm in lignocellulose biomass. One of the large-scale studies was conducted, where the municipal aerobic sludge was disrupted by jetting the sludge to collide with a collusion plate at 30 bar pressure. The homogenization process involves four steps while applying in the industrial level. The first step is the addition of alkalinity to weaken biomass cell membranes and to decrease the viscosity. The alkaline biomass (pH 8.5–10) is kept in a holding tank for a stipulated time. The second step is to mechanically disintegrate the biomass floc into relatively fine particles. Then it is homogenized at high pressure up to 12,000 psi. Electrohydrolysis (Pulsed-Electric Field) Electrohydrolysis pretreatment involves application of short burst of high voltage to a sample placed between the electrodes (Veluchamy et al. 2017). This is also named as pulsed-electric field pretreatment. The principle of electrohydrolysis pretreatment relies on electrophoresis, ohmic heating, and electroosmosis resulting in the disintegration of particles and microbial cell lysis (Zhen et al. 2014). The electrophoresis is the process in which shifting of ions is relative to a static phase depending on its electrical charge and molecular size (Mahmoud et al. 2010). The process in which thermal energy is generated by passing electric current through organic materials is known as ohmic heating (Varghese et al. 2014). The motion of solid particles suspended in a liquid, under the influence of an electrical field, is known as electroosmosis (Mahmoud et al. 2010). Electrohydrolysis is the process of passing direct current (DC) through an ionic substance to solubilize the organic matter by breaking the bonds between polymers induced by application of current through electrodes. An electrode when connected to a DC, one electrode becomes a positively charge and another one becomes a negatively charged electrode. This initiates the movement of electrolyte toward electrodes, i.e., positive ions move to cathode and negative ions move to anode. When an electric field is generated between the parallel plate electrodes, the field strength (E) is given by E = V/d, where V is the voltage and d is the distance between two electrodes. Normally in electrohydrolysis pretreatment, the high field strengths in the range of 5–20 kV/cm are used to significantly rupture the lignocellulose biomass. When a high-intensity external electric field is applied, a critical electrical

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potential is induced in the biomass cell membrane that leads to breakdown and structural changes of the cell wall of the lignocellulose biomass. This could result in a dramatic increase in mass permeability and mechanical rupture of the cell wall of the lignocellulose materials. By applying high-strength electric pulses, this can create pores in the cell membrane and hence facilitate the entry of acids or enzymes to break down the holocellulose into its constituent sugars. The advantage of this pretreatment is that it can be carried out at ambient environment conditions and the energy usage and its cosumption is low. This is because of the few seconds of pulse times. This pretreatment can affect the structure of the lignocellulose biomass. Ultrasonic Ultrasonic pretreatment is one of the physical pretreatment methods. The reaction involved in this pretreatment is the rapid collapse and expansion of the microbubbles caused by the localized high-temperature and high-pressure gradient in the liquid phase that could rupture the cell membrane, releasing the intercellular matter to the bulk solution. This pretreatment could be accomplished by the exposure to high-frequency sound waves generated by vibrating probes, commonly known as a horn. Ultrasonic pretreatment uses the high-intensity ultrasound ranging between 20 and 40 kHz. This pretreatment can disrupt cell wall structure, increase the specific surface area, and reduce the degree of polymerization, leading to increased biodegradability of lignocellulose biomass materials. The rupturing of cell wall occurs when the local pressure in the aqueous phase is reduced below the evaporating pressure, resulting in the formation of gas bubbles. Recently several studies have been conducted on the use of ultrasound to pretreat the lignocellulose materials for bioenergy production. This pretreatment could generate monolithic cavitations that resulted in physical and chemical effects, causing the destruction in the cell wall structure of lignocellulose biomass. The physical effects are due to the collapse of cavitation bubbles, which in turn produce an elevated alteration in the chemical structure through the formation of free radicals (Kumakura and Kaetsu 1983; Saha et al. 2011). The factors such as viscosity, background pressure, solid content, and solubilized gas concentration can influence the degree of cavitation produced by sonication pretreatment (Roxburgh et al. 2005). The usage of sonication pretreatment has been demonstrated in laboratory and in full scale level treating the municipal wastewater, but limited application for the lignocellulose biomass. Thermal Pretreatment In thermal pretreatment, lignocellulose biomass is rapidly heated by different heating sources such as hot air, liquid hot water, microwave irradiation, and steam over a period of stipulated time to encourage the holocellulose hydrolysis with rapid decompression. The biomass is subjected to temperature in the range of 150–200  C, although lower temperature has also been reported. The pressure adjoining these temperatures is in the range of 600–2,500 kPa. The major key factors for thermal pretreatment are the treatment time, exposing temperature, moisture content, and size of the particle. Thermal pretreatment is highly effective in

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increasing the accessible and susceptible surface area of cellulose and enhancing the degradability of cellulose to microbes and enzymes (Veluchamy and Kalalamdhad 2017; Veluchamy and Kalamdhad 2017c). All studies report a positive impact of thermal pretreatment on anaerobic digestion. Liquid Hot Water This pretreatment is used occasionally throughout the world. This is one of the hydrothermal pretreatment technologies that does not use chemicals and is used mostly in industrial level for bioethanol production. This pretreatment uses water at elevated temperature at high pressures to maintain its liquid form in order to promote disintegration and separation of lignocellulosic matrix (Negro et al. 2003; Rogalinski et al. 2008). Temperatures can range from 100  C to 250  C (depending on the type of sugar formation) and the length of time also ranging from few minutes to an hour (depending on the amount of sugar formation) (Yu et al. 2010a). This process uses many of the same features of steam explosion pretreatment, primarily autohydrolysis without the rapid decompression. This pretreatment is to completely solubilize hemicellulose and separate it from the rest of the solid material while reducing the formation of inhibitors. To avoid the formation of inhibitors, the reaction pH should be kept between 4 and 7 during the pretreatment. Maintaining the pH between this range could minimize the formation of monosaccharides and degradation products that can further catalyze hydrolysis of the cellulosic material during pretreatment (Weil et al. 1997, 1998; Mosier et al. 2005; Veluchamy and Kalamdhad 2017b; Kohlmann et al. 1995; Mosier et al. 2005a). The generation of reactive cellulose fibers for the production of pentosans as well as disruption of the entire lignocellulose matrix is achieved through cell penetration of the biomass by water, along with solubilization of both hemicellulose and lignin by this liquid hot water acting as an acid (Yu et al. 2010b). Autoclave (Steam Pretreatment) During this pretreatment the lignocellulose biomass is pretreated in a vessel and steam with high temperature (up to 240  C) and pressure for a stipulated time duration. After a set time, steam in a vessel is released, and the biomass is quickly cooled down. This pretreatment is to solubilize the hemicellulose to make the cellulose better accessible for enzymes or microbes and to avoid the formation of inhibitors. In this process, the in-situ formed acids catalyze the process itself and so called autoclave steam pretreatment. The role of the formed acids is not to catalyze the solubilization of hemicellulose but to catalyze the hydrolysis of the soluble hemicellulose oligomers. The optimal hemicellulose solubilization and hydrolysis can be achieved by either high temperature and short residence time (270  C, 1 min) or lower temperature and longer residence time (190  C, 10 min). Depending on the biomass moisture content, the pretreatment time has been determined. The higher the moisture content, the longer the pretreatment times and vice versa. Steam explosion pretreatment is the most commonly used pretreatment method for the lignocellulosic biomass materials. In this pretreatment, biomass could be pretreated with a high-pressure saturated steam, at he end of the processes pressure is

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suddenly reduced resulted in explosive decompression of the biomass materials. The pretreatment uses the temperature of 160–260  C (corresponding pressure 0.63–4.83 MPa) for several seconds to a few minutes before the material is exposed to atmospheric pressure. The biomass in a vessel and a steam mixture is held for a period of time to promote hemicellulose hydrolysis, and the process is terminated by an explosive decompression. The rapid flashing to atmospheric pressure and turbulent flow of the material cause fragmentation of biomass material, thereby increasing the accessible surface area (Cara et al. 2008; Viola et al. 2008) This pretreatment causes hemicellulose degradation and lignin transformation because of high temperature, thus increasing the potential of cellulose hydrolysis. Lignin is removed only to a limited extent during this pretreatment, but it is redistributed on the fiber surfaces as a result of melting and repolymerization/depolymerization reactions (Kabel et al. 2007). The factors such as residence time, temperature, chip size, and the moisture content could affect the steam explosion pretreatment. The difference between the steam pretreatment (simple stream) and steam explosion pretreatment (high-pressure saturated steam) is the quick depressurization and cooling down of the biomass in simple pretreatment. Whereas in steam explosion pretreatment causes the water in the biomass to explode at the end of the pretreatment process. Microwave Oven Pretreatment through irradiation such as microwave, gamma-ray, and electron beam radiation has been used to improve the ethanol and biogas production from various sludge, but little research has been done on the application of them for the lignocellulose biomass. Microwave pretreatment is the most studied irradiation methods. In this pretreatment, energy is generated by an electromagnetic field and delivered directly to the material in order to provide rapid heating throughout the material with reduced thermal gradient. The microwave irradiation and the dielectric response of a material determine its ability to be heated with microwave energy (Eskicioglu et al. 2007). This pretreatment can more rapidly heat a large volume, reducing the treatment time that could lead to considerable energy savings. Similar to other traditional thermal pretreatments, this could have the side effect of producing heatinduced inhibitors such as phenolic compounds, furfural, and HMF. As a result, it is important to optimize the pretreatment conditions to avoid the formation of these inhibitors. Hence, this could not be used individually for lignocellulosic biomass pretreatment, but it has been used to provide heat for assisting acid or alkaline pretreatment at relatively low temperatures without compromising the effects. With the exception of microwave pretreatment, irradiation methods are usually expensive and have difficulties and challenges in handling large volume, process scaling up, and unreliable operation in industrial level.

Chemical Pretreatment Chemical pretreatment refers to the use of chemicals such as acids and bases and the ionic liquids, to alter the physical structure and chemical characteristics of

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lignocellulose biomass. Among the three categorized pretreatments, chemical pretreatment is one that received the most recent research interest. Although many chemical pretreatment methods have been studied for ethanol production, in recent years it has been applied to biogas production.

Acid Pretreatment Pretreatment of lignocellulose biomass with acids (dilute or strong aids) at ambient temperature is done to enhance solubilization of hemicellulose, by this making the cellulose more easily accessible for the enzymes. The main reaction that occurs during the pretreatment is the hydrolysis of hemicellulose, especially xylan as glucomannan that is relatively acid stable. During acid pretreatment, solubilized hemicellulose could be subjected to hydrolytic reactions producing monomers, furfural, HMF, and other volatile products in acidic environments (Ramos 2003). Solubilized lignin will quickly condensate and precipitate in acidic environments (Liu and Wyman 2003). The solubilization of hemicellulose and precipitation of solubilized lignin are more pronounced during strong acid pretreatment compared to dilute acid pretreatment (Hendriks and Zeeman 2009). The advantage of acid pretreatment is making the cellulose more easily accessible for the enzymes due to the solubilization of hemicellulose. However, there is some risk on the formation of volatile degradation products, and this carbon may not be used efficiently in the ethanol conversion. However, these volatile products can be converted to methane. Hence, strong acid pretreatment is attractive for methane production and not for ethanol production, because of the formation of inhibiting compounds. However, methanogens can handle these compounds like furfural and HMF to a certain concentration with an acclimatization period. Nevertheless, the soluble lignin component is risky and often inhibiting for both the processes like ethanol and methane production. Alkaline Pretreatment Alkaline pretreatment uses the bases, such as NaOH, Ca(OH)2 KOH, and NH3. H2O, to remove lignin and hemicellulose and/or cellulose, rendering lignocellulosic biomass more degradable through microbes and enzymes. The reactions taking place in alkaline pretreatment are solvation (cleavage of lignin-carbohydrate linkage) and saponification (Tarkow and Feist 1969). This process causes the biomass to a swollen state and makes the biomass more accessible for enzymes and bacteria. At strong alkali concentrations, dissolution such as “peeling” of end groups, alkaline hydrolysis, and decomposition of dissolved polysaccharides can take place in the lignocellulose biomass. This peeling is an advantage for the later conversion of lignocellulose biomass to ethanol and methane production. However, there is a risk on degradation and loss of carbon in the form of carbon dioxide also increased. An important aspect of alkali pretreatment is the solubilization, redistribution, and condensation of lignin and the modification of cellulose crystallinity state. This pretreatment has a positive effect on the digestibility of cellulose due to the removal of hemicellulose and part of lignin. However, there is a loss of hemicellulose to

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degradation products, and the solubilized lignin components often have an inhibitory effect. The loss of fermentable sugars and the production of inhibitory compounds make this pretreatment less effective for the ethanol production than methane production. This is because methanogens are often capable of adapting these inhibitory compounds with particular concentration.

Oxidative Pretreatment An oxidative pretreatment is one with the addition of oxidative compounds (hydrogen peroxide (H2O2), peracetic acid, dimethyldioxirane, and peroxymonosulfate) to the lignocellulose biomass that is suspended in water. In most of the research, the peroxides are transformed in situ into hydroxyl radicals (OH*). The most simple and frequently used peroxide is H2O2, and few studies were found on the application of other peroxides for the lignocellulose biomass for the bioenergy production. Due to this pretreatment, the lignocellulose biomass partially breaks down lignin and hemicellulose and the release of cellulose fraction for the enzymes and microbes. This pretreatment is a nonselective oxidation process; therefore loss of hemicellulose and cellulose could occur. In addition to that, inhibitors might be generated, due to oxidization of lignin to form the soluble aromatic compounds (Hendriks and Zeeman 2009). H2O2 is a strong oxidant that has been used for the lignocellulose biomass pretreatment for both ethanol and biogas production. Ammonia Fiber and CO2 Explosion Pretreatment Ammonia fiber explosion (AFEX) pretreatment is conducted with ammonia loading around 1:1 ratio (kg ammonia/kg dw biomass) at temperature ranging from ambient temperature with a duration of 10–60 days or to a temperature of up to 120 C and high pressure (17–20 bar) for 5–10 min (Alizadeh et al. 2005; Kim and Lee 2005). In AFEX pretreatment, the lignocellulose biomass is exposed to liquid ammonia at high temperature and pressure for a period of time, with a subsequent quick reduction of pressure (Alizadeh et al. 2005). AFEX pretreatment has shown good results on several lignocellulose biomass such as wood, switch grass, sugarcane bagasse, and corn stover. The advantages of AFEX, the ammonia could be recovered after pretreatment, unlike other pretreatments liquid waste fraction has not been formed. The AFEX pretreatment enables operating at high solid concentration. Recently, more research interests have started to determine the optimum conditions for the AFEX pretreatment of lignocellulose biomass. Carbon dioxide (CO2) explosion pretreatment is conducted with high-pressure CO2 at high temperature of up to 200  C with a duration of several minutes (Kim and Hong 2001). The idea behind using CO2 explosion is that it could have lower temperature than steam explosion and reduced expense compared to AFEX pretreatment. It was hypothesized that CO2 form carbonic acid when dissolved in water and this acid increases the hydrolysis rate. CO2 molecules are comparable in size to water and ammonia, and this could be able to penetrate small accessible pore size. Due to explosive release of the pressurized CO2, the distribution of cellulosic structure increases the accessible surface area of the substrate to

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hydrolysis (Zheng et al. 1998). Hence, this pretreatment can hydrolyze hemicellulose and cellulose that could be used for conversion of ethanol and methane production.

Ozonolysis Pretreatment The reaction taking place in ozonolysis pretreatment, when ozone is applied, this ozone molecules disintegrate into hydroxyl radicals (OH*) in water, resulting in a combination of oxidation by both ozone and the (OH*) radicals (Neely 1984). The pretreatment parameters are the water content in the reactor, particle size, and the ozone concentration in the gas stream. Among the parameters, water content is the most important factor affecting the solubilization of lignocellulose feedstocks (Vidal and Molinier 1988). The ozonolysis pretreatment is used to degrade lignin and hemicellulose in many lignocellulosic materials such as wheat straw, bagasse, green hay, peanut, pine cotton straw, and popular sawdust. Unlike other chemical pretreatments, it does not produce toxic inhibitory compounds. This pretreatment has an advantage that the reactions are carried out at room temperature and normal pressure. Furthermore, the ozone can be easily decomposed by using a catalytic bed or increasing the temperature means that could be designed to minimize environmental pollution. A drawback of this pretreatment is that a large amount of ozone is required that could make the process more expensive. Organosolv Pretreatment In recent year, organosolv pretreatment has attracted much attention for utilization of lignocellulose biomass. In this pretreatment, an organic or aqueous organic solvent mixture with inorganic acid catalysts (HCL or H2SO4) is used to break the integral lignin and hemicellulose bounds. The commonly used solvents are the methanol, ethanol, acetone, ethylene glycol, triethylene glycol, and tetrahydrofurfuryl alcohol (Thring et al. 1990; Chum et al. 1999). The organic acids such as oxylic, acetylsalicylic, and salicylic acids could also be used as catalysts in this pretreatment. The reaction involved is the simultaneous prehydrolysis by the organic solvents and the delignification of lignocellulosic biomass.

Biological Pretreatment Biological pretreatment for enhanced bioenergy production has mainly focused on fungal microbial consortium and enzymatic pretreatment. Comparing the physical and chemical pretreatment methods, biological pretreatments usually require lower energy input, and no chemicals are conducted under much milder environmental conditions. However, longer pretreatment time has limited the use of these processes in commercial application (Taherzadeh and Karimi 2008).

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Microbial-Consortium Pretreatment Microbial consortium pretreatment is conducted by the microbes that all screened from natural environments in rotten lignocellulosic biomass. Microbial pretreatment has high cellulose and hemicellulose degrading ability with minor affects in lignin content (Zhang et al. 2011). An approach for applying microbial metabolism to the challenges of biofuel production involves ensiling, which is commonly used for enhancing the digestibility of biomass for ruminants. This ensiling process exploits the capacity of naturally occurring bacteria to ferment the sugars within the lignocellulose biomass and produce the substrate that is more easily digested by ruminal microorganisms. While these bacterial consortia lack the ability to substantially degrade the lignin components (Zhong et al. 2011). Moreover ensiling is a relatively low input process that is anaerobic and therefore does not require mixing and aeration. For these reasons, ensiling could be incorporated into a biorefinery process at the earlier stages of energy production. Fungal Pretreatment In fungal pretreatment, brown-rot, white-rot, and soft-rot fungi are used to degrade the lignin and hemicellulose while utilizing little cellulose in lignocellulose materials. Brown rots mainly attack cellulose, while white- and soft-rot fungi attack both cellulose and lignin. On comparing these fungi, white-rot fungi are the most effective for biological pretreatment of lignocellulose materials (Sun and Cheng 2002). Fungal pretreatment is usually conducted in a sterile environment. Enzymatic Pretreatment To increase the bioenergy production from lignocellulose biomass and its biowaste, enzymes with hydrolytic activity were applied prior to or during the production step. The most commonly used enzymes are cellulase and hemicellulase. In most of the cases, the effect of enzymes on bioenergy production was minimal, but the cost of the enzymes was high; therefore application of enzymatic pretreatment was limited in commercial application. In general, most of the biological pretreatments are not efficient as chemical and physical pretreatments. This is because relatively high retention time is needed in biological pretreatment about 2–5 weeks (Muthangya et al. 2009).

Conclusion The focus on bioenergy production from the lignocellulose material has attracted a lot of research, development, and optimization in a number of fields related to conversion of lignocellulose into biorefinery products. Throughout the world, more pilot-scale facilities for pretreatments are being constructed, facilitating much greater evaluation of the technologies, their constraint, and opportunities. The ideal pretreatment therefore depends on the local conditions such as type, cost of raw material, and the need of products. The lignocellulose biomass pretreatment and the

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intrinsic structure of the biomass itself are primarily responsible for its subsequent steps. The conditions used in any chosen pretreatment method will affect various substrate characteristics which in turn govern the susceptibility of the substrate to subsequent hydrolysis step. Therefore, pretreatment is an extremely important step in the bioenergy production from the lignocellulosic biomasses. There is a critical need to understand the fundamentals of various processes that helps in making a suitable choice depending on feedstock structure. Future research is needed to focus on development and validation of large-scale systems. In particular handling of biomass with high solid concentration, large particle size, and varied substrate from forest and agricultural residues has to be tested and verified to ensure the long-term stability of a bioenergy production.

Cross-References ▶ Agrowaste Materials as Composites for Biomedical Engineering ▶ An Introduction to Sustainable Materials Management ▶ Biosynthesis and Assemblage of Extracellular Cellulose by Bacteria ▶ Development In-House: A Trap Method for Pretreatment of Fat, Oil, and Grease in Kitchen Wastewater ▶ Potential of Biogas Technology in Achieving the Sustainable Developmental Goals: A Review Through Case Study in Rural South Africa

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Daniel Jato-Espino, Luis A. Sañudo-Fontaneda, and Valerio C. Andrés-Valeri

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Types of Green Infrastructure (GI) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Materials Used in Green Infrastructure (GI) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Green Roofs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Swales . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Grassed Swales . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biofiltration Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Bioswales . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Wetlands . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Rain Gardens . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Permeable Pavement Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Porous Paving . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Interlocking Block Pavers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Grassed Surfaces . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Benefits of Green Infrastructure (GI) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusions and Recommendations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cross-References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

Green infrastructure (GI) is a concept that encompasses a set of treatment technologies aimed at providing an opportunity to reduce the impact of climate change and urbanization by delivering diverse environmental, social, and

D. Jato-Espino (*) · V. C. Andrés-Valeri GITECO Research Group, University of Cantabria, Santander, Cantabria, Spain e-mail: [email protected]; [email protected]; [email protected] L. A. Sañudo-Fontaneda GICONSIME Research Group, University of Oviedo, Mieres, Asturias, Spain e-mail: [email protected] © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_46

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economic benefits, including flood mitigation, water purification, climate adaptation, carbon sequestration and storage, enhancement of biodiversity, job creation, food production, and generation of spaces for recreation. As such, GI can be a cornerstone of spatial planning strategies intended to restore the health of ecosystems and improve the quality of life of people through a more livable and sustainable environment. GI is at the very heart of nature-based solutions (NBS) policies, which are designed to recover natural processes in cities by deploying properties inherent to natural ecosystems (ecosystem services). Thus, GI systems were initially designed and constructed as sustainable drainage systems (SuDS), also known as Stormwater Best Management Practices (BMPs), becoming rapidly in multipurpose systems which broadly responded to the NBS goals of sustainability and resilience. The construction of GI is usually based in using high-porosity media and vegetation for enhancing water infiltration, pollutants sequestration, and environmental benefits. Nonstructural GI techniques are normally built up using vegetated surfaces and open-graded aggregate materials, while structural GI systems rely on high-strength surface materials to achieve adequate mechanical performance and maintain infiltration and filtration capacities. Geosynthetics can also be incorporated into GI systems to boost soil stabilization, as well as water purification and retention.

Keywords

Climate change · Green infrastructure · Materials · Nature-based solutions · Sustainable drainage systems · Sustainability

Introduction The combined effects of urban growth and climate change (CC) are one of the biggest challenges which humans will have to face in the future as a society, in order to ensure the sustainability of well-being over time (Hoornweg et al. 2011). Increased urbanization involves rising amounts of carbon dioxide (CO2) released to the atmosphere, aggravating the condition of the main human-induced cause of CC. In fact, about 70% of the world population is forecasted to live in urban areas by 2050 (Tucci 2001), which highlights the need for taking actions against the unstoppable environmental harm provoked by human-related activities. Urbanization also contributes to exacerbate the two main consequences derived from CC, such as global warming (GW) and flooding. On the one hand, the transition from natural to built-up areas entails a reduction in the amount of vegetated cover, which results in a decrease of both evapotranspiration and solar reflectance that facilitates the occurrence of the urban heat island (UHI) effect and, by extension, the warming of cities (Frumkin 2002). Besides, the growth of global average temperature of 0.76  C experienced from 1850 to 2005 is also being heightened by CC through the observed increase in greenhouse gases (Tett et al. 2002). GW can not only result in a decrease of comfort caused by warmer

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air but also contribute to creating chemical reactions that favor air pollution and, therefore, endanger human health. On the other hand, the increase in the degree of development involves a decrease in the permeability of the urban skin, which facilitates the rapid generation of runoff that is conveyed throughout conventional drainage systems, disregarding soil moisture replenishment and groundwater recharge. Furthermore, the severity and frequency of extreme rainfall events are expected to be boosted by CC, which might exceed the capacity of traditional drainage networks (Huntington 2006). The final consequence of these considerations is the occurrence of floods, which are likely to provoke important economic, environmental, and social damage (Tingsanchali 2012). Hence, these phenomena can result in loss of life, harmful impacts on the environment and human well-being, or deterioration of property. The concept of green infrastructure (GI) emerged from the alternative designs taken to deal with the effects of urbanization on drainage. Thus, instead of resorting to conventional practices consisting of drainage systems that capture water to minimize runoff and then transfer it to a sewer network formed of a series of pipes and manholes connected to each other, these designs aimed at reproducing the natural water cycle (Woods-Ballard et al. 2015). They have been postulated as measures to attenuate peak flow and control water pollution through infiltration, transport, and retention means, enhancing the processes of evapotranspiration, water table recharge, and rainwater reuse (Roy et al. 2008). Fletcher et al. (2015) provided an extensive overview of the nomenclature proposed over the years to name these systems, which included GI as one of the most relevant and recurrent terms used for this purpose, along with the following: Stormwater Best Management Practices (BMP), integrated urban water management (IWUM), low-impact development (LID), low-impact urban design and development (LIUDD), stormwater control measures (SCMs), stormwater quality improvement devices (SQIDs), sustainable drainage systems (SuDS), sustainable urban drainage systems (SUDS), and watersensitive urban design (WSUD). Traditionally, the benefits provided by these solutions were arranged according to the called sustainable drainage triangle, which considered water quantity, water quality, and amenity as the three main fields where the implementation of GI might make a difference (Woods-Ballard et al. 2007). This triangle was revised to add a fourth pillar, namely, biodiversity, due to the contribution of GI to creating attractive habitats for plants and wildlife (Woods-Ballard et al. 2015). The potential of GI was further extended recently, emphasizing the role these systems can play regarding other aspects related to urbanization and CC, such as carbon sequestration and climate adaptation (Charlesworth 2010). This new dimension given to GI is extremely linked to the concept of nature-based solutions (NBS), which represent the increasing recognition and consensus that nature is a trigger to provide sustainable, cost-effective, viable, and multifunctional solutions to several challenges (Nesshöver et al. 2017). GI are categorized as structural or nonstructural measures depending on whether they are intended to provide certain mechanical properties, such as withstanding traffic loads, or not. In any case, these systems consist of physical constructions that

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can be combined to form treatment trains to improve their performance at different temporal and spatial scales (Jato Espino 2016). They are usually built using several layers made of different materials, e.g., open-graded aggregates, vegetation, or geotextiles, which are responsible for providing the aforementioned features, including water retention, pollution sequestration, and social recreation. Under these premises, the aim of this chapter was to deepen the concept of GI, providing a description of the main types into which these practices can be divided and highlighting the potential benefits their implementation can result in according to the principles of sustainable development. The approach taken to deal with these aspects focused on the role played by the construction materials used to build GI systems, whose proper management can make a difference in protecting the environment and ensuring human well-being.

Types of Green Infrastructure (GI) The European Commission for Environment defines GI as “a strategically planned network of natural and semi-natural areas with other environmental features designed and managed to deliver a wide range of ecosystem services such as water purification, air quality, space for recreation and climate mitigation and adaptation” (EU Commission 2016). This network contains green and blue spaces, with the former being related to land while the latter reflecting the spaces occupied by water. This intimate relation between land and water epitomizes the design of GI systems, becoming an inherent element to classify the different types of GI techniques. GI influences directly on land-use planning, not only by modifying the water management paradigm (Morison and Chesterfield 2012) but also by acting as the foundations of urban retrofitting activities, such as restoration and enhancement of urban landscape (Kati and Jari 2016). GI, when understood as a useful tool for planning, becomes fundamental to create and protect urban ecosystems and, then, turns out to be a better alternative to conventional drainage systems (Perales-Momparler et al. 2017). There are different ways of naming GI, which mainly depend upon the country where they are applied (Fletcher et al. 2015). In order to avoid confusion, an academic research especially tailored to this topic was carried out, following up from a similar scientific analysis on GI by SañudoFontaneda et al. (2017b). GI is very often related to stormwater management, and therefore the most used terminology is “green stormwater infrastructure” (GSI) (Sañudo-Fontaneda et al. 2017b). Thus, a search combining the keywords green infrastructure and stormwater was undertaken in the bibliographic database Scopus (Elsevier). The results amounted to 458 documents divided into 244 scientific articles, 156 conference papers, 26 reviews, 13 book chapters, and 4 books, among other types of documents (see Fig. 1). The high number of documents found in a range of 12 years up to nowadays clearly shows the great impact of GI on stormwater engineering, highlighting the importance of GI types within the context of stormwater management.

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Fig. 1 Publications distributed by document type

Fig. 2 Number of publications organized by subject area

Moreover, the subject area where all the documents found were published is key to define GSI typologies and their main applications. Fig. 2 provides evidence of the large number of subject areas studying GSI techniques, demonstrating the crosscutting and transversal vision of how multidisciplinary and highly impactful on society GSI is. Environmental science is, by a great distance, the most relevant subject area. As depicted by Pataki et al. (2011), GI portrays a vital role in coupling biogeochemical cycles in urban environments through the ecosystem services provided by GSI. Wendel et al. (2011) also remarked the importance of environmental science for assessing equitable access to urban spaces by means of GI related to water infrastructure.

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Engineering and social sciences also marked an important status as dominant subject areas. Engineering covers a rather large area of research mainly based on the hydrological impact of GSI (Emerson and Traver 2008), monitoring (Mitsova et al. 2011) and designing optimization (Lee et al. 2012). Social sciences have boarded GSI from a wide range of views, highlighting those related to the improvement of the livability in cities (Larsen 2015), water scarcity (Coutts et al. 2013) and flooding management (Zahmatkesh et al. 2015), citizen participation (Mayer et al. 2012), stormwater governance (Porse 2013), and willingness to implement GSI (Baptiste et al. 2015), among others. A high number of publications on GSI have been related to education and urban planning. Education has been included in many works related to GSI, mainly from growing greener cities (Vitiello 2008), higher education both at graduate and postgraduate levels (Bradford and Drake 2010; Scott et al. 2014; Ahn 2016), and the application of a hands-on approach to learn GSI while practicing (Hensley 2014; Price 2015). The main GSI techniques were identified in the same search within those publications related to specific practices. Table 1 shows that green, blue, and living roofs have been the most researched techniques for GSI, representing 21.1% from all the published research across the world. The high influence of the Green Street Programs implemented worldwide, especially focused on GSI techniques used to retrofit urban areas, has had their impact in the results from this study, being undertaken in 19.1% of the published work (see Table 1). Some of the most utilized GSI techniques within Green Streets are bioretention cells, rain gardens, and street planters (SañudoFontaneda et al. 2017b). Other widely used techniques in these programs are permeable pavements (17.1%); street trees, tree box planters, and urban forest (7.2%); bioswales and swales (6.8%); and rain barrels or rainwater harvesting devices (5.6%). Table 1 Main green stormwater infrastructure techniques (Source: Scopus 2000–2017)

Green stormwater infrastructure (GSI) techniques Green/blue/living roofs Bioretention cells, rain gardens, and street planters Permeable pavements Street trees, tree box filter, urban forest Bioswales and swales Constructed wetlands/urban lakes Rain barrels/rainwater harvesting devices Filter drains/French drains/infiltration trenches Dry and wet ponds/infiltration basins Soil bioretention/bioretention column Green walls/facades Stormwater biofilters Stream restoration Treatment train Floating wetland

% of publications 21.1 19.1 17.1 7.2 6.8 6.8 5.6 4.0 3.2 3.2 1.6 1.6 1.2 1.2 0.4

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Permeable pavements have probably become the most successful GSI technique used so far, becoming part of the “superficial skin” of many cities across the globe, allowing infiltration and recovering the preexisting natural water cycle before urbanization (Andrés-Valeri et al. 2016). These techniques have had special repercussion in the design and construction or parking lots, right-of-way drainage, cycle and pedestrian paths, and low-speed roads in urbanizations (Castro-Fresno et al. 2013). Street Trees and Tree Box Planters are part of a greener approach to Sustainable Drainage (Asawa et al. 2017), being highly utilized in Green Street Programs and other engineering and architectural Water Sensitive Urban Designs (WSUD) (Roseen et al. 2009). The main aim of their implementation is the achievement of the so-called urban forest, where local species of trees are planted back in urban environments to mimic the hydrological behavior of a natural watershed (McPherson et al. 2005). In addition to Green Street approaches, Rainwater Harvesting Systems have reached a high level of implementation in households and other private and public properties and buildings, completing a WSUD approach in stormwater management (Jones and Hunt 2010). These techniques, with special attention to rain barrels, work as storage devices for stormwater, reducing runoff flows and benefitting this natural resource by reutilizing water for flushing toilets and gardening (irrigation), among other demonstrated properties. Furthermore, they play an important role in urban agriculture, serving as storage tanks for crop irrigation in urban environments (Smith-Nonini 2016). They can also work as part of bigger treatment train systems consisting of more GSI, such as permeable pavements or green roofs. Constructed wetlands and urban lakes consume more surface than those systems described before. They are commonly designed alongside some urban forest approaches to generate natural landscapes in urban environments (Xu et al. 2013). Floating wetland devices are under development over the last years as a new field of research, becoming inherently linked to constructed wetlands as demonstrated by Winston et al. (2017) and Schwammberger et al. (2017). Dry and wet ponds and infiltration basins are also high area consumers and have been researched under many different perspectives. One of the most holistic approaches was taken by Abrahams et al. (2017). This approach, called wetland ecosystem treatment (WET), consisted of the link between permaculture design, stormwater management to improve flood resilience, biomass production, and biodiversity enhancement. Stream restoration is a field recently incorporated into GSI from a more naturalized approach. Stream restoration has been defined by BAE (2017) as a comprehensive process related to physical and biological components, which intends to recover the structure and function of urban natural streams before disturbance. Bonneau et al. (2017) also identified the impact of stream restoration on the recovery of the pre-development conditions of flow regime. Filter drains, also known as French drains, are a type of GIS with a low number of research studies undertaken worldwide but a high level of implementation, as demonstrated by Sañudo-Fontaneda et al. (2017a). Large stretches of highways in

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the UK are drainage systems by means of this technique. Filter drains are usually identified with infiltration trenches; however, there are some differences in the hydraulic objective pursued by each of them. While infiltration trenches aim to capture runoff and infiltrate it through them, filter drains are often designed to capture runoff and convey it toward an infiltration point. Filter drains can also act as infiltration trenches, but there are some structural differences between them depending upon their hydraulic aim (Coupe et al. 2016). Finally, the treatment train scheme represents the combination of several GSI techniques to achieve even a higher level of water quality purification and water quantity management (Lashford et al. 2014). Treatment trains are often designed and constructed in places where high-intensity rainfall regimes are common and conventional drainage systems cannot cope with stormwater, producing flooding events periodically (Sañudo Fontaneda 2014).

Materials Used in Green Infrastructure (GI) The construction of GI relies on natural vegetated surfaces, growing media, and open-graded aggregates. Sometimes, sustainable materials such as recycled or synthetic materials are also included in GI with both functional and aesthetic purposes (Garg et al. 2017). The design and construction of any GI practice should consider the physical and chemical characteristics of the study area, including topography, soil type, depth to water table, and surrounding land use (Charlesworth et al. 2016). Environmental conditions, construction, and operation constraints and the type and purpose of the GI practice to be built are other important factors to bear in mind for selecting suitable materials for the construction of these techniques.

Green Roofs Green roofs are one of the most extensively used and researched GI technique. The typical structure of a green roof consists of a vegetated surface coverage, a substrate or growing media, and a set of permeable layers. A separation geotextile is commonly included under the substrate to act as a filtering layer, along with a geocomposite or lightweight granular bed under the geotextile for draining. Green roofs also have a thick and strength geofabric installed at their bottom for protecting the rooftop waterproofing and insulating from the potential injuries caused by the plants rooting system (Andrés-Valeri et al. 2014a). Vegetal selection in green roofs is challenging due to the rooftop adverse conditions for plant survival and growth. Severe droughts, extreme temperatures, high light intensities, and high wind speeds increase the risk of desiccation and physical damage to vegetation and substrate (Dvorak and Volder 2010). The composition of roof vegetation depends on many factors, especially substrate depth conditions. Intensive green roofs enable installing a wide variety of vegetal species, with the only limitation of the geometrical, climatic, and structural conditions of the building. In contrast, plant

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selection for extensive green roofs is more complex due to substrate depth restrictions. Suitable plants for extensive green roofs have stress-tolerant characteristics, compact growth, evergreen foliage, twiggy growth, and drought tolerance (Oberndorfer et al. 2007). Succulent and herbaceous species are one of the most intensively investigated taxa on green roof sites (Dvorak and Volder 2010), since they are shallow root systems with the ability of using water efficiently and tolerating extreme conditions on rooftops. In general, Sedum species have been found to be very reliable for green roofs, because of their great drought tolerance, with some species surviving up to 4 months without precipitations (Durhman et al. 2006). Green-roof substrates are composed of a high amount of mineral materials, including organic matter in the range of 10–15% by weight. The mineral component may come from different sources and can have different weights depending on the rooftop load capacity. However, light expanded clay granules and crushed bricks are two of the most widely used materials with this purpose (Oberndorfer et al. 2007).

Swales Swales can be successfully built and operated in most soil types. Soils must be non-compacted to promote adequate root and biological organism developments, promote water retention for dry periods, and provide filtering and infiltration properties (Jurries 2003). They must also contain carbon and nutrients to enable initial vegetation and biota establishment. Soil requirements for swale structures include clay, sandy, and loam soils. Clay and silt hold moisture adequately, but delay the downward movement of water and can stop infiltration completely, which makes this material ideal for berm construction to prevent water flow. Moreover, sandy soils allow rapid water movement but do not retain it for long-term uses, proving to be unable to hold water and nutrients and causing their quick drying. Finally, loam soils consist of a mix of sand, silt, or clay and organic matter (Jurries 2003). These soils are loose and rich, allowing water infiltration and supporting healthy vegetation because of their water absorption and moisture storage.

Grassed Swales Plants are incorporated into swales to favor the deposition of suspended solids dragging pollutants and contribute to reducing flow rates (Rahman et al. 2011). Overall, perennial species are preferable to maintain vegetation cover during the winter season. Vegetal selection depends on the type of swale under construction. Grassed swales are very easy to construct and do not require large amounts of materials. Although native grasses are normally recommended for their role in enhancing biodiversity and wildlife (Charlesworth et al. 2016) using ad hoc designed grass types, there is room to improve swale performance in terms of both water quality and quantity (Woods-Ballard et al. 2015). The inclusion of geofabrics and open-graded aggregates under the growing media demonstrated to be favorable for

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reducing suspended solids in grassed swale structures (Andrés-Valeri et al. 2014b). Swales can be built with bulk materials such as leaves, rotten wood, or straw. A perforated pipe or other drainage system is often placed below the substrate to help in conveying runoff without producing overflows during severe storm events. Concerning the material used to form berms, most recommendations point to low permeability soils like clay and rocks. Sometimes, the natural soil under the swale can be waterproofed in order to avoid the infiltration of conveyed water (Andrés-Valeri et al. 2014b). Geocomposites, bituminous membranes, and high-density polyethylene (HDPE) coverages are often used with this purpose. Furthermore, rock and gravel are also used for building up a drainage layer under the substrate of grassed swales (Andrés-Valeri et al. 2014b) to improve stormwater quantity management and facilitate the infiltration of conveyed water.

Biofiltration Systems Bioswales, wetlands, and rain gardens can be classified as bioretention areas, also called biofilters. Biofiltration systems rely on the use of vegetation to delay runoff velocities, filter sediment, and pollutant accumulation, guaranteeing water quality (Ogle and Hoag 2000), as illustrated in Fig. 3. Their structure and performance are very similar to each other, mainly differing in the floor dimensions of the plants and their capacity for water treatment.

Bioswales A bioretention swale or a bioswale is a bioretention system located within the base of a swale (Fig. 3). Bioswales are vegetated channels designed to treat pollutants

Fig. 3 Bioretention area

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derived from stormwater runoff. Stormwater conveyed by sewer networks enters the channel at the inlet and receives treatment before it leaves through the outlet. In general, bioswales are at least 30 m long and 0.6 m wide, have a longitudinal slope in the range from 0.5% to 6%, and are located in series with detention ponds, which store runoff and reduce peak discharge (Mazer et al. 2001). Since pollutant reduction is the primary objective of bioswales, a great variety of vegetal species, including macrophytes, can be used for building the vegetal coverage. They should be tolerant to flooding conditions and high organic and inorganic loadings, as well as adaptable to local weather conditions and diseases (Leroy et al. 2017). The selection and planting of vegetation must be accomplished in accordance with both the characteristics of the pollutants to be removed and the flow and velocity design requirements for the bioswale (Jurries 2003). In addition, the vegetation selected for a biofiltration swale must have characteristics providing either vegetal delay of water or Manning’s roughness coefficient between 0.20 and 0.24 (Ogle and Hoag 2000). Choosing between turf or woody plants depends on the desired capacity and residence time of stormwater and pollutants in the bioswale. Aboveground plant parts are capable of inducing sedimentation of particulates and pollutants, while plant roots stabilize sediment deposits, which prevent sediment re-suspension (Mazer et al. 2001). Bioswales are generally composed of three basic vegetation zones: high, middle, and low. Vegetation in the low zone should tolerate standing water and fluctuating water levels. Plants in the middle zone deal with slightly drier conditions and more infrequent fluctuating water levels, being often selected for erosion control purposes. Finally, the high zone can be planted with species adapted for drier conditions. Woody plant materials should only be located on side slopes, while trees must be planted along the edge of the bioswale, in order to provide shade to minimize temperature increases in water during dry months. A lower canopy of shrubs and grasses can also be planted underneath the trees (Jurries 2003).

Wetlands Wetlands are frequently constructed by excavating, backfilling, grading, diking, and installing water control structures to establish hydraulic flow patterns. If the construction site is founded on highly permeable soils, an impervious and compacted clay liner is usually installed, placing the original soil over the liner. Wetland vegetation is then planted or allowed to grow naturally (Kadlec and Knight 2004). A common concern with wetlands is the potential loss of water from infiltration and contamination of groundwater below the construction site (EPA 1999). Although there are some applications where infiltration is desirable, the majority of wetlands require the installation of a barrier to prevent groundwater contamination. Under ideal conditions, the wetland site consists of natural soils with low permeability (between 104 and 106 cm/s) that restrict infiltration (EPA 1999; Winogradoff 2002). If necessary, a barrier can be placed over the natural soil for preventing excessive infiltration. This barrier is normally built up using clay fills,

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bentonite soil layers, chemical treatment of existing soils, asphalt, and synthetic membrane liners such as polyvinyl chloride (PVC) or HDPE. Wetland substrates support vegetation and provide sites for both biochemical and chemical transformations and storage of removed pollutants. Commonly used substrates include soil, sand, gravel, and organic materials (Davis 1995). The soil substrate for wetland vegetation should be agronomic in nature, well loosened, and at least 150 mm deep. Depending on the material of the liner, deeper soil substrates might be required to protect it. Although soils like loam and silt are suitable for plant growth, they can enable large vegetation mats to float when large water level fluctuations occur in the wetland. Denser soil substrates such as a sandy loam or loam gravel mixes can be used to prevent this potential problem. Constructed wetlands receiving water with high nutrient content, such as domestic and agricultural wastewaters, can be built with sand or gravel (Davis 1995). Wetland vegetation is basically composed of six types of vegetal species: freefloating aquatic species, rooted floating aquatic species, submerged aquatic species, emergent aquatic species, shrubs, and trees. Plant selection should be carried out using locally available species, avoiding invasive or aggressive plants, and searching for tolerance to high pollution loads and continuous flooding. Hence, perennial species with a slow growth rate and wildlife benefits are normally preferred. The types of plants that are most often used in constructed wetlands are persistent emergent plants, such as bulrushes (Scirpus), spikerush (Eleocharis), other sedges (Cyperus), rushes (Juncus), common reed (Phragmites), and cattails (Typha) (Davis 1995). However, plant communities in wetlands undergo significant changes with respect to their initial planting conditions (EPA 1999). Only a few constructed wetlands maintain the original species composition and density distributions envisioned by their designers.

Rain Gardens Rain gardens are depressed areas in the landscape designed to manage stormwater (Dietz 2007). Hence, the aim of rain gardens is to capture and filter stormwater runoff using a permeable soil substrate and plants tolerant to both drought and inundation (Richards et al. 2015). Rain garden designs used for stormwater retention often rely on planted depressions placed downstream from drainage areas. Since adequate infiltration and percolation rates are essential properties for rain gardens, careful investigations at potential construction sites must be done. Percolation rate should be higher than 2.5 cm/h, and organic matter content should be above 5% (Richards et al. 2017). If the existing soils fail to meet these criteria, they can be amended by adding sand and compost until obtaining a suitable performance in relation to those characteristics. High infiltration soils, such as sandy soils, are recommended for improving rainwater infiltration (Yu et al. 2013). High infiltration materials like gravel, sand, or plastic cells usually replace underlying soils, in order to produce draining layers in which perforated pipes are often included for limiting possible overflows. A geofabric is usually placed over the drainage layer for avoiding the scouring of fine particles into the drainage outlet.

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Perennial native plant species are recommended for rain gardens (Richards et al. 2015), mainly because their capability to develop deep root systems, which enable building and maintaining high organic matter content and soil porosity. Traditionally, low growing grasses compose the border of the rain garden, while its bottom consists of a variety of plant species that bloom throughout the growing season. The selection of these species should consider lower growing native plants whose eventual height is limited to 1 m. The surface of rain gardens is usually covered by mulch material, which helps in the erosion control of the soil (Dietz 2007).

Permeable Pavement Systems In general, neither specific design nor aggregate gradation is used in permeable pavement systems. The cross section of a permeable pavement cross section consists of a permeable surface on top, an open-graded bedding layer, a subbase reservoir layer, and the compacted subgrade. Non-woven geotextiles can also be used to separate the bedding layer and the reservoir bed and/or to be enable water storage over the subgrade soil. Perforated pipes, geocomposites, or waterproofing membranes can also be incorporated into the cross section of permeable pavements for specific applications (Castro-Fresno et al. 2013). Geotextiles can be used as separation and strengthening layers under roads and car parks, as well as for filtering purposes in other GI practices. Geotextiles can enhance organic matter removal by trapping the pollutants deposited on their surface, enabling microbial biodegradation. Moreover, although geotextiles may support drainage, their use might lead to potential problems with frost in cold climatic regions, such as Scotland and Canada (Scholz 2013). The role of geotextiles is usually divided into five main categories: separation, filtration, drainage, protection, and reinforcement. The open-graded bedding layer is normally composed of medium- to small-sized aggregate materials, usually limestone, in order to provide slightly higher air void contents than those of the upper layer and prevent stability damages. The subbase reservoir layer can consist of clean crushed stone with few fine particles, in order to ensure a minimum void ratio of 40%. Commonly used aggregate materials include limestone, gravel, basalt, and sandstone (Mullaney and Lucke 2014). Some experiences concerning the use of recycled or alternative aggregates such as basic oxygen furnace slag (BOF-slag) in permeable pavements have been reported (Sañudo-Fontaneda et al. 2014). Other options for building the subbase layer of these systems consider plastic cells (Andrés-Valeri et al. 2014a), which provide the bearing capacity required for withstanding light traffic volumes and ensuring a high storage capacity in the reservoir layer. There are different types of permeable surfaces that are frequently used in permeable pavement systems, including porous materials, interlocking block pavers, and vegetated surfaces reinforced with concrete or plastic grids (Andrés-Valeri et al. 2014a; Rodriguez-Hernandez et al. 2016), as depicted in Fig. 4. Other possible surfaces include soft paving materials, such as wood mulch and crushed shells, which are typically used for foot traffic.

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Fig. 4 Permeable surfaces. (a) Porous asphalt. (b) Porous concrete. (c) Interlocking block pavers. (d) Grassed surface

Porous Paving Porous materials, which are a particular type of surface used in permeable pavement systems, are made of a granular skeleton coated with a binder. Depending on the binder used, two main types of porous surface might be defined: porous concrete (PC), if cementitious material is used as a binder, and porous asphalt (PA), in case the binder is made of bitumen. PC and PA mixtures are designed to ensure high air void contents, normally greater than 18–20%, while providing the required bearing capacity and durability to resist the traffic loads applied during their service life (Tennis et al. 2004; Alvarez et al. 2011). To guarantee adequate infiltration capacities, open-graded aggregate materials with low or no fine content are used for building up the granular skeleton (Andrés-Valeri et al. 2016). Aggregates should ensure angularity, roughness, chemical stability, and an adequate mechanical performance. The criteria for selecting aggregates are normally based on local availability; however, the most widely used aggregates in PA include limestone, granite, quartzite, basaltic, and porphyric materials. Although some experiences with promising results have been also reported in relation to the use of recycled and synthetic aggregates, further research is needed in this line (Shen et al. 2008; Frigio et al. 2013).

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Binders used for PA mixtures include materials like styrene-butadiene, styrenebutadiene rubber, crumb rubber, and high-viscosity bitumen (Chen et al. 2013a; Ibrahim et al. 2014; Rodriguez-Hernandez et al. 2015). Binder content in PA mixtures is variable, but normally ranges between 5% and 12% by weight, depending on the binder type, mix design, additives, expected traffic level, and climatic conditions (Alvarez et al. 2011). The low content of fine particles and the presence of open-graded aggregates can result in draindown problems for high bitumen ratios. To prevent this situation, stabilizing fibers are often incorporated into PA mixtures, including cellulose or mineral materials (Alvarez et al. 2011). Other materials like Portland cement, hydrated lime, limestone filler, crumb rubber, or polymeric fibers are also used in PA materials for improving their mechanical performance, durability, and service life (Alvarez et al. 2006). Cementitious binders used in PC materials can also vary depending upon the additives used in the cement paste. Portland cement is the most extensively used cementitious material for PC mixtures (Tennis et al. 2004), normally in addition to water reducers, silica fume, viscosity modifiers, and air entraining admixtures (Yang and Jiang 2003; Andrés-Valeri et al. 2016). Cement content normally exceeds 300 kg/m3 for trafficked applications. Water to cement ratios are generally beyond 0.4 in conventional PC materials, but can be lowered down to 0.2 when additives or polymeric solutions are incorporated into concrete mixes. These additives are usually dosed according to manufacturer’s specifications. Interesting developments have been achieved by adding polymeric solutions to fresh cement paste, resulting in the so-called polymer-modified porous concrete (PMPC). Various types of polymeric additions have been successfully tested in PMPC mixtures, such as styrenebutadiene polymers, vinyl acetate ethylene (VAE) polymers, acrylic emulsions, or polyvinyl alcohol formaldehyde solutions (Yang and Jiang 2003; Chen et al. 2013b; Shen et al. 2013). Among them, VAE polymers have shown the best performance so far, probably due to their positive chemical interactions with Portland cements.

Interlocking Block Pavers Permeable interlocking pavers are common pavers, but designed for allowing water infiltration. Compared with other porous paving materials, properly constructed block pavements are solid and stable, so that they can bear almost any traffic, including moving vehicles (Ferguson 2005). They can be divided into two categories: porous blocks or impervious blocks with open joints. Both systems allow water to infiltrate across the pavement surface. The air void content in porous materials normally exceed 20%, while the joints in impermeable blocks cover between 8% and 20% of the total pavement surface area and are usually filled with 2–5 mm aggregate material (Mullaney and Lucke 2014). This type of surface relies on the geometry of blocks to provide interlocking and structural strength. Blocks can be used to create a permeable surface with the aesthetic appeal of brick, stone, or other interlocking paving materials. Although they are often made with high-strength concrete, there are also some investigations concerning the use of ceramic pavers.

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Porous paving blocks are made of porous concrete reinforced with polymers and exhibit better fatigue behavior than those without plastics. Still, these improvements have been found to decrease for low values of stress, to the extent to appear to be negligible in the case of traffic loads associated with main and highway roads (Scholz and Grabowiecki 2007). These concrete products can act as pollution sinks, because of their particle retention capacity during filtration. Hence, the high porosity of this type of concrete leads to improved infiltration and air exchange rates. Those pollutants that are filtered out can sometimes be removed by cleaning the pavement.

Grassed Surfaces The design and role of grassed surfaces reinforced with plastic or concrete grids generally have significant differences in relation to the types of surfaces previously described, especially in terms of the impervious area of the pavement. Grass reinforced with concrete grids is much larger than individual interlocking block pavement surfaces, having more open void space to promote infiltration. The percentage of open voids ranges from 20% to 50% for concrete grids and between 90% and 98% for plastic grids (Mullaney and Lucke 2014). The materials used for filling the voids in plastic and concrete grids are normally composed of a suitable combination of mineral aggregates and organic material, in order to ensure adequate infiltration capacities, sustain vegetation, and filter runoff. Locally available grasses proving to be resistant to droughts conditions and high pollutant loads should be selected as vegetal coverage. The grass planted in the gaps of the paving often experiences severe heat stress in dry periods, which can shorten their life cycle. In the end, this can erode the soil in the paving apertures during heavy rainfall. Therefore, this type of system has been reported to be more suitable for cooler climates (Lucke and Beecham 2011).

Benefits of Green Infrastructure (GI) The broadening of the concept of GI, whose consideration rapidly evolved from solutions limited to improve water management to multifunctional systems capable of responding to a number of additional economic, environmental, and social challenges, has highlighted the potential positive contribution these techniques might have to sustainability. An enlightening means to provide evidence of the role that GI can play in addressing different issues related to the sustainability of people, planet, and prosperity can be obtained based on the Sustainable Development Goals (SDGs). The SDGs are the 17 objectives included in the 2030 Agenda for Sustainable Development approved in the United Nations Conference on Sustainable Development, held in Rio de Janeiro in 2012 (UN-DESA 2012). A summary of the SDGs and the targets into which they are divided is given in Table 2.

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Table 2 Sustainable Development Goals (SDGs) and targets into which they are divided

SDG 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17

Concept No poverty Zero hunger Good health and well-being Quality education Gender equality Clean water and sanitation Affordable and clean energy Decent work and economic growth Industry, innovation and infrastructure Reduced inequalities Sustainable cities and communities Responsible consumption and production Climate action Life below water Life on land Peace, justice, and strong institutions Partnerships for the goals

1541 Targets 5 8 13 10 9 8 5 12 8 10 10 11 5 10 12 12 19

The description provided about GI techniques, the different types into which they can be categorized, and the materials that can form them enabled establishing their relationship to some of the SDGs listed in Table 2. Hence, Table 3 highlights the specific targets considered by the United Nations in the area of sustainable development that can be addressed through the implementation of GI. The materials used to build GI generally consist of porous media and vegetated surfaces, which provide a variety of environmental, energetic, and social benefits that can contribute to achieving up to 28 targets framed in 12 SDGs, as demonstrated in Table 3. The main original purposes of all GI types were to improve water quantity and quality management, thanks to the increased porosity of their layers and the pollutants retention and treatment capacity of some materials (Bayon et al. 2015), such as geotextiles or biochar. These properties strengthen the resilience of urban areas to flooding phenomena and also reduce the concentration of pollutants in runoff (Ellis 2013), enabling their future reuse with non-potable purposes and protecting receiving water bodies. Aspects related to the management of water-related disasters and the purification of water are considered by the United Nations through SDGs 1, 3, 6, 11, 13, 14, and 15. The higher solar reflection provided by open-graded and vegetated covers in comparison with built-up surfaces enables reducing near-surface air temperature and, therefore, attenuating the effects of global warming. Furthermore, this characteristic also contributes to mitigate the urban heat island effect experienced in urban areas due to an excessive presence of dark surfaces (Solecki et al. 2005). Consequently, solar reflection plays an important role in increasing resilience to climaterelated hazards expressed in SDGs 1 and 13.

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Table 3 Sustainable Development Goals (SDGs) and targets to which green infrastructure (GI) might contribute SDG 1 2

3

6

7

8

9

11

12

Target By 2030, build the resilience of the poor and those in vulnerable situations and reduce their exposure and vulnerability to climate-related extreme events [. . .] By 2030, double the agricultural productivity and incomes of small-scale food producers, in particular women, indigenous peoples, family farmers, pastoralists, and fishers [. . .] By 2030, [. . .] implement resilient agricultural practices that increase productivity and production, that help maintain ecosystems, that strengthen capacity for adaptation to climate change [. . .] By 2020, halve the number of global deaths and injuries from road traffic accidents By 2030, substantially reduce the number of deaths and illnesses from hazardous chemicals and air, water, and soil pollution [. . .] By 2030, improve water quality by reducing pollution, eliminating dumping, and minimizing release of hazardous chemicals and materials, halving the proportion of untreated wastewater [. . .] By 2030, substantially increase water-use efficiency across all sectors and ensure sustainable withdrawals and supply of freshwater to address water scarcity [. . .] By 2030, ensure universal access to affordable, reliable, and modern energy services By 2030, increase substantially the share of renewable energy in the global energy mix By 2030, double the global rate of improvement in energy efficiency By 2030, enhance international cooperation to facilitate access to clean energy research and technology, including renewable energy, energy efficiency [. . .] Improve progressively, through 2030, global resource efficiency in consumption and production and endeavor to decouple economic growth from environmental degradation [. . .] Develop quality, reliable, sustainable, and resilient infrastructure [. . .] to support economic development and human well-being, with a focus on affordable and equitable access for all By 2030, ensure access for all to adequate, safe, and affordable housing [. . .] By 2030, provide access to safe, affordable, accessible, and sustainable transport systems for all, improving road safety [. . .] Strengthen efforts to protect and safeguard the world’s cultural and natural heritage By 2030, significantly reduce the number of deaths and [. . .] people affected and substantially decrease the direct economic losses [. . .] caused by disasters, including water-related disasters [. . .] By 2030, reduce the adverse per capita environmental impact of cities, including by paying special attention to air quality and municipal and other waste management By 2030, provide universal access to safe, inclusive, and accessible, green and public spaces, in particular for women and children, older persons, and persons with disabilities By 2020, substantially increase the number of cities [. . .] adopting [. . .] integrated policies and plans toward inclusion, resource efficiency, [. . .] adaptation to climate change, resilience to disasters [. . .] By 2030, achieve the sustainable management and efficient use of natural resources By 2020, achieve the environmentally sound management of chemicals and all wastes throughout their life cycle [. . .] and significantly reduce their release to air, water, and soil [. . .] By 2030, substantially reduce waste generation through prevention, reduction, recycling, and reuse (continued)

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Table 3 (continued) SDG 13 14 15

Target Strengthen resilience and adaptive capacity to climate-related hazards and natural disasters in all countries By 2025, prevent and significantly reduce marine pollution of all kinds, in particular from land-based activities, including marine debris and nutrient pollution By 2030, combat desertification, restore degraded land and soil, including land affected by desertification, drought, and floods, and strive to achieve a land degradation-neutral world By 2020, integrate ecosystem and biodiversity values into national and local planning, development processes, poverty reduction strategies and accounts

In line with the enhanced draining capacity and solar reflectance of GI, permeable pavements can also have impacts on the safety perception of users, since their implementation results in both runoff being removed from the surface and increased visibility for drivers. Therefore, the splash and spray effects produced by runoff accumulation are decreased, while the reflection of road lights and vehicle headlights are increased (Rungruangvirojn and Kanitpong 2010). These issues are represented in SDGs 3, 9, and 11, which concern road safety and traffic accidents reduction. Those GI practices involving arable vegetated surfaces, such as rain gardens or green roofs, can help in increasing the amount of cultivated land in urban areas and improve the socioeconomic progress and well-being of their inhabitants through enhanced agricultural productivity (Wilkinson and Torpy 2016). This factor impacted on SDG 2, since the implementation of green arable infrastructures may provide opportunities to boost the rates of food productivity. Mitigating carbon dioxide is a great challenge for future societies, since this gas is the main human-induced cause of climate change and also increases ozone and particle concentrations (Charlesworth 2010). Vegetation and trees can decrease ground-level ozone or smog through the attenuation of air temperatures, as well as absorb particulate matter. Besides, particular GI technologies like green roofs offer an enhanced behavior in terms of acoustic insulation, providing extra protection against noise in relation to conventional roofs. As a result, GI might have beneficial health effects and, therefore, help meeting SDGs 3, 11, and 12. The presence of GI practices in urban areas also provides an opportunity for improving energy efficiency, either through its generation from biomass derived from plant-related surfaces or by reducing its demand in buildings and other urban spaces using green layers that act as cooling and heating regulators (Saadatian et al. 2013). The objectives established by the United Nations regarding energetic efficiency and facilitated access to energy are included in SDG 7. The planning and management of landscape patterns must be based on the presence of natural surfaces, which can act as a trigger for ecological sustainability (Benedict and McMahon 2012). Furthermore, green urban areas also provide pleasant and comfortable environments for social recreation, while helping to control the loss of biodiversity produced by the increased degree of development of cities

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Table 4 Summary of some of the benefits provided by green infrastructure (GI) reported by O’Neil (2014) Dimension Economic

Environmental

Social

Contribution of green infrastructure to sustainable development 50% of capital costs in relation to traditional drainage solutions 4.5  C of temperature regulation using green roofs and walls £2.3M of annual contribution of trees through air purification +40% commercial trading within town centers including green infrastructure 9.1% of suspended particles removed by urban vegetation 3.16 kg of carbon stored in 1m2 of green space 3 dB reduction due to the use of grass instead of built-up surfaces Up to 8  C decrease in urban temperature from vegetated cover 17–20% reduction in runoff using green roofs in residential areas +50% species transfer when habitats are connected by green areas People happier when living in urban areas covered by large amounts of green space 83% of respondents believe green areas provide focal points for communities 24% of people more likely to be physically active if have access to green space 40% less chances to be obese if living in a highly green urban area

(Andersson et al. 2014). For these reasons, GI can contribute to fulfilling SDGs 11 and 15 by guarantying the existence of natural inclusive areas and integrating ecosystem and biodiversity values into planning strategies, respectively. The construction materials commonly used to build GI, especially aggregates, polymers, substrate, concrete, or asphalt, can be partially or totally recycled without compromising their properties and behavior. In addition, these systems can also contribute indirectly to resource efficiency by protecting from sunlight, atmosphere, water, and land (Demuzere et al. 2014). These facets of GI are linked to SDGs 8 and 12, which stand for the sustainable management of natural resources, as well as the recycling and reuse of materials. As a more tangible proof of the potential benefits of GI to sustainable development, Table 4 compiles some facts and figures reported in a series of investigations and reports related to the impact of green space in urban areas, arranged according to their economic, environmental, and social contributions.

Conclusions and Recommendations This chapter explored the concept of green infrastructure (GI) and provided evidence of their potential efficiency and the main environmental and social benefits they may induce, with a special focus on the materials commonly used to build them. Based on the introduction of some relevant terms related to GI, a review about the release of scientific documents and technical reports related to the use and application of these techniques was presented to give way to a description of their main types. This task

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laid the foundations needed to overview the materials forming different GI practices, which in turn enabled discussing the positive impacts that these technologies might have on sustainable development. The research carried out so far in relation to GI demonstrated that the choice of materials to build these systems can have a positive influence on a variety of aspects, including stormwater management, temperature regulation, air purification, and noise reduction, among others. Overall, suitable plants for vegetated systems should have a number of characteristics related to stress, pollution, drought, and flooding tolerance, organic matter content, sediment stabilization, and localness, in order to avoid invading surrounding ecosystems. Soils to be used as substrate in these vegetated layers should be loosened to facilitate the growth of microorganisms and enhance water retention. Regarding the materials used to build open-graded layers in GI practices, the valorization of plastic cells and alternative aggregates derived from industrial by-products provides an opportunity to improve both the bearing and water storage capacities of these systems. In summary, GI are comprehensive systems capable of mitigating the harmful effects of two of the greatest threats that humans will have to face over the next years: climate change and urbanization. Therefore, GI technologies are multifunctional measures that can make a difference in the achievement of sustainable development. As such, their implementation should be considered in urban planning designs and strategies aimed at safeguarding the environment and protecting human health and well-being.

Cross-References ▶ Constructed Wetland: A Green Approach to Handle Wastewater Acknowledgments The investigation presented in this chapter was possible thanks to the research project SUPRIS-SUReS (Ref. BIA2015-65240-C2-1-R MINECO/FEDER, UE), financed by the Spanish Ministry of Economy and Competitiveness with funds from the State General Budget (PGE) and the European Regional Development Fund (ERDF).

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Ehsan Banayan Esfahani, Fatemeh Asadi Zeidabadi, Alireza Bazargan, and Gordon McKay

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . The Importance of Nitrogen and Phosphorus Removal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Types of Biological Nutrient Removal Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Advantages of Biological Nutrient Removal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . The History of the Modified Bardenpho Process . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biological Phosphorus Removal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Source of Phosphorus in Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Importance of Phosphorus Removal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chemical Removal of Phosphorus: Advantages and Disadvantages . . . . . . . . . . . . . . . . . . . . . . Biological Phosphorus Removal Process . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biological Nitrogen Removal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Source of Nitrogen in Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Importance of Nitrogen Removal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chemical Removal of Nitrogen: Advantages and Disadvantages . . . . . . . . . . . . . . . . . . . . . . . . . Biological Nitrogen Removal Process . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Types of Biological Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Suspended Growth Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Attached Growth Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Performance of the Modified Bardenpho Process . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Reactors in the Modified Bardenpho Process . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Comparison of the Modified Bardenpho Process with Other BNR Systems . . . . . . . . . . . . . . Modified Bardenpho Process Design . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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E. Banayan Esfahani · F. Asadi Zeidabadi Department of Chemical and Petroleum Engineering, Sharif University of Technology, Tehran, Iran A. Bazargan (*) Department of Civil Engineering, K. N. Toosi University of Technology, Tehran, Iran e-mail: [email protected] G. McKay (*) Division of Sustainable Development, College of Science and Engineering, Hamad Bin Khalifa University, Qatar Foundation, Doha, Qatar e-mail: [email protected] © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_87

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Nutrient Removal Establishment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . COD and Nutrient Ratio . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Hydraulic Retention Time (HRT) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sludge Retention Time (SRT) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Recycling Ratio (R) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Temperature, pH, and Bubble Size . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Modified Bardenpho Process Applications . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . COD, BOD, TSS, Nutrient, and Heavy Metal Removal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Viruses Removal in Modified Bardenpho Process . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Estimation of Greenhous Gas Emissions of Modified Bardenpho Process . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

Excess presence of nitrogen and phosphorous, two fundamental prerequisites for plant photosynthesis, in water can cause noteworthy problems such as eutrophication and health issues for humans. Hence, efforts have been made to find solutions decreasing their concentrations. In addition to methods such as ion exchange, air stripping and breakpoint chlorination for nitrogen removal, and coagulantion-flocculation for phosphorous removal, biological methods for nutrient removal have also been used. In this chapter the following three categories will be discussed: nitrogen removal processes, phosphorous removal processes, and combined nitrogen/phosphorous removal processes. The basic concept of biological nitrogen removal processes relies on nitrification and denitrification which are two major steps in the nitrogen cycle. A biological nitrogen removal process should normally consist of at least one aerobic and one anoxic reactor. The Modified Ludzack and Ettinger (MLE) process and the Bardenpho process are two biological nitrogen removal processes. Biological phosphorus removal (BPR) depends on the incorporation of phosphorus into cell biomass and subsequent phosphorus removal by sludge wasting. BPR processes generally consist of an anaerobic reactor followed by anoxic or aerobic reactors such as the A/O (anaerobic/aerobic) and PhoStrip processes. Biological combined nitrogen/phosphorous removal processes contain all the three main biological conditions (aerobic, anaerobic and anoxic conditions). The A2/O (anaerobic/anoxic/aerobic) and the modified Bardenpho processes remove both nitrogen and phosphorous simultaneously. The modified Bardenpho process is a biological process which provides special conditions for both nitrogen and phosphorous removal. This system consists of five distinct reactors which are respectively: anaerobic reactor, first anoxic reactor, first aerobic reactor, second anoxic reactor, and second aerobic reactor. Each reactor provides appropriate conditions to play its special role in the removal of wastewater impurities. Also, each reactor has specific conditions such as pH and temperature. The modified Bardenpho process’s performance in the removal of nitrogen and phosphorous is respectively excellent and good. Since the modified Bardenpho process has five distinct bioreactors and every reactor has specific functions and required conditions, there are some critical

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parameters in designing the process. These parameters are nutrient removal establishment, chemical oxygen demand and nutrient ratio, hydraulic retention time, sludge retention time, recycling ratio, temperature, pH and bubble size. These parameters play specific roles in the modified Bardenpho process’s efficiency in removing wastewater impurities and should be optimized. Due to the utilization of five biological stages in the modified Bardenpho process, not only does the modified Bardenpho process remove nitrogen and phosphorus efficiently as its main function, but it also has other benefits such as reducing chemical oxygen demand, biological oxygen demand, total suspended solids, heavy metals and viruses. Keywords

Nitrogen · Phosphorous · Biological nutrient removal · Modified bardenpho process · Optimization

Introduction The Importance of Nitrogen and Phosphorus Removal Nitrogen, as the 7th element of the periodic table, and phosphorus, as the 15th element, are considered as nutrients since they are two fundamental prerequisites for microorganism growth and plant photosynthesis. However, the excess presence of these elements in water can cause serious problems; hence, nitrogen and phosphorus concentration in water are measured and should be limited. In the initial section of this chapter, some of the most noteworthy problems due to the presence of excess nutrients in water, namely, eutrophication, health issues for humans, and the increase of the chemical oxygen demand of the water, will be discussed.

Eutrophication When the amount of nutrients in the water increases, this is will result by an overgrowth of plants and algae. Such events are sometimes referred to as algal blooms. Algal blooms block sunlight from reaching the bottom of the lake or river; and hence, plants which live in the depths will die due to a lack of energy via sunlight. The bacterial decomposition of these biomasses will subsequently consume the oxygen in the water. A decline in the oxygen concentration of the water will cause hypoxia which impairs the survival of other marine organisms and decreases water quality (Chislock et al. 2013). Eutrophication might naturally occur in lakes and rivers through centuries, but human activities dramatically accelerate the rate of this phenomenon. Human activity, chiefly pollutants from agriculture, industry, and sewage disposal containing high amounts of nutrients are the main culprits (Chislock et al. 2013). Eutrophication can cause serious problems such as damaging drinking water resources, destroying recreation and aesthetic advantages of lakes and rivers, and causing taste and odor

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issues (Dodds et al. 2008). Such problems, in addition to harming the environment, result in economic losses. Hence, the release of nutrients into bodies of water should be limited by policies and relevant standards. The total costs, including recreational water usage, waterfront real state, recovery of endangered species, and drinking water, have been estimated at approximately $2.2 billion annually as a result of eutrophication in US freshwaters (Dodds et al. 2008). The combination of phosphate, nitrogen, and carbon resources causes eutrophication, but each has a different impact. A series of experiments which started in 1973 in a small lake (Lake 226 of the Experimental Lakes Area of Northwestern Ontario) have yielded interesting results. The body of water was separated into two sections using a sea curtain (Schindler 1974). Both of these two areas contained the same amounts of nitrogen and carbon, but phosphorus was added just to the northeast basin of the lake. The investigations illustrated that the main limiting factor for eutrophication is phosphate since the basin with phosphorus was covered with microorganisms within 2 months, while the south basin did not show signs of eutrophication (Schindler 1974). In another experiment in Lake 227, it was demonstrated that a shortage of carbon did not prevent eutrophication (Schindler 1974). Due to the catastrophic consequences of eutrophication, various actions have been taken to combat this phenomenon all around the world. In China, there are over 110,000 lakes, and through the 1990s, most of these lakes faced drastic eutrophication problems (Liu and Qiu 2007). The main causes of this mishap are rapid population increase and the disposal of industrial and domestic wastewater into lakes which cause a dramatic degradation of water quality (Liu and Qiu 2007). Prevention strategies for confronting eutrophication are divided into two categories: external and internal nutrient loading control. External methods control point and nonpoint sources of pollution such as improving sewage treatment systems prior to discharge into bodies of water, or the inhibition phosphorus detergents and controlling the amount of pesticides used in agriculture (Liu and Qiu 2007). Internal methods consist of sediment dredging, water flushing, and aeration technology to regulate internal sources of pollution (Liu and Qiu 2007). In addition, planting aquatic macrophytes such as hyacinth, bulrush, cat’s-tail, and canna is a feasible method to refine eutrophic lakes. The roots of these plants provide aerobic and anaerobic conditions for bacterial growth and consequently remove nutrients by degrading nitrogen and phosphorus compounds (Liu and Qiu 2007). Laws can play a prominent role in control of water quality and decreasing eutrophication such as the “Law Concerning Special Measures for Conservation of the Environment of the Seto Inland Sea” of Japan which has substantially decreased red tides in this country since the mid-1970s (Imai et al. 2006).

Health Issues for Human Nitrogen is a basic element of amino acids and protein, and its use in fertilizers has led to the much needed increase of global food supplies (Wolfe and Patz 2002). However, if its concentration passes safe thresholds, it will adversely affect terrestrial, aquatic, and atmospheric domains (Wolfe and Patz 2002; Galloway et al. 2008).

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Particularly, nitrogen in the form of nitrite (NO2) and nitrate (NO3) in water can pose several hazards to human health. Methemoglobinemia Methemoglobinemia or blue baby syndrome which is a fatal blood disorder, especially in infants younger than 6 months, is mainly attributed to high concentrations of nitrite. Nitrates can also be reduced to nitrites. Nitrite ions oxidize soluble Fe2+ to insoluble Fe3+ as well as normal hemoglobin to methemoglobin in human blood (Wolfe and Patz 2002; Nitrate and nitrite in drinking-water 2011). Methemoglobin’s ability to carry oxygen molecules to the tissues is fundamentally lower than hemoglobin. According to the severity of a baby’s condition, the infant may develop bluegray skin or become lethargic. In extreme cases, methemoglobin concentrations higher than 50% can cause coma and death (Wolfe and Patz 2002). The normal methemoglobin level in humans is less than 2%; and in infants under 3 months of age, it is less than 3% (Nitrate and nitrite in drinking-water 2011). Carcinogenicity N-nitroso compounds are mostly carcinogenic compounds which are a result of nitrite and nitrosatable compounds reacting in the human or animal body (Nitrate and nitrite in drinking-water 2011). Several investigations about the correlation of nitrate or/and nitrite and nitrosatable compounds intake and cancer risk have been published which have concluded that high nitrate concentrations in water have a positive correlation with increasing gastric and/or esophageal cancer. However, a number of case-controlled studies have found other conceivable causes for the cancer (Wolfe and Patz 2002; Nitrate and nitrite in drinking-water 2011). Weyer et al. (2001) analyzed cancer incidence from 1955 through 1988 in a group of 21,977 Iowa women who were between 55 and 69 years old. Among 3150 cancer incidences, there was no correlation between nitrate concentrations in drinking water and cancers of the colon, breast, lung, pancreas, or kidney, while there were positive associations for bladder and ovarian cancers and negative relations for uterine and rectal cancers (Weyer et al. 2001). Finally, yet importantly, nitrate like similar anions prohibits iodine uptake and as a result causes antithyroid effects on the human body (Nitrate and nitrite in drinkingwater 2011). Studies in Slovakia, Bulgaria, Germany, and USA have found a relation between nitrate intake and thyroid function (Radikova et al. 2008; Hampel et al. 2003; Gatseva and Argirova 2008; Ward et al. 2010); on the other hand, a clinical study in the Netherlands did not find a correlation between nitrate intake and thyroid structure or concentration (Hunault et al. 2007; Blount et al. 2009). Due to the abovementioned hazards of nitrate and nitrite, the United States Environmental Protection Agency (USEPA) has set the Maximum Contaminated Level Goal (MCLG) for total nitrate/nitrite at 10 mg/l and for nitrite at 1.0 mg/l (US Environmental Protection Agency 1996). Similarly, too much phosphorus concentration can cause health problems such as kidney damage and osteoporosis. The Wisconsin Department of Natural Resources (WDNR) has set a phosphorus limit for point sources at 1.0 mg/l (NR 217).

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Chemical Oxygen Demand Increase Chemical Oxygen Demand (COD) is an indicator of wat quality which measures the amount of oxygen required for the oxidation of all organic substances in water. The COD of raw municipal wastewater in the United States is about 400 mg/l (can vary tremendously), phosphorus concentration is 6–10 mg/l, and nitrogen concentration is about 30–40 mg/l. According to Randall et al. (1998) which takes into account the growth of new biomass due to photosynthesis, releasing 1 kg of phosphorus into nature results in 111 kg of biomass with a COD of 138 kg; hence, 6 mg/l phosphorus produces 828 mg/l COD, more than double the COD of sewage. Also, 1 kg of nitrogen generates 16 kg of biomass with a COD of 20 kg; so, 30 mg/l of nitrogen in the untreated sewage is equivalent of 600 mg/l COD, more than the COD of the organic compound of raw sewage (Randall et al. 1998).

Types of Biological Nutrient Removal Systems Considering the adverse hazards of nitrogen and phosphorus in water as discussed in the previous section, efforts have been made to find the best practical, environmental friendly, and economical solutions to reduce nitrogen and phosphorus concentration, at least to the levels required by standards. Due to the characteristics of nitrogen and phosphorus, various chemical and physical methods have been proposed to remove these nutrients from water. Alternatively, biological methods for nutrient removal have also been investigated. There are three major chemical and physical processes for nitrogen removal, namely, ion exchange, air stripping, and breakpoint chlorination. In the first method, wastewater flows over an ion exchanger which is very selective to nitrogen ions (such as ammonium) in the presence of other ions like sodium, magnesium, and calcium. In this method, either synthetic or natural resins are utilized as ion exchangers. While natural zeolite clinoptilolite as a microporous structure of silica and alumina is a natural resin for nitrogen removal from water, modern ion exchangers are polymer based. In the air striping method, the pH of the water is increased, with, for example, lime to pH 10.5–11.5, then through an air-striping tower ammonium is removed from the water and desorbed into the air by providing enough air and water contact. In the breakpoint chlorination approach, sufficient amount of free chlorine (via chlorine gas or hypochlorites) is added to the wastewater in order to remove nitrogen-ammonium from wastewater by oxidizing it to nitrogen gas. Equation 1 shows the overall equation of this process.  þ NHþ 4 þ 1:5 HOCL ! 0:5 N2 þ 1:5 H2 O þ 2:5 H þ 1:5 Cl

(1)

Chemical and physical removal of phosphorus is fundamentally based on bringing phosphorus out from soluble state, and creating particles by adding coagulants and flocculants, followed by using solid-liquid separation processes like sedimentation or filtration for their removal. Aluminum and iron compounds are typical

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coagulants for the precipitation of phosphorus from wastewater. In this process, aluminum and iron compounds react with phosphate in the wastewater and generate FePO4 and AlPO4 which are highly inclined to sedimentation. On the other hand, biological processes have major advantages when compared to physical and chemical methods. Therefore, they will be focused on for the remainder of the chapter. These processes are divided into three main categories of: nitrogen removal processes, phosphorus removal processes, and combined nitrogen/phosphorus removal processes.

Biological Nitrogen Removal Processes The basic concept of biological nitrogen removal processes relies on nitrification and denitrification which are two major steps in the nitrogen cycle. Ammonium is biologically oxidized to nitrite, and nitrite is oxidized to nitrate in the nitrification process which is accomplished in aerobic reactors, in which oxygen is the electron acceptor. In the denitrification process, nitrate is reduced as an electron acceptor in anoxic reactors to nitrogen gas. Hence, a biological nitrogen removal system should consist of at least one aerobic and one anoxic reactor. Among the biological nutrient removal systems, the Ludzack and Ettinger process, the Modified Ludzack and Ettinger (MLE) process, and the Bardenpho process remove only nitrogen from wastewater. An oxidation ditch and batch reactor can remove nitrogen, if the amount of air transferred and the time of different phases are controlled. Biological Phosphorus Removal Processes Biological phosphorus removal (BPR) is basically derived from interpolation of phosphorus into cell biomass and subsequent phosphorus removal by sludge wasting. Phosphorus accumulating organisms (PAOs) are responsible for phosphorus consumption; so, reactors configuration in BPR systems should provide conditions for PAOs growth. Anaerobic condition in the absence of nitrate and oxygen is a mandatory first step in BPR systems in which PAOs take in volatile fatty acids (VFAs) and acetate and produce storage products such as polyβ-hydroxyal-kanoates (PHA) and polyhydroxybutyrate (PHB). Concurrently, orthophosphates are released from polyphosphate, and as a result the amount of phosphorus in the anaerobic condition is increased. The anaerobic condition is followed by an anoxic or aerobic zone in which storage products are oxidized and produce energy which is required for new microorganisms to grow. Meanwhile the dissolved orthophosphate is removed from wastewater and stored in biomass as polyphosphate. Finally, sludge wasting reduces the phosphorus concentration in the sewage. So generally, biological phosphorus removal process consists of an anaerobic reactor following by anoxic or aerobic reactors. A/O (anaerobic/aerobic) and PhoStrip processes are two such systems. Biological Combined Nitrogen/Phosphorus Removal Processes Biological combined nitrogen/phosphorus removal processes contain all the three main biological conditions, anaerobic, anoxic, and aerobic conditions.

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A2/O (anaerobic/anoxic/aerobic), the modified Bardenpho, and the University of Cape Town (UCT) processes remove both nitrogen and phosphorus simultaneously. Although this chapter will provide a general review of all these processes, it will specifically focus on the modified Bardenpho process.

Advantages of Biological Nutrient Removal Many chemical and physical methods have been used for nutrient removal in the past, but biological processes developed simple and practical alternatives for nitrogen and phosphorus removal. Each of these methods has its special merits and demerits. For instance, chemical and physical methods need a high amount of chemical compounds for nitrogen and phosphorus precipitation; on the other hand, biological processes require aeration and have a larger land footprint for the various anaerobic, anoxic, and aerobic reactors. Therefore, after considering the budget, the desired quality of effluent, the amount of available area for plant construction, and other aspects, the best and most practical approach should be selected. Nonetheless, biological processes have found relative popularity in recent years. During biological nutrient removal processes, nitrogen and phosphorus are removed from wastewater by utilizing microorganisms in various environmental conditions; in these processes, the amount of consumable chemical compounds critically decreases in comparison to chemical and physical methods. For instance, a small amount of acetate is required during biological phosphorus removal to improve biomass growth; on the contrary, a lot of aluminum or ferrous salts are needed for chemical and physical phosphorus removal. The biological nutrient removal, hence, is a more economical approach. Reduction in the amount of consumable chemical compound causes substantial decline in excess sludge volumes; hence, biological approaches have both economical and operational benefits. Furthermore, in most BNR methods, anaerobic and/or anoxic reactors are placed ahead of aerobic reactors which causes significant reduction in energy consumption in biological processes in comparison to not only chemical and physical approaches, but also other conventional biological methods like activated sludge. The presence of an anaerobic zone decreases the amount of required aeration in the aerobic zone since aerobic influent has negligible dissolved oxygen and according to mass transfer equations the oxygen transfer driving force is greater. Also, nitrate as an electron acceptor in anoxic zones causes COD stabilization. Biological phosphorus removal systems are the most economical choice for an effluent standard of 1 mg/l. Adalbert Oneke Tanyi, in a modeling study, illustrated that biological phosphorus removal can result in concentrations of as low as 0.4, while for higher phosphorus removal, auxiliary chemical addition is needed (Tanyi). Finally, anaerobic and anoxic zones in biological nutrient removal systems improve the Sludge Volume Index (SVI), a measure of sludge settling properties, defined as the volume of 1 gram of sludge after settling the aerated liquor for 30 min.

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The History of the Modified Bardenpho Process Nitrification processes were firstly studied in the nineteenth century but were altered after the invention of the activated sludge method by Arden and Lockett (1914). Sawyer and Bradney (1945) investigated excess sludge production during their work on nitrification and denitrification (Sawyer and Bradney 1945). Biological nutrient removal (BNR) from wastewater developed in the 1960s. Ludzack and Ettinger (1962) and Wuhrman (1964) evolved biological nitrogen removal systems. Ludzack and Ettinger’s process utilized biodegradable organic compounds in the influent as a carbon source for denitrification. Their system consisted of a series of anoxic and aerobic reactors which were not completely separated; hence, it caused lower control on wastewater flow between two reactors and also caused different performances in nitrogen removal. Levin and Shapiro (1965) examined biological phosphorus removal processes and developed the Phostrip system (Levin and Shapiro 1965). They illustrated that adding an anaerobic reactor in the wastewater treatment plant increases polyphosphate storage in microorganisms; so, they utilized an anaerobic tank in order to release phosphorus. They did not realize the role of carbon source and orthophosphates, but by adding an anaerobic reactor, they released phosphorus and used chemical methods and precipitation to remove the phosphorous (Fig. 1). Simultaneous biological nitrogen and phosphorus removal progresses were developed in the 1970s predominantly through the work of James Barnard in South Africa. Barnard (1973) proposed a modification for the Ludzack-Ettinger system in which he completely separated the anoxic and aerobic zones and returned a proportion of the wastewater from the aerobic reactor to the anoxic one. With this,

Aeration Tank Clarifier Influent Effluent

Returned Activated Sludge Excess Sludge Stripper Tank Chemical Sludge

Lime Feed

Fig. 1 The Phostrip process, Reproduced with permission from Randall et al. 1998

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he successfully removed nitrogen (Barnard 1973). Figure 2 illustrates the Modified Ludzack-Ettinger process (MLE). The importance of the anaerobic reactor was firstly realized by Barnard (1974) who studied previous investigations about biological phosphorus removal with activated sludge. Subsequently, he used an anaerobic reactor before an aerobic reactor and designed the Phoredox system which is illustrated in Fig. 3 (Barnard 1974). Fuhs and Chen (1975) proved that this mechanism is biological and not chemical (Fuhs and Chen 1975). Barnard developed several biological processes in order to remove nitrogen and phosphorus separately or together. He devised the 4-stage Bardenpho process in 1973 for biological nitrogen removal which is illustrated in Fig. 4. After that, Barnard in 1978 added an anaerobic zone before the 4-stage Bardenpho to remove

Recycle flow Aerobic Tank

Clarifier Influent Effluent

Aeration

Anoxic Tank

Excess Sludge

Returned Activated Sludge

Fig. 2 The MLE process, Reproduced with permission from Randall et al. 1998

Aerobic Tank Anaerobic Tank

Clarifier

Influent Effluent

Aeration

Returned Activated Sludge

Excess Sludge

Fig. 3 The Phoredox process, Reproduced with permission from Randall et al. 1998

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Fig. 4 The 4-stage Bardenpho process, Reproduced with permission from Randall et al. 1998

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Fig. 5 The modified Bardenpho process, Reproduced with permission from Randall et al. 1998

phosphorus from wastewater, and as a result, he invented the Modified Bardenpho. The 5-stage advanced (Modified) Bardenpho is shown in Fig. 5 (Randall et al. 1998). Marshal Spector (1979) realized that the anaerobic/aerobic configuration of activated sludge can remove phosphorus and invented the A/O (Anaerobic/Oxic) system which was similar to the Phoredox system. Also, he added an anoxic zone to the A/O process in order to remove nitrogen from wastewater. This system is known as the A2/O (Anaerobic/Anoxic/Oxic) process, and it was identical to a system configured by Barnard (Fig. 6) (Spector 1979). Table 1 briefly categorizes the historical development of biological nutrient removal systems.

Biological Phosphorus Removal Phosphorus is highly reactive and typically found in the form of PO4 or PO3 in nature. Phosphorus is a necessary element in the human body since it plays a role in energy distribution throughout the body as well as in the DNA molecule structure.

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Fig. 6 The A2/O process, Reproduced with permission from Randall et al. 1998

Table 1 Historical review of Biological Nutrient Removal (BNR) No. 1

Year 1914

Author(s) Arden–Lockett

2

1945

Sawyer–Bradney

3

1962

Ludzack–Ettinger

4

1964

Wuhrmann

5

1965

Levin–Shapiro

6

1973

Barnard

7

1974

Barnard

8

1975

Fuhs–Chen

9

1973

Barnard

10

1978

Barnard

11

1979

Marshall Spector

12

1979

Marshall Spector

Description Changes in the nitrification process with the invention of activated sludge Investigation of problems of increasing sludge in nitrification and denitrification Using organic biodegradable materials in the inlet flow as denitrification source (Ludzack-Ettinger process) Proposing a denitrification step after nitrification, in sludge systems for nitrification of high-load processes Developing a biological phosphorus removal system known as PhoStrip by adding an anaerobic reactor in the wastewater treatment plant Proposing a modification for the Lufzack-Ettinger system by separating anoxic and aerobic zones completely and using an internal recycle (modified Ludzack-Ettinger (MLE)) Designing the Phoredox system by using an anaerobic reactor before an aerobic reactor Proving that the Phoredox system is biological and not chemical Developing a 4-stage Bardenpho process for nitrogen removal Designing a 5-stage Bardenpho process (Modified Bardenpho) by adding an anaerobic zone as a first stage ahead of the 4-stage Bardenpho process Designing a process known as the Anaerobic-Oxic (A/O) process, which was identical to the Phoredox system for BPR Combining the A/O process with an anoxic zone in order to remove nitrogen (Anaerobic-Anoxic-Oxic (A2/O) process)

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Total phosphorus (TP) represents all forms of phosphorus, either dissolved or particulate. Orthophosphate (e.g., phosphate ion and phosphoric acid) and polyphosphates (e.g., pyrophosphate and trimetaphosphate) are soluble in water. The main difference between these two forms of phosphorus is that polyphosphate cannot be removed by precipitation with metal salts and it must first be changed to orthophosphates with biological activity (Moore 2010). The third form of phosphorus is organic phosphorus which is either dissolved or suspended in water. Particulate organic phosphorus can be removed by precipitation. Dissolved organic phosphorus is divided into biodegradable and nonbiodegradable, and the biodegradable portion can be converted to organophosphate with biological processes (Moore 2010). Unlike nitrogen or other compounds, the phosphorus cycle is limited in terrestrial and aquatic environments. Phosphorus is not naturally found in the gaseous state. Hence, phosphorus slowly cycles between the water, soil, and sediments. Phosphorus is typically in the form of phosphate salts in the soil or marine sediments, and it is a limiting factor for plant growth. Phosphorus is taken up from sediments by plants, then moves to animals’ bodies through eating, and finally, when plants or animals dye or urinate/defecate, the phosphorus moves to the soil or marine sediments again. Human activities like utilizing fertilizers can disturb this cycle.

Source of Phosphorus in Wastewater There is a wide variety of phosphorus sources in the wastewater. Domestic wastewater, especially in region with high detergent and cleaning products, is one of the main point sources of phosphorus. Also, septic tank wastewater disposal system is another possible source of phosphorus (Lee et al. 1978). Industrial and municipal wastewater are point sources of phosphorus. Protein-rich foods such as meat, nuts, beans, milk and soya have high amounts of phosphorus and consequently wastewaters from these industries have high concentration of phosphorus. Drainage from agricultural and urban lands, livestock’s excreta and atmospheric deposition are nonpoint sources of phosphorus which can increase this nutrient amount in the wastewater (Lee et al. 1978).

Importance of Phosphorus Removal Any added phosphate is almost immediately uptaken by plants, and only small amounts can cause exponential growth in plants. Phosphorus concentration in water in more critical since cyanobacteria (blooms of cyanobacteria and eukaryotic algae cause eutrophication) are capable of absorbing their required nitrogen from fixation of atmospheric nitrogen (N2) (Seviour et al. 2003); so, as mentioned in section “The Importance of Nitrogen and Phosphorus Removal,” phosphorus is the main cause of algal blooms in the aquatic environment.

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Chemical Removal of Phosphorus: Advantages and Disadvantages Chemical and physical processes in wastewater treatment rely on chemical and physical properties of elements such as reactivity, density, and mass transfer. Chemical processes for phosphorus removal are based on phosphorus reactions with for instance lime or metal salts (e.g., aluminum or iron). Aluminum salts, ferric salts, and lime react with the phosphate and form AlPO4, FePO4 and Ca5OH(PO4)3 which precipitate in water and consequently remove the phosphate from water. Aluminum sulfate (Al2(SO4)3) and ferric chloride (FeCl3) are widely used in comparison to other aluminum and iron salts. Chemical removal of phosphorus is a common undertaking in wastewater treatment plants (WWTP) which is implemented either alone or with biological processes to reduce costs and sludge production. Chemical methods provide higher phosphorus removal efficiency by consuming more chemical compounds. Also, chemical phosphorus removal has a lower sensitivity to environmental conditions such as temperature and wastewater influent composition in comparison to biological processes. These methods, however, are costly and produce more sludge, and they increase water salinity with unwanted by-product salts.

Biological Phosphorus Removal Process Enhanced biological phosphorus removal (EBPR) systems remove phosphorus economically, microbiologically, and in an environmentally friendly manner. Such systems are historically designed by putting an anaerobic zone ahead of anoxic and aerobic zones. For example, if an anaerobic zone is placed upstream of the MLE and Bardenpho processes which only remove nitrogen biologically, the resulting 3-stage Phoredox and Modified Bardenpho processes remove phosphorus as well (Seviour et al. 2003). It is generally believed that in the anaerobic zone, PAOs assimilate VFAs and convert it to PHA and also synthesize acetate to PHB which are stored in PAOs structure as an energy and carbon source for the subsequent steps. Meanwhile, polyphosphates are hydrolyzed to orthophosphates, and phosphorus concentration in the anaerobic zone is substantially increased. Figure 7 schematically shows this process in the anaerobic reactors. In the following aerobic (or anoxic) reactor, as shown in Fig. 8, PAOs utilize PHA and PHB to convert orthophosphates to polyphosphates and store them in their structure. Through sludge removal, the phosphorus concentration in the wastewater is decreased to the desired level. Oxygen and nitrate are, respectively, electron acceptors in the aerobic and anoxic zones, but in the anaerobic zone, PHA synthesis needs a source of reducing power (Seviour et al. 2003). There are various suggestions about derivation of the required electrons in anaerobic reactors. Comeau et al. suggested the tricarboxylic acid (TCA) cycle which is shown in the Fig. 9 (Comeau et al. 1986). Bacteria use three main procedures to maintain the proton motive force (pmf) (Comeau et al. 1986): (1) the first method drives out the H+ from the cell when a carbon substrate and an electron

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Fig. 7 Biological phosphorus removal microbiology – anaerobic zone, Reproduced with permission from (Seviour et al. 2003)

Fig. 8 Biological phosphorus removal microbiology – aerobic zone, Reproduced with permission from (Seviour et al. 2003)

acceptor (oxygen in aerobic zone and nitrate in anoxic zone) are present. In this process, nicotinamide adenine dinucleotide (NADH) plays the electron donor role and is produced via glycolysis and/or TCA cycle. (2) In the absence of electron acceptors, ATP breakdown at the ATP-ase site takes place to displace protons. (3) The enzyme NADH-transhydrogenase breaks down NADH to NAD+ to transfer protons. Mino et al. suggested that Glycogen degrades anaerobically to generate electrons for PHA synthesis (Mino et al. 1998). This process is illustrated schematically in the Fig. 10. Fig. 11 illustrates the PHB synthesis pathway. PHB is produced in a three steps synthesis pathway. In the first step, two Acetyl-CoA coupled to form Acetoacetyl-CoA

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Fig. 9 Bacterial strategies to maintain proton motive force, Reproduced with permission from (Comeau et al. 1986)

Fig. 10 Biochemical model for the anaerobic uptake of organic substrates and their conversion to PHA by PAOs, Reproduced with permission from (Mino et al. 1998)

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Fig. 11 PHB metabolic pathway, Reproduced with permission from (Kessler and Witholt 2001)

by β-ketothiolase catalysis. Then, Acetoacetyl-CoA reduced to (R)-3hydroxybutyryl-CoA by NADPH-dependent acetoacetyl-CoA reductase catalysis. Finally, (R)-3-hydroxybutyryl-CoA molecules polymerize to PHB. Dashed arrows illustrate the negative regulatory of HSCoA and NADPH on β-Ketothiolase and citrate synthase, respectively (Kessler and Witholt 2001). Acinetobacter bacteria, Aeromonas, and Pseudomonas are kinds of bacteria which can store polyphosphates and carbon in the form of PHB; hence, these bacteria are responsible for biological phosphorous removal (Comeau et al. 1986). Sidat et al. have prepared a table containing different organisms and their phosphorus uptake (Sidat et al. 1999).

Effects of Environmental Conditions on Phosphorus Removal Mamais and Jenkins (1992) examined the effects of temperature and solid retention time (SRT) on EBPR. Tests were performed as batch experiments at 10–37  C. The optimum temperature for aerobic phosphorus removal was between 28  C and 33  C. SRT values were in the range of 2–4 days, and EBPR efficiency was independent of SRT values until the SRT values were regulated above 2.9 days (Mamais and Jenkins 1992). Smolder et al. (1994) investigated pH effects on phosphorous release, which illustrated that P-release in the anaerobic zone is critically affected by the pH (varied between 5.5 and 8.5) resulting in a variation of 0.25–0.75 p-mol/C-mol (Smolders et al. 1994). Also, phosphorus release in the anaerobic zone is increased as the pH is increased (Mulkerrins et al. 2004).

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Dissolved oxygen (DO) in the anaerobic zone must be negligible (0.0–0.2 mg/l oxygen), and it is highly recommended that DO in the aerobic condition be in the range of 3.0–4.0 mg/l. DO values which are above of 4.0 mg/l are wasteful since it cannot induce biological phosphorus removal (therefore the energy used to increase the DO above 4.0 mg/l is wasted). The ratio of phosphorus to total organic carbon (TOC) is another important parameter. When low P/TOC in the influent is utilized, PAOs growth is suppressed, and obviously high P/TOC increases the growth of PAOs over glycogen accumulating organisms (GAOs) (Mulkerrins et al. 2004).

Importance of Acetate in Biological Phosphorus Removal Comeau et al., in a biological phosphorus removal plant, evaluated the concentration of phosphorus, oxidized nitrogen, and PHB of three samples during 8 h of experiment in three agitated but not aerated containers. Initially, three solutions of sodium acetate (0, 30, and 60 mg/l acetate as HAc) were added to the containers. Subsequently, a solution of sodium nitrate (10 mg/l nitrate as N) was added after 4 h. They illustrated the changes of these variables according to time. Regarding their observations, acetate presence and increase in its concentration causes enhancement in PHB production, and polyphosphates highly convert to orthophosphates and also increase reduction of nitrate to nitrite or gaseous nitrogen. While nitrate addition has opposite effects since phosphate and PHB are uptaken by biomass, and also denitrification and biological phosphorus removal (bio-P) bacteria enter the competition for available substrates (Comeau et al. 1987).

Biological Nitrogen Removal Nitrogen is a nutritious element for humans, animals, and plants, which occupies approximately 80% of the atmosphere. The main source of nitrogen in wastewater is from human activities such as cooking, bathing, and waste disposal. Nitrogen exists in the environment in the forms of organic nitrogen and inorganic nitrogen which consists of ammonium (NH4+), nitrate (NO3), nitrite (NO2), nitrous oxide (N2O), nitric oxide (NO), and gaseous nitrogen (N2). The various processes of the nitrogen cycle are fixation, ammonification, nitrification, and denitrification which allow for the circulation of nitrogen in the entire ecosystem. Figure 12 shows the nitrogen cycle in nature. Kjeldahl Nitrogen (KNT) is the sum of organic nitrogen and ammonium which was developed by Johan Kjeldahl in 1883. Nitrogen mostly enters wastewater in the form of organic nitrogen and ammonium. Ammonium is derived from breaking down of urea. Total Nitrogen (TN) is the sum of KTN, nitrate, and nitrite.

Source of Nitrogen in Wastewater Municipal, petrochemical, agricultural, food industry, paper making, leather making, artillery making, and slaughterhouse wastewaters and leachate of solid wastes have

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Fig. 12 The nitrogen cycle in nature, Reproduced with permission from (Stein and Klotz 2016)

various specific amounts of nitrogen. Fertilizers have great amounts of nitrogen in their structures which can pollute run-offs or ground waters.

Importance of Nitrogen Removal As it mentioned in section “The Importance of Nitrogen and Phosphorus Removal,” nitrogen causes some irreparable problems in the water such as eutrophication which increases water turbidity, decreases dissolved oxygen, and degrades water quality leading to human health hazards, especially for the cardiovascular system.

Chemical Removal of Nitrogen: Advantages and Disadvantages As explained in section “Types of Biological Nutrient Removal Systems,” there are several chemical and physical methods for nitrogen removal. In the following, the advantages and disadvantages of these methods are summarized. The advantages of the ion exchange method are insensitivity to temperature, ammonium removal up to required standards, operation even with substantial total dissolved solids (TDS), not forming harmful compounds; on the other hand, this method’s demerits are requiring reclamation, reduction of efficiency due to dissolved solids, sensitivity to organic compounds and particles, and high cost of construction and operation. Alternatively, the air stripping process removes ammonium and phosphorus up to intended standards and is not susceptible to toxic materials, but requires huge amounts of aeration and chemical compounds for pH regulation. Also,

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this process is sensitive to low temperatures. Breakpoint chlorination’s advantages are high removal of ammonium, demanding a small space, insensitivity to high temperatures and toxic materials, and disinfection of the effluent. On the other hand, considering chlorine consumption, this process is costly and produces trihalomethanes.

Biological Nitrogen Removal Process Biological nitrogen removal from wastewater follows two mechanisms: (1) biomass synthesis and (2) nitrification–denitrification.

Nitrification–Denitrification Nitrification The main source of nitrogen in sewage is ammonium. Nitrification is a biological process oxidizing ammonium to nitrite and then nitrite to nitrate. Bacteria responsible for the nitrification process are mostly autotrophs, while there are some heterotrophs as well. Autotroph bacteria utilize carbon dioxide as carbon source and gain their required energy from inorganic compounds, while heterotrophs gain their energy and carbon from organic compounds. The nitrification process consists of two steps. Firstly, ammonia oxidizing bacteria (AOB) oxidize ammonium to nitrite. These kinds of bacteria are nitroso organisms which include nitrosomonas, nitrosospiras, and nitrosococcus. The following equation shows the nitrification process. þ 2NH4 þ 3O2 ! 2NO 2 þ 2H2 O þ 4H þ New cells

(2)

Secondly, nitrite oxidizing bacteria (NOB) oxidize nitrite to nitrate. These kinds of bacteria are nitro-organism which include nitrobacter, nitrospira, nitrococcus, and nitrospina. The following equation shows the process:  2NO 2 þ O2 ! 2NO3 þ New cells

(3)

Koops and Pommerening-Röser (2001) investigated the distribution and ecophysiology of ammonia-oxidizing and nitrite-oxidizing bacteria. Ecophysiological parameters and preferred habitats of ammonia-oxidizing and nitrite-oxidizing bacteria are, respectively, illustrated in Figs. 13 and 14 (Koops and Pommerening-Röser 2001). Effects of Environmental Conditions on Nitrification Process

Nitrification is sensitive to pH and the rate of nitrification declines when pH values decrease, with nitrification practically ceasing under pH 6.0. The optimum pH is between 7.5 and 9.0 (Mulkerrins et al. 2004). Long aeration times and SRT induce complete nitrification. DO in the nitrification zone (aerobic condition) should be

The Modified Bardenpho Process

Fig. 13 Dendrogram based on 16S rDNA sequences of ammonia-oxidizing bacteria, Reproduced with permission from (Koops and Pommerening-Röser 2001)

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Fig. 14 Dendrogram based on 16S rDNA sequences of nitrite-oxidizing bacteria, Reproduced with permission from (Koops and Pommerening-Röser 2001)

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between 2 and 4 mg/l. Also, since the growth rate of nitrifiers is lower than heterotrophs, long SRT is mandatory for the nitrification process. Similar to phosphorus removal, the recommended mixed liquor suspended solid (MLSS) should be between 1500 and 1700 mg/l (Mulkerrins et al. 2004). Increase in temperature causes enhancement in the nitrification rate up to the optimum temperature of 30–35  C (Jeyanayagam 2005). Temperature affects microbial growth mechanism and mass transfer of oxygen. Nitrifiers are highly sensitive to environmental inhibitors such as heavy metals and some organic compounds (Tomlinson et al. 1966). Denitrification Nitrification is followed by denitrification in order to remove nitrogen from wastewater. In the denitrification process, nitrate is reduced to gaseous nitrogen (N2), and consequently N2 is released to the atmosphere. Denitrification process requires anoxic conditions. The following equation illustrates the overall denitrification process in which various reductions take place: nitrate to nitrite, nitrite to nitric oxide, nitric oxide to nitrous oxide, and nitrous oxide to gaseous nitrogen.  NO 3 ! NO2 ! NO ! N2 O ! N2

(4)

Bacteria responsible in the denitrification process are heterotrophs which include thiobacillus, micrococcus, serratia, and pseudomonas reducing nitrate and/or nitrite in the absence of dissolved oxygen. Pseudomonas are the most common denitrifiers which can utilize various energy sources. In the denitrification process, wastewater alkalinity increases and carbon dioxide decreases. Carbon dioxide reduction causes pH increase which is the opposite of the nitrification process; so, one of the most significant advantages of the denitrification process is pH regulating which inhibits corrosion of treatment plant facilities like piping and pumps. Effects of Environmental Conditions on the Denitrification Process

Optimum pH for the denitrification process is between 7.0 and 8.0 (Mulkerrins et al. 2004). Increasing wastewater temperatures induces microbial activity and as a result the rate of denitrification is enhanced. For instance, reduction in wastewater temperature from 20  C to 10  C reduces the denitrification rate by up to 75% (Jeyanayagam 2005). Due to the heterotrophic nature of denitrifiers, one of the most effective parameters on denitrifiers’ performance is the ratio of BOD or COD to TKN, with higher ratios of BOD to TKN causing higher denitrification (Jeyanayagam 2005).

Types of Biological Processes Different wastewater unit operations exist that utilize biological processes to remove pollutants from wastewater such as trickling filters (TF), rotating biological contactors (RBC), moving bed bioreactors (MBBR), activated sludge process

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(ASP), and aerobic and anaerobic fluid beds. According to the growth condition of microorganisms, these systems are divided into two general categories which are suspended growth and attached growth. In suspended growth processes, microorganisms are suspended in the wastewater during biological operation, and recycling of deposited biomass is mandatory. In attached growth, microorganisms are attached to a media such as polymeric, ceramic, or rocky bed to grow on their surface, and microorganism recycling is not necessary. Microorganisms form a microbial layer on the surface of the media and treat wastewater by this biofilm. Set-up, control, and managing biofilm systems are easier. Since microorganisms are attached to media, fluid flow cannot wash the biomass out of the wastewater treatment plant, and microorganism recycling is not required. Also, the effluent of the attached growth system has a better quality than suspended growth in terms of microbial contamination. Biofilm processes have an operational benefit that excess sludge production is lower in this process when compared to suspended growth processes. But regarding to utilized media, costs of biofilm processes are usually more than suspended growth systems.

Suspended Growth Processes Microorganisms in suspended growth processes such as the activated sludge process (ASP) are suspended in the unit in order to hydrolyze and consequently remove pollution such as organic matter, nitrogen and phosphorous. Microorganisms are mixed in the tank with aeration in aerobic zones, or with agitators in the anaerobic or anoxic zones. As wastewater flows through the suspended growth tank, food (COD or organic compounds) decreases due to microbial activity and cell mass is increased. For instance, in the ASP, biomass oxidize organic material to grow and form flocs of biofilms (Dabi 2015). The aeration not only provides oxygen as an electron acceptor but also mixes the suspended microorganisms. Subsequently, the mixture of treated wastewater and biomass flocs pass to the clarifier unit in which microorganism and wastewater flow are separated. The sludge is either returned to the aeration tank to increase the biomass concentration (MLSS) to the required level, or discarded when the biomass concentration in the tank is adequate. Figure 15 illustrates an activated sludge schematic.

Attached Growth Processes Biomass attach and grow on the surface of different media in attached growth processes. These media have different shapes and are made of different materials, all having high surface to volume ratios to support microorganism growth. Unlike suspended growth processes, the substrate and oxygen must transfer from the wastewater to the media and diffuse through the biofilm layers to be available for all microorganisms. Subsequently, products of microbial activity follow diverse routes from the biofilm to the wastewater (Jenkins and Sanders 2012).

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Activated Sludge Process Clarifier

Effluent

Influent

Aeration

Returned Activated Sludge

Excess Sludge

Fig. 15 Activated sludge process, Reproduced with permission from (Kayser 2005)

Kumar et al. (2009) investigated bacterial dynamics of living cells and dead cells in a biofilm membrane bioreactor by confocal laser scanning microscopy (CLSM) counts and considered 5 stages for microorganism growth and detachment during biofilm developments (Kumar et al. 2009). These steps are, respectively, cell attachment, pollutant limitation, biofilm stabilization and colonization, colonized biofilm, biofilm erosion. Burkholderia Vietnamiensis G4 bacteria (BVG4) are utilized to biodegrade the toluene as an impurity and also energy source in the wastewater. The investigations proved that living and dead cells have a direct relation with the efficiency of toluene removal (Kumar et al. 2009). Worden and Donaldson (1987) developed governing equations for the biofilm phase, wastewater phase and biomass growth in a well-mixed reactor for phenol oxidation. In this research, the biofilm grew on coal particles (Worden and Donaldson 1987). Horn et al. (2003) analyzed biofilm growth stages in homogenous growth, quasi steady state, and washout experiments (Horn et al. 2003). A trickling filter is a kind of attached growth processes in which wastewater is sprinkled on the top of the filter media and flows downward through the media due to gravitational force. The trickling filter provides an appropriate surface for microorganisms’ growth, and wastewater flow provides oxygen and energy sources for this purpose. Limitation in oxygen and energy source transfer, low pace of biofilm growth, and clogging of the filters are some of trickling filters’ problems (Fig. 16). Due to the above-mentioned problems associated with trickling filters, rotating biological contactors (RBCs) which consist of many disks rotating on one axis to provide the required surface for microorganism growth emerged in the 1960s and 1970s (Jenkins and Sanders 2012). In this system, due to the rotation of halfsubmerged disks around a shaft, microorganisms obtain their necessary carbon source from the wastewater and required oxygen from the air (Fig. 17). Up-flow anaerobic fixed bed (UAFB) is another attached growth process in which wastewater flows upward through the bed under anaerobic conditions. Zayen et al. (2016)

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Fig. 16 A trickling filter, Reproduced with permission from (Filipe and Grady 1998)

Fig. 17 A rotating biological contactor, Reproduced with permission from (Filipe and Grady 1998)

used UAFB for landfill leachate treatment. They concluded that UAFB has a great COD removal efficiency and biogas production (Zayen et al. 2016) (Fig. 18). A moving bed biofilm reactor (MBBR) is another attached growth system in which microorganisms grow on media as the wastewater smoothly moves through them, since their density is near water density. Unlike the trickling filter in which fixed media are utilized, in the MBBR system biomass grow on free-floating media such as small plastic carriers. These media are moved around the aerobic tanks by aeration and around the anoxic and anaerobic tanks by mixers (Fig. 19). This system

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Fig. 18 An up-flow anaerobic fixed bed, Reproduced with permission from (Zayen et al. 2016)

Fig. 19 A moving bed biofilm reactor. (a) Aerobic reactor, (b) anaerobic and anoxic reactor, Reproduced with permission from (Rusten et al. 2006)

was first developed in Norway (Seviour et al. 2003), and the media utilized in them were marketed as Kaldnes media which is shown in Fig. 20. MBBRs are very flexible systems which can be utilized for various purposes such as organic compound oxidation, nitrification, denitrification, and phosphorous removal. High microbial activity, relatively small reactors, low energy consumption, and low sludge production are some of important merits of moving bed biofilm reactors, while the cost of media might be a hindrance, and set-up and control of these systems take times and needs specific skills.

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Fig. 20 Kaldnes media

Fig. 21 An integrated fixed-film activated sludge process, Reproduced with permission from (Borner and Trubenbach 2017)

An integrated fixed-film activated sludge (IFAS) system is a combination of suspended and attached growth systems. This system provides excess biomass for an activated sludge process to enhance system capacity. This approach is mainly used for aerobic zones and can be used for upgrading systems to increase biological oxygen demand (BOD) removal and nitrification (Jenkins and Sanders 2012). Activated sludge is returned in this system to enhance biomass concentration in the tank, and microorganisms’ carriers are either fixed or free-floating media. Figure 21 illustrates an integrated fixed-film activated sludge.

Performance of the Modified Bardenpho Process The Modified Bardenpho Process is a biological approach combining both nitrogen and phosphorus removal process configurations together. This system consists of 5 distinct reactors in each of which special environmental

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conditions are provided for biomass growth and removal of wastewater impurities, especially nitrogen and phosphorus, with high efficiency. The Modified Bardenpho zones are, respectively: anaerobic reactor, first anoxic reactor, first aerobic reactor, second anoxic reactor, and second aerobic reactor. As it was mentioned in previous sections, phosphorus biological removal is highly dependent on PAOs’ function in absorbing energy and carbon sources and hydrolyzing polyphosphates to organophosphates in the anaerobic zone and using energy and carbon sources and up-taking polyphosphates in the aerobic zone; so, biological phosphorus removal units should include anaerobic and aerobic reactors. While for the biological nitrogen removal process, providing appropriate conditions for nitrification and denitrification is a mandatory requirement. As nitrification and denitrification processes, respectively, take place in aerobic and anoxic zones, these two conditions are required, too. Therefore, all three biological conditions – anaerobic, anoxic, and aerobic conditions – are necessary in simultaneous removal of nitrogen and phosphorus, but what was the reason of this sequence in the Modified Bardenpho process? The Modified Ludzak-Ettinger (MLE) process is a biological process to remove nitrogen with nitrification and denitrification processes. In this process, an aerobic tank is placed after the anoxic zone since if the aerobic zone is the first step, the oxygen concentration in sewage increases and disrupts the denitrification process. Also, in order to complete the nitrification-denitrification cycle, a recycle flow from aerobic zone to the anoxic zone is mandatory. Considering the MLE, Barnard devised a 4-stage Bardenpho by adding an anoxic and an aerobic reactor to the MLE process in order to enhance biological nitrogen removal so excess nitrate in the first aeration zone could be reduced to gaseous nitrogen. Subsequently, Barnard realized the importance of the presence of an anaerobic reactor in phosphorus removal; so, he decided to put an anaerobic zone ahead of the 4-stage Bardenpho to increase the efficacy. Wastewater flow through the modified Bardenpho process can be provided by gravitational force, while an important recycle flow from the first aerobic zone to first anoxic zone should be considered which is usually provided by pumping. Figure 22 illustrates the modified Bardenpho process and wastewater flow through all reactors. In the following, all 5 stages of the modified Bardenpho process are discussed in detail.

Reactors in the Modified Bardenpho Process Anaerobic Reactor The first reactor in the modified Bardenpho process is an anaerobic reactor (depleted of oxygen or oxidized nitrogen). In this reactor, mixers circulate wastewater in a tank in order to deliver carbon and energy sources to all the biomass.

1st

Anoxic Tank

Aeration

1st Aerobic Tank

Recycle flow

Fig. 22 The main flows in modified Bardenpho process

Influent

Anaerobic Tank

Pump

2nd Anoxic Tank

Returned Activated Sludge

Aeration

2nd Aerobic Tank

Clarifier

Excess Sludge

Effluent

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• Function: Anaerobic zones provide appropriate conditions for phosphorus accumulating organism (PAOs) growth and uptake of volatile fatty acids or other carbon sources such as acetate to produce storable compounds such as PHA and PHB. Also, polyphosphates are hydrolyzed to orthophosphates. The main purpose of this step in the modified Bardenpho process is increasing phosphorus removal efficiency. An important role of the anaerobic tank before the aerobic tank is that oxygen concentration in the sewage is very low and it can increase motive force of oxygen transfer in the aerobic tank. • Environmental conditions: Phosphorus release in the anaerobic zone increases as the pH increases. Also, ratio of the phosphorus concentration to the total organic carbon concentration has a direct relation with PAOs growth and phosphorus removal. • Contaminant concentration: Phosphate concentration in the anaerobic reactor is significantly increased since it is released from polyphosphates. BOD and COD amounts decrease since they are consumed as carbon sources. Nitrate, nitrite, and DO concentrations are very low in the anaerobic zone. Ammonia concentration in the anaerobic reactor is lower than the feed concentration but is higher than other subsequent reactors.

First Anoxic Reactor The second reactor in the modified Bardenpho process is the first anoxic reactor. Effluent of the anaerobic reactor and recycle flow from the first aerobic reactor enter this tank, and effluent of first the anoxic reactor flows to the first aerobic reactor. In this reactor, dissolved oxygen should be low in order for the biomass to utilize bonded oxygen in nitrate. If dissolved oxygen exists in the anoxic zone, biomass will reduce oxygen as an electron acceptor and after depletion of oxygen, reduce nitrate; hence, denitrification efficiency is declined. Mixers provide homogenous conditions in all parts of the anoxic reactor. • Function: The anoxic zone provides appropriate conditions for growth of denitrifiers. A group of heterotrophs such as pseudomonas reduce the nitrate and/or nitrite in the absence of dissolved oxygen. The anoxic zone provides conditions for the denitrification process to complete the nitrogen cycle and remove nitrogen as a gas (N2). As a result, the recycle flow from the first aerobic zone to this reactor is full of nitrate and denitrifiers reduce them in the denitrification process. • Environmental conditions: Temperature has a direct relation with denitrifiers’ performance. It goes without saying that increasing the temperature increases denitrifiers’ activity and denitrification. Optimum pH for the denitrification process is a rather neutral pH between 7.0 and 8.0. Also, a high ratio of COD to TKN stimulates denitrification. • Contaminant concentration: COD and BOD amounts in this reactor are substantially reduced since denitrifiers utilize BOD and COD in the influent to reduce nitrate to gaseous nitrogen. Nitrate and nitrite concentrations are very low in this zone as well. Ammonia and phosphorus concentrations are lower than the anaerobic reactor, but their concentration are still high.

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First Aerobic Reactor The third reactor in the modified Bardenpho process is the aeration or aerobic tank which is aerated with air compressors. Aeration provides dissolved oxygen as an electron acceptor in the biological nitrogen and phosphorus removal processes and also provides mixing in the tank. This reactor, as the heart of the modified Bardenpho process, plays the most important role in the performance of the unit. One reason is that the first aerobic reactor provides appropriate conditions for PAOs to use carbon sources such as PHA and PHB in order to absorb polyphosphates; so, phosphorus concentration is significantly reduced in this section. Also, ammonia is oxidized to nitrate through two reactions in which oxygen is as electron acceptor. In this way, the aerobic tank provides conditions for the nitrification process. Finally, a proportion of the effluent is sent to the first anoxic tank in order to reduce nitrate to gaseous nitrogen. • Function: The first aerobic reactor has two functions in the modified Bardenpho process. Firstly, this reactor completes biological phosphorus removal since orthophosphate in the sewage is absorbed and stored in biomass as polyphosphates. Secondly, this reactor helps the nitrification process in which nitroso organisms such as nitrosomonas cause ammonia to oxidize into nitrite and nitroorganisms such as nitrobacters cause nitrite oxidation to nitrate. In both these reactions, oxygen performs as an electron acceptor and its existence is mandatory. • Environmental conditions: Optimum pH for the nitrification process is near neutral conditions between 7.5 and 9.0, and a reduction in pH reduces the nitrification process. Dissolved oxygen concentration in the aerobic reactor should be between 2 and 4 mg/l. Optimum temperature is between 30  C and 35  C. • Contaminant concentration: BOD and COD amounts in the first aerobic zone are significantly reduced since PAOs, nitroso, and nitrobacter organisms need carbon sources for their growth and activity. Also, phosphorus concentration is substantially reduced to the desired standards. Ammonia concentration is greatly diminished since it is oxidized during the nitrification process, but nitrate concentration increases to its highest amount in the modified Bardenpho process. Nitrite concentration is increased a little, too.

Second Anoxic Reactor Actually, the main difference between the modified Bardenpho process and the A2/O or 3-stage Phoredox system, which consist of anaerobic, anoxic, and aerobic reactors, is the 4th and 5th steps of the modified Bardenpho process which are the second anoxic and second aerobic reactors. The fourth reactor in the modified Bardenpho process is the second anoxic reactor which provides another anoxic tank for unreduced nitrate which has not been recycled to the first anoxic reactor and has flown directly out from the first aerobic reactor; hence, it can improve biological nitrogen removal in the modified Bardenpho process in comparison to A2/O or 3-stage Phoredox reactors.

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• Function: The second anoxic zone truly enhances biological nitrogen removal in the modified Bardenpho process since unrecycled nitrate is reduced. • Environmental condition: The required environmental conditions in the second anoxic tank are similar to the first anoxic reactor. • Contaminant concentration: BOD and COD amounts in the influent of the second anoxic reactor are very low, but their amounts are reduced further in this step. If anaerobic conditions are created in some parts of the anoxic reactor, orthophosphate may release and phosphorus concentration may increase a little in this section, while in the following second aerobic tank released phosphorus is absorbed as polyphosphates and removed.

Second Aerobic Reactor The final stage in the modified Bardenpho process is an aerobic reactor which aerates wastewater in order to increase oxygen concentration and enhance wastewater quality. Also, the presence of oxygen can decrease phosphorus concentration, if any phosphorus is generated during the second anoxic tank or at the septic or clarifier tanks. • Function: As mentioned above, the second aerobic tank increases oxygen concentration in the wastewater to improve wastewater quality as the final stage and also inhibit phosphorus release in the next steps of the WWTP. In biological phosphorus removal, it is mandatory to waste sludge in the aerobic zone because sludge contains the maximum amount of the phosphorus (Mulkerrins et al. 2004). • Environmental conditions: The required environmental conditions in the second aerobic tank are similar to the first aerobic reactor. • Contaminant concentration: Similar to previous stages, microorganisms use up BOD and COD as a carbon source for their growth. Also, the phosphorus amount reaches its lowest value. Total nitrogen concentration in the influent and also the effluent of the second aerobic reactor is very low, and in effect all contaminations’ concentrations have reached their desired levels. In addition to the reactors discussed above, there are some other required facilities such as at least two pumps for wastewater circulation from the first aerobic reactor to the first anoxic reactor and from the precipitated sludge in the clarifier to the anaerobic tank to return sludge. Also, the two aerobic tanks need air compressors and diffusers, while the anaerobic and anoxic tanks require mixers. Wastewater recycle flow from first the aerobic to the first anoxic tank needs an adjustable reservoir tank since wastewater flow from the aerobic tank is saturated with dissolved oxygen, and DO should be minimized before entering the anoxic tank. Furthermore, similar to all wastewater treatment units, the modified Bardenpho process needs appropriate piping, heaters or coolers to adjust temperature and chemical compounds, and tanks for adjusting pH.

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Comparison of the Modified Bardenpho Process with Other BNR Systems A wide range of chemical and physical, environmental and operational parameters such as hydraulic retention time, influent distribution, temperature, pH, recycling ratio and so on determine the performance of biological nutrient removal (BNR) systems; therefore, comparison of the modified Bardenpho process function with other BNR systems requires an investigation in which nitrogen and phosphorus removal efficiencies are compared in equal or analogous aforementioned parameters. In what follows, biological nutrient removal efficiencies of different BNR systems will be compared by mentioning some experiments’ results which compare two or more processes and general comparison of advantages and limitations of BNR processes. As it stated before, BNR processes are divided into three categories. Some BNR processes like modified Ludzack and Ettinger (MLE) process include anoxic and aerobic tanks and remove only nitrogen. Austin and Nivala compared energy and area requirement of MLE process with some wetland processes like aerated, tidal flow and pulsed flow wetlands and concluded that MLE process has the highest energy requirement and needs the lowest area in comparison to wetland processes (Austin and Nivala 2009). On the other hand, the biological phosphorus removal processes are highly dependent on the presence of anaerobic condition prior to the anoxic or aerobic reactors. Lee et al. compared phosphorous removal between different BNR processes. They concluded that both anaerobic/aerobic (A/O) and anaerobic/anoxic/aerobic (A2/O) processes are dependent on influent COD/TP ratio, but in the same COD/TP ratio (=44) A/O process has higher phosphorus removal efficiency than A2/O. Therefore, A2/O process requires higher COD/TP ratio to remove phosphorus efficiently. Also, in comparison between A/O and Phostrip processes in lower COD/TP ratio (=25), Phostrip process phosphorus removal efficiency remains high while A/O removal efficiency is dramatically decreased (Lee at al. 1997). The third category of BNR processes remove nitrogen and phosphorus simultaneously such as modified Bardenpho process which provide all biological conditions to remove nutrients. In a comprehensive investigation, Ontiveros and Campanella compared various aspects of conventional, modified University of Cape Town and modified Bardenpho processes. In comparison with the conventional process, nitrogen removal efficiencies of modified University of Cape Town and modified Bardenpho process are respectively 60% and 85%, while phosphorus removal efficiencies are respectively 51% and 24%. Also, according to land use, climate change, eutrophication potential and some toxicity aspects, modified Bardenpho process illustrated the most environmentally friendly alternative (Ontiveros and Campanella 2013) Jeyanayagam (2005) and Linden et al. (2001) provided general comparisons between different BNR processes. According to these investigations, modified Bardenpho process has an excellent performance in biological nitrogen removal and lower efficiency in biological phosphorus removal. Also, large reactor volume requirement can be mentioned as one of the most significant limitations of this process (Jeyanayagam 2005; Linden et al. 2001).

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Modified Bardenpho Process Design The modified Bardenpho process as a biological system has various complications in design since it has 5 distinct bioreactors, and every reactor has especial functions and required conditions. For instance, even the two aerobic reactors or two anoxic reactors have different applications and responsibilities in the modified Bardenpho process, and as a result their specifications are different. Also, beside wastewater flow, there is a significant recycle flow from the first aerobic reactor to the first anoxic reactor which plays a prominent role in nitrogen removal. There are some imperative parameters in designing modified Bardenpho process as discussed below.

Nutrient Removal Establishment Start-up of every wastewater treatment unit, especially biological plants, takes time in order to habituate microorganisms with environmental conditions such as the COD/N/P ratio of the wastewater influent and the kinds of contaminations. Oldham and Stevens (1984) studied commissioning a new wastewater treatment plant with the modified Bardenpho process in Kelowna, British Colombia (Oldham and Stevens 1984). They illustrated that the biological nitrogen removal process (nitrification and denitrification) was established within 1 month of plant start-up, and nitrogen concentration in the effluent was lower than 5 mg/l with an influent of 30 mg/l. Also, 2 months of operating time was needed to steady out the phosphorus removal process, and phosphorus concentration declined from 6 mg/l in the influent to less than 0.5 mg/l in the effluent (Oldham and Stevens 1984). Due to the complexity of the modified Bardenpho process, a small physical problem can decrease nutrient removal of the plant and cause irregular results in some points. For instance, when air supply is out of service in first the aerobic tank it can reduce both nitrogen and phosphorus removal efficiencies.

COD and Nutrient Ratio Shortage of carbon, nitrogen, and phosphorus cause a wide range of problems, and each tank optimally functions at a specific C:N:P ratio. For example, the effects of three different influent C/N ratios on nutrient removal and the CO2, CH4, and N2O emission rates were investigated. Higher C/N ratios (5:1 or 10:1) have higher removal efficiencies rather than lower C/N ratio (2.5:1). Also, moderate C/N ratio (5:1) illustrated better greenhouse gas emission rates (Zhao et al. 2014). In another work, Puig et al. investigated biological nutrient removal by sequence batch reactor (SBR) technology and the effects of C/N/P ratio. According to the results, C/N/P ratioe of 100:12:1.8 in the influent caused the successful biological nutrient removal (Puig et al. 2007). Sewage of the modified Bardenpho plant of Oldham and Stevens (1984) had a COD: NH3-N ratio of 9:1 (Oldham and Stevens 1984). Also, for efficient biological phosphorus removal (phosphorus concentration in the effluent is lower than 1.0 mg/l), the COD: P ratio should be higher than 40:1 (Mulkerrins et al. 2004).

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Hydraulic Retention Time (HRT) HRT is a paramount and influential parameter in designing wastewater treatment plants since it designates the time that the wastewater remains in a reactor. HRT can determine a reactors’ volume and consequently required constriction space and costs. There is a common tendency to reduce reactors’ HRT, but hydraulic retention time should be sufficient for the removal of the targeted impurities; so, one of the most important design parameters that should be carefully evaluated is the HRT. Sakuma (2005) studied hydraulic retention time effects of 7 different A2/O systems on phosphorus and nitrogen removal efficiency. The anaerobic, anoxic, and aerobic reactors’ HRTs are regulated, respectively, between 1.4 and 3.3 h, 1.7 and 6.1 h, 7.1 and 9.9 h. The maximum phosphorus removal efficiency (91.4%) was obtained when anaerobic, anoxic and aerobic reactors’ HRTs are, respectively, 1.4, 4.8, and 7.7 h, and the maximum nitrogen removal efficiency (75.4%) was obtained when anaerobic, anoxic, and aerobic reactors’ HRTs are, respectively, 1.5, 6.1, and 9.4 h. HRT has critical influence on the BNR performance. Firstly, anaerobic reactor retention times should be optimized since if HRT of the anaerobic tank is very low, PAOs cannot produce storable carbon sources (PHA and PHB), and if the HRT of anaerobic tanks is very high, PAOs substantially utilize COD of wastewater which adversely affects the denitrification process. Also, long HRT is one of the main causes of secondary P-release which occurs in the absence of VFA and adversely affects the BPR process (Mulkerrins et al. 2004). Secondly, similar to the anaerobic reactor, the anoxic reactor retention time should be optimized, since short HRT of the anoxic tank causes improper denitrification and long HRT of anoxic tank harmfully affects phosphorus removal; denitrifying PAOs (DPAOs) are 40% less efficient in producing energy and also 20–30% lower cell yield in comparison to PAOs (Patel et al. 2005).

Sludge Retention Time (SRT) SRT is the average time which biomass stays in the system and is another important parameter in designing biological nutrient removal systems. There are various studies investigating SRT’s impact on BNR performances. Studies have illustrated that SRT has two opposite influences: longer SRT causes better effluent quality, while if the SRT is too long, it will cause high nitrification rates and a reduction in the poly-P microbial fraction of the mixed culture. There are many investigations about optimized SRT which have analyzed SRTs between 5 and 20 days, and have suggested that in BPR systems the optimized SRT is approximately 10 days (Mulkerrins et al. 2004).

Recycling Ratio (R) The recycling flow is an effective wastewater flow from the first aerobic reactor to the first anoxic reactor which completes the nitrification-denitrification process.

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This is because the generated nitrate in the aerobic tank is reduced in the anoxic tank to gaseous nitrogen and removed from the wastewater. The Recycling ratio (R) should be optimized according to influent properties (Patel et al. 2005). Nitrogen removal efficiency decreases with the increase of R at low COD/N ratios. At medium COD/N ratios there is a peak. As the nitrogen removal efficiency first increases and then decreases with the increase of R past the optimal point. At high COD/N ratios an increase in R also increases the nitrogen removal efficiency. Phosphorus removal efficiency generally increases with increase of R (Patel et al. 2005).

Temperature, pH, and Bubble Size Temperature and pH are two important parameters which should be measured during wastewater treatment and set to the optimized values at which microorganisms have their highest performance. Optimum temperature and pH for the anaerobic, anoxic, and aerobic reactors have been discussed in section “Reactors in the Modified Bardenpho Process.” Also, the modified Bardenpho process consists of two aerobic reactors which require aeration. One point regarding aeration needs particular attention: the bubble size. This is because small bubbles have more efficient mass (oxygen) transfer to wastewater, while large bubbles circulate wastewater or attached growth media better; so, bubble size should not be very large or very small and should also be optimized.

Modified Bardenpho Process Applications The main application and purpose of the modified Bardenpho process is nutrient removal. Meanwhile, due to the utilization of 5 biological stages, removing other impurities such as chemical oxygen demand (COD), biological oxygen demand (BOD), total suspended solids (TSS), heavy metals, and viruses can be considered as added benefits of the modified Bardenpho process. In the following sections, different applications of the modified Bardenpho process will be touched upon.

COD, BOD, TSS, Nutrient, and Heavy Metal Removal Emara et al. (2014) investigated the Fisha Selim wastewater treatment plant which utilized rotating biological contactors (RBC) during 2013, but the effluent quality was not pleasing. Average removal efficiencies of chemical oxygen demand (COD), biochemical oxygen demand (BOD), total suspended solid (TSS), total nitrogen (TN), and total phosphorus (TP) in this system were 82%, 86%, 63%, 54%, and

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50%, respectively. In order to improve the effluent quality, a 4-stage Bardenpho and a modified Bardenpho processes were designed. In this study, biological processes were suspended growth such as the activated sludge process. COD, BOD, TSS, TN, and TP removal efficiencies in the 4-stage Bardenpho process increased to 97%, 98%, 97%, 97%, and 50%, respectively, while these efficiencies in the modified Bardenpho process dramatically increased to 99%, 99%, 99%, 99%, and 90% (Emara et al. 2014). The modified Bardenpho process efficiency is on all accounts higher, particularly for phosphorus removal. Emara et al. (2014) also investigated nickel and iron contents in the wastewater influent and effluent from various systems. They observed that the removal efficiency of heavy metals from the Fisha Selim wastewater ranged between 10% and 30% in the RBC system (after 12 h), while the 4-stage Bardenpho and 5-stage Bardenpho, respectively, removed these heavy metals by 70 and 90–100% after 14 h (Emara et al. 2014).

Viruses Removal in Modified Bardenpho Process The initial focus and purpose of the modified Bardenpho process is biological nutrient removal, although a limited number of studies have pertained to virus removal as a valuable side effect of this process. A particular study compared virus removal via the modified Bardenpho process as a secondary wastewater treatment approach to other conventional aeration basin and trickling filter processes (Schmitz 2016). According to this study, the five-stage Bardenpho process is more effective at reducing viruses in wastewater than other conventional processes. Before 2014, two conventional biological processes, the trickling filter, and activated sludge aeration basin were utilized for secondary wastewater treatment in Tucson, Arizona. In recent years, however, Tucson WWTPs have started to utilize the modified Bardenpho process. In this way, a great opportunity for comparing pathogen removal efficiency of the modified Bardenpho process with other conventional biological processes came about (Schmitz 2016). In the mentioned study, the removal of 11 different virus types (pepper mild motlle virus, aichi virus, genogroup I, II, and IV noroviruses, enterovirus, sapovirus, group-A rotavirus, denovirus, and JC and BK polyomaviruses) was analyzed in a 12-month period in southern Arizona. The results indicated that the modified Bardenpho process had a higher efficiency in pathogenic virus removal than conventional biological processes. In addition, the pepper mild mottle virus seemed to be a useful process indicator in the investigation of virus removal during wastewater treatment (Schmitz 2016). According to this study, the five-stage Bardenpho process removes viruses by adsorption onto the biomass flocs in various anaerobic, anoxic, and aerobic zones. Actually, the modified Bardenpho provides more opportunity for virus adsorption to suspended particles; so, viruses are not killed, but rather concentrated in the solid phase and transformed to the sludge waste.

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Results illustrated that average virus removal efficiency in the modified Bardenpho process with sedimentation as preliminary treatment was 71.2% compared to 44.9% and 47.9% for other conventional setups. Combing the modified Bardenpho with a DAF system did not improve virus removal efficiency (Schmitz 2016).

Estimation of Greenhous Gas Emissions of Modified Bardenpho Process Recently, global warming has become one of the top concerns of all societies, and greenhouse gas emissions are hypothesized to play a prominent role in this phenomenon. Wastewater treatment plants are a sources of greenhouse gas emissions since during wastewater treatment, some greenhouse gases such as carbon dioxide (CO2), methane (CH4), and nitrous oxide (N2O) are produced. Sources of greenhouse gases are on-site and off-site in WWTPs. CO2, CH4, and N2O are by-products of biological activities, while greenhouse gases can be generated from off-site sources such as electrical and chemical operations that support the WWTPs (Kyung et al. 2015). Kyung et al. (2015) developed a model to assess greenhouse gas emissions from WWTPs. In this model, greenhouse gas emissions from a treatment plant consisting of a clarifier, a modified Bardenpho process, second clarifier, filter bed, and an ultraviolet disinfection system were determined. The aim of this WWTP was to remove carbon, nitrogen, and phosphorus from 5500 m3/day wastewater with 200 mg/l BOD. Each greenhouse gas emission is evaluated in every step of the modified Bardenpho process, and on-site greenhouse gas emissions occur only in the primary clarifier, modified Bardenpho process, and secondary clarifier but does to indirect emissions (chemical and electricity production, building materials, transport, etc.) all parts of the WWTP release off-site greenhouse gases. Table 2 illustrates greenhouse gas emissions from primary on-site resources. Based on this table, the first aerobic Table 2 On-site greenhouse emissions in a WWTP consisting of a clarifier, a modified Bardenpho process, second clarifier, filter bed and an ultra-violet disinfection system, Reproduced with permission from (Kyung et al. 2015) Unit process Primary clarifier Anaerobic 1st Anoxic 1st Aerobic 2nd Anoxic 2nd Aerobic Secondary clarifier Total

CO2 emission (kgCO2/day) 7.7  0.7

CH4 emission (kgCO2/day) 287  27

N2O emission (kgCO2/day) 71.2  7.9

Total (kgCO2/day) 366  36

1.3  0.1 1.8  0.2 3673  265 2.1  0.3 2.3  0.3 13.2  2.0

497  35 0.4  0.1 776  58 0.1  0.0 0.6  0.1 1.1  0.2

11.4  0.8 14.6  0.9 2646  247 12.9  2.8 6.5  0.7 238  29

510  36 16.8  1.2 7095  570 15.1  3.1 9.4  1.1 252  31

3701  269

1562  120

3001  289

8264  678

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reactor is the main source of greenhouse gases, especially CO2 since in this step, nitrification, BOD oxidation, and microorganism respiration are occurring. The anaerobic reactor generates CH4 more than other gases, while anoxic reactors produce higher amounts of N2O (Kyung et al. 2015).

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Sakuma M (2005) A2O process introduced to 7 WWTPs in regional sewerage office, Tokyo. Proc Water Environ Fed 2005(16):604–605 Sawyer CN, Bradney L (1945) Rising of activated sludge in final settling tanks. Sewage Works J 17:1191–1209 Schindler DW (1974) Eutrophication and recovery in experimental lakes: implications for lake management. Science 184(4139):897–899 Schmitz BW (2016) Reduction of enteric pathogens and indicator microorganisms in the environment and treatment processes. The University of Arizocna Seviour RJ, Mino T, Onuki M (2003) The microbiology of biological phosphorus removal in activated sludge systems. FEMS Microbiol Rev 27(1):99–127 Sidat M, Bux F, Kasan HC (1999) Polyphosphate accumulation by bacteria isolated from activated sludge. Water SA 25(2):175–179 Smolders GJF et al (1994) Model of the anaerobic metabolism of the biological phosphorus removal process: stoichiometry and pH influence. Biotechnol Bioeng 43(6):461–470 Spector ML (1979) US Patent 4,162,153, 24 July Stanier RY, Adelberg EA, Ingraham JL (1976) The microbial world, 4th edn. Prentice-Hall, Englewood Cliffs Stein LY, Klotz MG (2016) The nitrogen cycle. Current Biology, 26(3):R94–R98 Tanyi AO (2006) Comparison of chemical and biological phosphorus removal in wastewater–a modelling approach (Doctoral dissertation, Master’s thesis. Water and Environmental Engineering Department of Chemical Engineering Lund University, Sweden) Tomlinson TG, Boon AG, Trotman CNA (1966) Inhibition of nitrification in the activated sludge process of sewage disposal. J Appl Microbiol 29(2):266–291 US Environmental Protection Agency (1996) Drinking water regulations and health advisories Ward MH et al (2010) Nitrate intake and the risk of thyroid cancer and thyroid disease. Epidemiology 21(3):389–395 Weyer PJ et al (2001) Municipal drinking water nitrate level and cancer risk in older women: the Iowa Women’s Health Study. Epidemiology 12(3):327–338 Wolfe AH, Patz JA (2002) Reactive nitrogen and human health: acute and long-term implications. AMBIO J Hum Environ 31(2):120–125 World Health Organization (2011) Nitrate and nitrite in drinking-water: background document for development of WHO guidelines for drinking-water quality. Revised and Expanded Worden RM, Donaldson TL (1987) Dynamics of a biological fixed film for phenol degradation in a fluidized-bed bioreactor. Biotechnol Bioeng 30(3):398–412 Wuhrmann K (1964) Nitrogen removal in sewage treatment processes: with 9 figures in the text and on 2 folders. Internationale Vereinigung für theoretische und angewandte Limnologie: Verhandlungen 15(2):580–596 Zayen A, Schories G, Sayadi S (2016) Incorporation of an anaerobic digestion step in a multistage treatment system for sanitary landfill leachate. Waste Management, 53:32–39 Zhao Y, Zhang Y, Ge Z, Hu C, Zhang H (2014) Effects of influent C/N ratios on wastewater nutrient removal and simultaneous greenhouse gas emission from the combinations of vertical subsurface flow constructed wetlands and earthworm eco-filters for treating synthetic wastewater. Environmental Science: Processes & Impacts, 16(3):567–575

Diagnostic and Treatment by Different Techniques of Leachates from Municipal Solid Waste in Morocco Using Experimental Design Methodology

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Meriem Abouri, Salah Souabi, and M. Abdellah Bahlaoui

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Materials and Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . The Study Areas . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sampling and Recovery of Leachates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chemical Products . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Equipment and Materials . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Ventilation Device . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Anaerobic Treatment Device Combined with Aerobic Treatment . . . . . . . . . . . . . . . . . . . . . . . . . Measurement of Global Parameters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Experimental Design . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Results . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Characterization of Raw Leachate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Characterization of Settled Leachate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Application of Composite Plans for the Treatment of Leachate of MSW . . . . . . . . . . . . . . . . . Biological Treatment of Leachate from MSW . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cross-References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

The management and treatment of leachate issues in Morocco are of technological and environmental orders. Indeed, the pressure on the environment as on the quality of life of the populations and health conditions are becoming more and M. Abouri (*) · S. Souabi Laboratory of Process Engineering and Environment, Faculty of Sciences and Techniques, Mohammedia, University – Hassan II, Mohammedia, Morocco e-mail: [email protected] M. A. Bahlaoui Laboratory of Materials, Catalysis and Development of Natural Resources, Faculty of Sciences and Techniques, Mohammedia, University – Hassan II, Mohammedia, Morocco © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_88

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more alarming, which directs researchers to find adequate solutions to the leachate treatment with less expensive technology investment and exploitation. The present work shows the results of leachate treatment coming from two site: leachates from Municipal Solid Waste of the city Mohammedia and leachates from the landfill of the city Kenitra by the coagulation flocculation process, using several coagulants (FeCl3 a rejection of the steel industry rich in FeCl3 (SIWW), Al(Cl)n) and flocculants (Himoloc DR3000 and cactus), by biological processes (aerobic and anaerobic) and by physicochemical and biological techniques. This study aims to implement a simple technique and less expensive investment and operation for the reduction of pollution of leachate based on the Response Surface Methodology (RSM). The latter can significantly reduce the time and cost. Tests are carried out in the laboratory to assess the optimal conditions for effectively reducing pollution of leachate. To achieve this goal, the mastery of pollution is needed to quantify the different types of pollutants. Indeed, leachate discharges are rich in organic matter, NH4+, detergents, phenols, etc. The results showed that the pH without adding reagents has considerable influence on the reduction of pollution in terms of COD, BOD5, NH4+, detergents, phenols, color, organic matter measured at 254 nm. The comparative study of treatment of leachate from MSW of the city of Mohammedia and that of the discharge Ouled Berjal Kenitra by coagulation flocculation process has shown that the effectiveness of the latter is based on the quality of leachate selected for the study (stabilized leachate, slightly stabilized, young, etc.). The optimal concentrations evaluated over time will vary with the quality of leachate from Mohammedia and from landfill of the city of Kenitra. The comparative study of different coagulants showed that removal of pollution depends on coagulants, flocculants, pH, and function of the physicochemical characteristics of leachate. Sludge formation is high at basic pH (700 ml/L), whereas its production during coagulation flocculation is a function of nature of coagulants and flocculants added. The optimization of the Response Surface Methodology of coagulation flocculation process in the treatment of leachate from municipal solid waste of city of Mohammedia by studying the combination of coagulants and flocculants has allowed us the valuation of a biofloculant such as cactus and a wastewater of steel industry rich in FeCl3. The second technique concerns the treatment of leachate from Municipal Solid Waste by aerobic and anaerobic biological pathway, on the one hand by studying the evolution of indicators of pollution parameters such as COD, phenol, detergent, and the ammonia in the absence and presence of ventilation, on the other hand, the combination of anaerobic and aerobic treatment. The technique using the continuous and discontinuous aeration significantly reduced pollution leachate from the city of Mohammedia. Removal ranging from 90% to 60% of surfactant was obtained during aeration of leachate from MSW, while the phenol removal efficiency varies around 63%. Good reduction of COD and BOD5 during aeration was obtained during the study cycle. In addition, the treatment of leachate with a high organic load by the combined biological process – anaerobic for 43 days followed by intensive aeration at 3.5 L of air per

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minute per 10 l of leachate over a period of 30 days – has allowed elimination of 75%, 66%, and 60% of the COD, BOD5, and turbidity, respectively, after 73 days. Keywords

Leachate · Solid waste · Coagulation-flocculation · Biological treatment · Response surface methodology · Experimental design

Introduction In Morocco, solid waste production is steadily increasing as a result of population growth and the intensification of economic activity (Ez Zoubi et al. 2010). According to these authors, this production is estimated at 4.5 million tons per year of household waste and 800,000 tons per year of industrial waste. The rate of waste production varies from one region to another and can vary from 0.6 to 0.7 kg/inhabitant/day (El Maguiri et al. 2014; Souabi et al. 2011; Tazi 2001). The treatment of household waste, in particular, remains very little developed outside the landfill, almost universalized by Moroccan municipalities. This method can no longer survive in view of the significant damage it causes to the environment (pollution of groundwater and surface water, soil pollution, impact on human health, greenhouse gas emissions) (Idlahcen et al. 2014; Magda and Gaber 2015). In the interior of the wild dumps, the dumped wastes are only rarely totally inert and many physicochemical and biological reactions intervene between the waste and the environment in which it is located (rock, soil, groundwater, seepage water), but also within waste of various origins. The evolution of waste in landfills and their interactions with the external environment lead to the dispersion of pollutant flows, mainly through the emergence of leachate resulting from the dissolution of physicochemical and biological elements pollutants in water percolation. This water is loaded with organic and mineral substances and gives rise to leachates (Magda and Gaber 2015). The infiltration of these pollutants into the water table or their flow to the creek can lead to an insidious degradation of groundwater and surface water (Idlahcen et al. 2014). In fact, surface water pollution by leachates may occur by overflow and flow of liquids in the river system, either abruptly or gradually, whereas groundwater pollution is the result of infiltration and the spread of leachate in permeable or cracked subsoil (Souabi et al. 2010). The dissolution and chemical precipitation mechanisms that occur during infiltration are closely related to the physicochemical, biological, and hydrodynamic conditions of the aquifer system compartments traversed by the leachate (soil, unsaturated zone, and saturated zone of the aquifer), the latter regulating the attenuation, the transfer delay, the level of propagation, and the definitive retention of the polluting product (Baun et al. 2004). The risk of groundwater pollution from landfill leachate has been the subject of several studies. Thus, since 1960, some authors (Magda and Gaber 2015; Renou et al. 2008) have shown in their work that the pollution of the aquifer by the discharge is almost undetectable; others (Ragle et al. 1995) have shown the presence of a real hazard on many landfills studied in Wales and Canada.

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At the level in Morocco, several studies are made on the characterization and treatment of landfill leachates and household waste: Chofqi et al. (2004) have shown that the uncontrolled landfill of El Jadida generates stabilized leachates with a high nonweakly or poorly biodegradable, undrained, pollutant leachate, which seeps through the subsoil and contaminates the shallow circulating water table at the bottom of the quarries. Another study that was done by the Ministry of Land Use Planning, Water and Environment (MATEE) and the GTZ Management and Protection of the Environment Program (PGPE) in 2006 on the Tangier landfill shows that these leachates are generally loaded with highly colored organic matter and increased levels of heavy metals (Chouaouta 2002). They are charged bacteriologically and especially chemically with substances both mineral and organic. They can mix with both surface and groundwater and thus constitute a pollutant element both in terms of quantity and quality (ecotoxicological elements). El Kharmouz et al. (2013) studied the impact of leachates from the former Oujda city dump on groundwater and surface water. They have found that these leachates reveal high levels of biodegradable organic matter (BOD5 equal to 13,520 mg/l) and mineral matter (EC equal to 170 ms/cm) and that the analysis of ground and surface water has levels in NO3- (50 mg/l), chloride (400 mg/l), and iron, Zn, and Ni contents which far exceed the French standard values set for groundwater. Souabi et al. (2010) did a study on the characterization of leachate from the city dump of Mohammedia and Fez. The results of this study have highlighted the pollution generated by leachates that affects surface water and groundwater. The analysis of heavy metals in leachates has shown, according to the authors, a high concentration of chromium (5 mg/l) in the case of the Mohammedia discharge and 9 mg/l of that of Fes, while the Pb reaches a maximum concentration of 2.1 mg/l for the Fez landfill and 0.7 mg/l for the Mohammedia landfill. Following the impact of leachate on the environment through percolation into the groundwater and surface water pollution found in these previous studies, further work is being done to address this leachate treatment problem to minimize impact on the environment and human health. Benradi et al. (2013) studied the treatment of leachate at the Intercommunal Waste Center of Oum Azza (Rabat-Morocco) by reverse osmosis and the treatment of concentrate by the combination of coagulation flocculation using chloride ferric FeCl3 as chemical coagulant followed by decantation. These authors found that this treatment made it possible to obtain a clarified supernatant liquid with a decantation yield of up to 96%. Jirou et al. (2014) used another method of treating leachates from fresh household waste received at the Agadir city dump, using aeration to reduce the pollutant load of leachate. This technique of air injection has given variable results depending on the importance of the pollutant load and the chemical nature of the organic constituents. The purification yields are quite high with 99.3% reductions in COD, 99.1% of BOD5, 94.4% of total nitrogen, and 82.9% of suspended solids. Zalaghi et al. (2014) studied the characterization and treatment of leachate from the uncontrolled landfill in the city of Taza. They evaluated the technical efficacy and socio-economic performance of infiltration-percolation treatment on fixed natural supports appropriate to the local context. This technique results in a reduction of the suspended matter (MES),

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respectively, of the order of 90%, 93%, and 96%, a reduction of the chemical oxygen demand (COD) of the order of 91%, 89%, and 83%, respectively, and a reduction, respectively, of the order of 94%, 90%, and 91% in terms of the biological oxygen demand (BOD5) at the outlet of the ash filter, clay, and that of the mixture of the two. Benyoucef et al. (2015) have used another method of treating household waste leachate from the city of Kasbah Tadla, that is, the technique UASB (Upflow Anaerobic Sludge Blanket). The authors have shown that these leachates have a high concentration of organic matter in COD of 12,240 mg O2/l and a high acidity (pH = 5.16). According to the same authors, leachate treatment using the UASB technique showed a 93% reduction in COD and a shift from pH to neutrality (pH = 6.9). In addition, Bouaouine et al. (2015) studied the treatment of leachate from the controlled public landfill of the city of Fez by the method of electrocoagulation with the aluminum blades. The efficiency of the process is evaluated in terms of metal pollution, turbidity, sludge production, absorbance, and color removal. The results showed a significant elimination of metallic elements, in particular As, Cd, Cr, Cu, Fe, Mn, Ni, Pb, and Zn elements. The results also showed that the elimination of Cr, Zn, Ni, and Fe varies, respectively, by 79%, 86%, 89%, and 90%. Our present study targeted the leachate treatment produced by the cities Mohammedia and Kenitra. The inhabitants of Mohammedia city produce an average of 23,156 tons per year of waste, or 0.8 kg/inhabitant per day. The introduction of selective collection will enable the recovery of 7687 tons/year of glass, 5109 tons/year of plastic, 8670 tons/year of cardboard, and paper, and 1690 tons/ year of metals (El Maguiri et al. 2014). This household waste mixed with industrial waste is collected and transported to the public dump of the city Mohammedia. Before their evacuation, this waste is compacted in order to minimize transport costs. Leachate from compaction contains bacteriological, organic, and mineral pollutants. The purpose of our study is to characterize the physicochemical quality of leachates resulting from the compaction of municipal solid waste of Mohammedia city and the leachates produced by the Kenitra city landfill, physicochemical and biological treatment of effluent as well as improving the efficiency of pollution removal efficiency through modeling and optimization using the Surface Response Methodology (SRM) of the Coagulation Flocculation (CF) process. The works carried out comprise three distinct parts. The first part presented in this chapter deals with the materials and methods used for the treatment of leachates, description of the study areas (Mohammedia city and Kenitra city discharge), the sampling method and presentation of the different methods of analysis used, and treatment techniques such as coagulation flocculation, aerobic, and anaerobic treatment. In the second part, we touched on the practical side: results and discussion, which spread over five subchapters: The first subchapter deals with the characterization and treatment of the leachates of the DSM of the city of Mohammedia and that of the Kenitra city discharge by the flocculation coagulation process by studying the effect of pH and the effect of the coagulants used (FeCl3 40%, polyaluminum chloride and SIWW) in terms of concentration and composition.

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The second subchapter examines the evolution of the concentration of phenol and detergent in the leachate of the DSM of Mohammedia city in the absence and in the presence of aeration as well as the biological treatment of this leachate by continuous and discontinuous aeration. The third subchapter focuses on anaerobic treatment combined with the aerobic treatment of leachate from DSM Mohammedia city for the elimination of turbidity, color, phenolic compounds, COD, BOD5, and nitrogen. The fourth subchapter aims at the optimization by the Surface Response Methodology of the Coagulation Flocculation Process in the treatment of the leachate of DSM in the city of Mohammedia by studying the combination of coagulants (FeCl3 40%, SIWW, Al (Cl) n) and flocculants (Himoloc DR3000 and the cactus) in the treatment of this leachate as well as the valorization of a biofloculant which is the cactus and a FeCl3-rich rejection which is the SIWW.

Materials and Methods The Study Areas Mohammedia City Mohammedia, formerly known as Fedala, is located on the coast of the Atlantic Ocean, between two of the most important cities of Morocco: the economic city of Casablanca (25 km to the east) and the administrative city of Rabat (65 km to the west), of which it is connected by the coastal road 111 and the secondary road 107 in addition to the Casablanca-Rabat highway and a double-track rail link. The city covers 34.03 km2 and is home to a population of 188,619 according to the 2014 census. The urban territory is divided into limited geographical areas, traditionally called sectors. A containerization plan indicating the location, number, and capacity of the installed bins is assigned to each area. The collection of these bins is done by a truck benne equipped with a container lifter and a tank for the recovery of leachate resulting from compaction of waste. This compaction is done at a transit station before being sent to the landfill for the purpose of reducing the cost of transport. Leachate composition is a function of the amount and composition of household waste generated by the seasons and the type of habitat. (El Maguiri et al. 2014). Table 1 shows that the majority of the household waste components of Mohammedia city are organic materials varying around 44.36%. The objective of this study is to treat leachate resulting from compaction of this household waste. These very young leachates are rich in biodegradable organic matter. Kenitra City The city of Kenitra is located 45 km north of Rabat, the capital. With an area of 76 km2, it is the fourth largest industrial city in the country and the capital of Gharb.

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Table 1 Household waste components of the city Mohammedia (El Maguiri et al. 2014)

Component Organic mater Glass Plastic bags Hygienic textile Plastic Paper Cardboard Textile Wood Minority waste Metals Food packaging Green waste

1599 Average (%) 44.36 9.78 8.17 7.45 6.50 5.87 5.16 3.05 2.48 2.23 2.15 1.90 0.90

It is considered one of the largest Moroccan cities and one of the largest in the north west of the Kingdom. Indeed, the city is on the south shore of the Oued Sebou 12 km from the mouth on the Atlantic Ocean at Mehdya beach. It is located in a crossroads of trade routes, connecting the cities of northern and eastern Morocco. The population is about 1,061,435, the climate is subhumid to semi-arid influenced by the ocean, and the average rainfall is about 600 mm. The city of Kenitra is characterized by two major phenomena: • Unprecedented population growth • Rapid expansion of the city This translates into pressure on collective social amenities and hyperproduction of solid waste. Annual production of household and similar waste is around 108,000 tons in 2004 and 120,000 tones in 2011. To cope with the continuous increase in the volume of waste produced by its population, the Urban Commune has entrusted the collection and cleaning service to two professional operators: the company SEGEDEMA which operates in the Mamora zone and the company TECMED which ensures the service in the Saknia area (cost of collection and cleaning is in the order of 40 million DH and the cost of landfilling at the landfill is 14 million DH) (Bolt et al. 2005). Table 2 illustrates the physical composition of household waste at the Kenitra landfill.

Ouled Berjal Dumps It is located northwest of the city of Kenitra on the secondary road from Kenitra to Sidi Allal Tazi, on the left bank of the Oued Sebou estuary. It covers an area of 20 ha and receives, on average, 329 tons per day and 120,000 tons per year of waste. Waste management in the landfill is handled by SOS. Fig. 1 shows the Ouled Berjal dump located in Kenitra.

1600 Table 2 Physical composition of municipal solid waste at the Kenitra landfill (Aouane et al. 2010)

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Components Organic waste Plastic Paperboard metals Glass Metals Various (wood, etc.)

Average (%) 70 12 10 2 4 2

Fig. 1 The Ouled Berjal dump located in Kenitra city

The technical installations of this landfill consist of: • • • •

Closing Administrative room for staff Landfill area consisting of a 4 hectare bin that is subdivided into 4 cells Three large ponds for leachate collection and one reserve pond: – The first has a capacity of 21,945 m3 (105  55  3.8) and has built since the year 2010. – The second basin has a capacity of 7,363 m3 (39  59  3.2) and has been built since 2012. – The third basin has a capacity of 19,220 m3 (88  78  2.8) and has been built since 2013. – The fourth reserve basin has a capacity of 19,200 m3 (100  60  3.2) and has been built since 2014. • Conducting leachates and rainwater

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Constraints: • • • •

Location of the landfill in a vulnerable site Untreated leachate Ecological recuperators (scavengers) working in the informal sector Risk of saturation of the discharge

Sampling and Recovery of Leachates The Dump Trucks The leachates of the city Mohammedia were sampled from the dump trucks. These trucks were chosen randomly, each truck has a capacity of about 3 tons. The dumpsters are trucks adapted to the recovery of garbage. They are equipped with an automatic recovery system adapted to the tank, equipped with a compacting mechanism. They consist of a cab chassis to which the bodywork is attached, an ejector plate in the body, a hydraulic system and a control mechanism. All functions and the distribution of the load of the clamshell bucket are properly designed with respect to the technical specifications of the chassis of the truck on which the bucket is mounted. Fig. 2 shows an example of a dump truck. In Situ Sampling: Case of the Kenitra City Landfill Two types of leachate removal from the Ouled Berjal landfill were carried out: • Sampling at 5 collectors in series (R1, R2, R3, R4, and R5 leachate) one time • Sampling at 3 storage basins (B1, B2, and B3) in three times

Fig. 2 Dump truck

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Leachate sampling is done in situ using a container in for each basin and from each collector as shown in (Fig. 3).

Chemical Products Ferric Chloride FeCl3 40% Table 3 illustrates the physicochemical characterizations of the coagulant FeCl3 40%:

Fig. 3 In situ sampling using a container Table 3 Characterization of the coagulant FeCl3 40% Determination Chemical formula Appearance Ferric chloride (FeCl3) Iron (III) Iron (II) Manganese Insoluble matter Density at 20  C pH à 20  C Melting/freezing point Boiling point

Unit – – % mass percentage % mass percentage % mass percentage of iron content (III)

kg/dm3 –  C  C

Data FeCl3 Dark brown 39.0–41.0 13.4–14.2 3; 0.4/ FD = 200 >3; 0.027/ FD = 200 >3; 0.02/ FD = 200 >3; 0.022/ FD = 200 7.22 124 2320

>3; 0.4/ FD = 200 >3; 0.023/ FD = 200 >3; 0.018/ FD = 200 >3; 0.019/ FD = 200 7.65 156 1880

0.35

0.41

0.27

0.30

860 – – – – – – – – – –

780 – – – – – – – – – –

910 – – – – – – – – – –

650 200 1290 1.03 1.45 2.5 0.27 0.53 0.9 0.6 7.5

leachate. In addition, the observed values of turbidity vary between 1850 and 1200 NTU. These values characterize young leachates with a high content of colloidal matter and suspended matter. The latter (SM) varies between 21.61 and 7.22 g/l. Turbidity and conductivity score very high values far exceeding the standards of treated wastewater. Chemical oxygen demand accounts for most of the organic compounds and oxidizable mineral salts (Makhoukh et al. 2011). The COD values observed in this study range from 43,000 to 64,000 mg O2/l. The lowest

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concentration is observed for the leachate L1. A high COD value greater than 64,000 mg O2/l was observed for L3 leachate, indicating high organic load associated with these very young leachates in the acetogenesis phase, while high concentrations of ammonia nitrogen (200 mg/l) and NTK (1290 mg/l) were detected. The concentrations of ammoniacal nitrogen detected remain high compared to the results detected in the leachates studied by Idlahcen et al. (2014). The biological oxygen demand, which is an indicator of organic water pollution, expresses the level of biodegradability of the effluent (Makhoukh et al. 2011). The four leachates (L1, L2, L3, and L4) had BOD5 ranged from 30 to 21,000 mg O2/l. The lowest value is observed for leachate L1, and the highest value is observed for the leachate L3, which shows a correlation between the COD and the BOD5. Kjeldsen et al. (2002) showed that the BOD5 leachate studied had values up to 81,000 mg/l for fresh leachate samples and only 4200 mg/l for old leachate samples. The BOD5/COD ratio is a good indicator of the biodegradability of an effluent. A BOD5/COD ratio of less than 0.1 indicates low biodegradability. A BOD5/COD ratio greater than 0.3 indicates significant biodegradability. Mean leachate characterization indicated a BOD5/COD ratio of around 0.5 which shows that the leachate is rich in biodegradable organic matter. Souabi et al. (2010) showed that for releases of Mohammedia leachate (stabilized leachates) the COD/BOD5 ratio varies between 5 and 7.5, showing that the leachate is not easily biodegradable (rich in humic and fulvic substances) and can therefore cause several impacts on surface waters (Oued El Maleh). The same authors have shown that the COD values obtained vary between 2301 and 2750 mg/l and remain much lower than the content detected by Navarro and Veron (1992) discharged at sea. According to Courant and Aimar (1996), the BOD5/ COD ratio, which indicates the biodegradable character of carbon pollution, is of the order of 0.5 for young leachates and decreases to 0 for stabilized leachates. The results of leachates obtained showed significant absorbances at 254 nm, which is composed of phenolic compounds present in the leachate, the measurement of which sometimes requires significant dilutions (up to 200 times). The absorbances measured at 254 nm remain very important, which shows that the leachates are rich in phenolic substances. The leachate coloration is produced by the organic material; the distribution of the MO was determined by UV absorbance measurement at 254 nm.

Characterization of Settled Leachate For the study of the effect of decantation without any prior treatment, one liter of leachate from municipal solid waste was transferred in a graduated glass test tube or in cones to be decanted. Leachate is left decanted overnight (24 h). Different parameters were analyzed on decanted leachate. Table 8 illustrates the parameters analyzed: Decantation is a mechanical separation operation, by difference in gravity of immiscible phases of which at least one is liquid. It is possible to separate liquid phases, a solid phase suspended in a liquid phase.

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Table 8 Characterization of decanted leachate for 24 h Parameter pH O2 (mg/l) Temperature ( C) Conductivity (ms.cm1) Turbidity (NTU) DCO (mg O2.L1) DBO5 (mg O2.L1) DBO5/DCO Detergent (mg.L1) Phenol (mg.L1) Volume of decanted sludge (mL.L1) Absorbance ʎ1 = 254 nm ʎ2 = 436 nm ʎ3 = 540 nm ʎ4 = 660 nm Suspended matter (g.L1) Total phosphorus (mg P.L1) Nitrates (mg N. L1) NH4 (mg.L1)

Decanted leachate 5.6 0.52 23 10.64 980 40.500 20.200 0.5 340 15 116 >3; 0.2/FD = 200 >3; 0.020/ FD = 200 >3; 0.014/ FD = 200 >3; 0.015/ FD = 200 3 1300 520 120

Removal efficiency by the settling effect (%) – – – – 47 37 33 – 18 73 – 50 26 30 32 58 44 43 40

If a solid suspension is allowed to stand in a liquid phase, it is observed that the particles under the action of gravity and buoyancy of Archimedes tend to fall to the bottom or to rise to the surface according to their density and size. This settling can however be relatively slow for very fine particles (sensitive to thermal agitation) and particularly viscous liquids. Table 8 shows the characterization of decanted leachate for 24 h. For decanted leachate, the pH of the effluent varies around 5.6, while the value of the conductivity obtained is 10.64 ms/cm remains lower than that detected in the case of raw leachate. In addition, the COD value is 40,500 mg/l. These results show that the pollutant load generated by the leachate and reflected on the value of COD depends on the composition and the nature of the compacted waste often producing discharges of leachates rich in easily settled organic matter. In addition, the detected phenol concentration admits as a value 15 mg/l which is much lower than that detected in the raw leachate with a yield by simple decantation of 73%. The detergent concentration is 340 mg/l with a lower detergent removal efficiency (18%), which causes foams to disturb the aquatic environment if leachate effluents are not treated. The value of the BOD5 is 20,200 mg O2/l, for a BOD5/COD ratio greater than 0.3 indicates a significant biodegradability which justifies that the leachate is young.

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Assou et al. (2015) showed that decantation of leachate releases significantly reduces leachate pollution (approximately 30% COD). The low concentration of dissolved oxygen (O2 = 0.52 mg/l) in decanted leachate indicates that the anaerobiosis phenomenon is predominant. Indeed, oxygen is highly stressed for the degradation of organic matter and the oxidation of minerals present in the effluent during aerobic degradation. In addition, the recorded temperature is 23.2  C; this value will be favorable to the maintenance of colonies of “mesophilic” microorganisms that develop at a temperature between 20  C and 40  C (Kouamé 2007). Staining is high for decant leachate due to high turbidity and very high Suspended Matter content. The leachate examined contains high levels of mineral and organic matter. Significant levels of nitrogen compounds, total phosphorus, phenols, and detergents were also detected.

Application of Composite Plans for the Treatment of Leachate of MSW As a first approach, the experimental design can be conceived as a means of knowing which factors or interactions have a statistically significant influence on the response studied. For the leachate treatment by the coagulation flocculation process, it is known in several cases that the pH, the concentration of the coagulant, and the concentration of the flocculant have an influence on the elimination of the pollution. These are the three factors chosen for this study. The obtained model can only be used within the field of study, hence the usefulness of a correct preliminary study. The preliminary study allowed us to choose a pH range between 5 and 7, a coagulant dose range between 32 and 48 ml/l equivalent to 9.44 and 14.16 g/l, and a flocculant dose range between 8 and 16 ml/l. Table 9 shows the values of the coded (Xi) and natural variables (Xj) obtained after the preliminary study which make it possible to define the whole of the experimental domain. The step of variation (Δxi) is the range between two successive natural variables, divided by that of the coded variables. The steps of variation for each of the factors are collected in the same previous table relating to the experimental domain of the chosen factors.

Table 9 Experimental domain of selected factors Variables (Xj)

Unite

X1 = pH X2 = coagulant dosage X3 = flocculent dosage

– g/l ml/l

Coded variables: X1, X2, X3* a – 0 4.31 5 6 7.82 9.44 11.8 5.3 8 12

+ 7 14.16 16

A 7.68 15.75 18.7

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An experimental design is elaborated in this study, whose input factors present a large number of levels; it is a centered composite plan. In order to obtain a good descriptive quality of the model and to allow a reliable prediction in all the experimental domain considered for a reasonable number of tests, we realized the matrix of experiments of the Composite Centered Design with the rotation isovariance (Louvet and Delplanque 2005; Khoder 2011; Anouzla 2014) shown in Table 10. All the tests carried out and the results obtained in terms of reduction of COD, BOD5, color, turbidity, phenol, and variations in the conductivity (ms/cm) and sludge produced (ml/l) are grouped in Table 11. Exploitation of experimental results is often quite fast. The principle of exploitation is simple: it consists in calculating the coefficients of the polynomial model; the higher the absolute value of the coefficient, the more the corresponding term (simple factor or interaction) will influence the response studied. The difficulty is rather to be able to distinguish a true influence and the role of uncertainty inevitably tainting any measure. Recall that a factor is significant when its value of Snedecor estimated experimentally (Fexp) is greater than or equal to the value of critical Snedecor (Fc) at a level of confidence that we have chosen a priori equal to 95%. The sign of a parameter gives the direction of variation of the response. Indeed, when a factor goes from level (1) to level (+1), we say that its effect is positive when its sign is positive and that its effect is negative when its sign is negative (Louvet and Delplanque 2005; Khoder 2011; Anouzla 2014).

Effect on the COD Removal Table 12 summarizes the different factors influencing the removal of COD and their meanings. Table 10 Experimental Logical order matrix of composite plane 1 centered with rotation 2 isovariance 3 4 5 6 7 8 9 10 11 12 13 14 15 16

X1 1 1 1 1 1 1 1 1 1.68179 1.68179 0 0 0 0 0 0

X2 1 1 1 1 1 1 1 1 0 0 1.68179 1.68179 0 0 0 0

X3 1 1 1 1 1 1 1 1 0 0 0 0 1.68179 1.68179 0 0

Code  + + ++ + ++ ++ +++ a00 A00 0a0 0A0 00a 00A 0 0

X1 5 5 5 5 7 7 7 7 4.32 7.68 6 6 6 6 6 6

X2 9.44 9.44 14.16 14.16 9.44 9.44 14.16 14.16 11.8 11.8 7.83 15.77 11.8 11.8 11.8 11.8

X3 8 16 8 16 8 16 8 16 12 12 12 12 5.3 18.7 12 12

COD (%) 25.28 15.74 29.63 23.11 30.33 44.28 34.69 60.15 12.51 40.7 14.66 52.33 41.24 48.67 52.92 52.75

BOD5 (%) 61.76 57.35 64.71 61.03 64.71 72.06 66.91 79.41 55.88 69.85 55.88 76.47 70.59 73.53 76.47 76.47

Table 11 Model configuration and answers (coagulant SIWW) Color (%) 85.27 84.8 66.8 69.47 84.53 87.4 93.6 90.8 75.4 84.47 81.73 72.07 82.67 86.53 73.33 68

Turbidity (%) 70.98 75.68 71.8 79.5 73.5 73.63 86.63 84.45 58.93 69.65 81.35 87.25 80.5 80.5 80.05 80.35

Conductivity (ms/cm) 14.3 14.19 15.6 15.8 15.12 15.13 15.61 15.54 15.02 15.55 14.02 16.62 15.06 15.22 15.77 15.55

Sludge (ml/l) 88 68 82 78 160 126 148 120 80 190 40 54 75 36 40 41

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Table 12 Effects tests for COD removal Source of variation X1(5,7) X2(32,48) X3(8,16) X1*X2 X1*X3 X2*X3 X12 X22 X32

Degree of freedom a 1 1 1 1 1 1 1 1 1

Sum of squares b 1109.59 665.06 94.08 9.05 384.61 26.39 838.75 464.26 84.90

F-Value c

p-Value d

Rapport t

35.03 20.99 2.97 0.28 12.14 0.83 26.48 14.66 2.68

0.0010e 0.0038e 0.1356 0.6121 0.0131e 0.3965 0.0021e 0.0087e 0.1527

5.92 4.58 1.72 0.53 3.48 0.91 − 5.15 − 3.83 − 1.64

a

Degrees of freedom: an estimate of the number of independent categories in a particular statistical analysis or experiment b Sum of squares: the sum of the squares is a mathematical approach to determine the dispersion of the data points. The sum of the squares is used as a mathematical means to find the function that best suits (varies less) from the data c F-value: The Fisher test, or F-test, is a statistical hypothesis test that tests the equality of two variances by reporting the two variances and verifying that this ratio does not exceed a certain theoretical value that we look for in the Fisher table (or Snedecor table). If F is greater than the theoretical value, we reject the hypothesis of equality of the two variances d p-value is the probability of getting the same (or even more extreme) value of the test if the null hypothesis was true. The usual procedure is to compare the p-value with a previously defined threshold (traditionally 5%). If the p-value is below this threshold, the null hypothesis is rejected in favor of the alternative hypothesis, and the test result is declared “statistically significant” (Wasserman 2004). In the opposite case, if the p-value is greater than the threshold, the null hypothesis is not rejected, and nothing can be concluded about the hypotheses formulated e Significant at the 95% confidence level

Table 12 shows that the two factors: pH and coagulant concentration are significant for a threshold of 5%, that is, for a 95% confidence level. pH and coagulant dosage are also significant at 95%. These two factors and the interaction between them have an effect on the reduction of COD. On the other hand, the flocculant dosage has no effect since the value P is greater than 0.05. The purpose of this statistical test is to know if there are coefficients that are not influential, that is, to say that have no effect on each of the responses. In the case where one or more noninfluential coefficients exist on all the answers, they can be removed from the mathematical model in order to simplify it and improve its quality. Equation of the Model: For three factors, the realization of a complete 5-level factor plan requires 53 = 125 experiments. While the composite plane centered with rotation isovariance only requires 16 experiments. If only second-order interactions are taken into account, the percentage of reduction in COD will be represented by the following polynomial model (Jacques 1999; Anouzla 2014): ^ y ¼ b0 þ

3 X j¼1

bj X j þ

3 3 X X 0

j¼1j ¼1 j6¼j

0

bjj0 Xj Xj0 þ

3 X j¼1

bjj Xj 2

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This model comprises 10 terms: a constant term, three linear terms, three square terms, and three rectangular terms. With: ^ y: Answer function b0: Constant term of the model bj: Coefficients of the model of the variable Xj bjj: Coefficients of the model of the square variable Xj bjj0 : Coefficients of the model of interaction between the variables Xj, between Xj The gross equation for the percent COD reduction of leachate rejection with SIWW coagulant is as follows: %DCOØliminØ ¼ 52:93 þ 9:01X1 þ 6:98X2 þ 2:62X3 9:51X1 2  7:08X2 2 3:03X3 2 þ 1:06X1 X2 þ 6:93X1 X3 þ 1:82X2 X3 (1) The evolution of the percentage of reduction of the COD during the leachate treatment by the SIWW according to the volume of the coagulant and the pH of the medium to 12 ml/l of the flocculant Hymoloc DR3000, as well as the graph of the percentages of COD observed according to the predicted percentages, are defined in Fig. 10. The response surface materializes the regression surface from a graph in a threedimensional space. The graphical of the model equation makes it possible to illustrate the variations of the response and to identify an area of the experimental domain in which the response is interesting (Bakraouy et al. 2016). From the response surface graph, it is noted that the increase in COD reduction is according with increasing pH and coagulant dosage until an optimal of pH = 6 and coagulant concentration = 11.8 g/l. The mathematical analysis consists of estimating, by means of the least squares method, the model coefficients p, and the N residues, namely, the differences

Fig. 10 COD observed according to predicted COD and 3D surfac6&e profiler at 12 ml/l of flocculant

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between the measured (or simulated) values and the values predicted by the model for each of the experimental designs treatments. The coefficient R2 reflects the contribution of the model in the restitution of the variation of the observed response. By definition, the coefficient of determination belongs to the interval [0, 1]. In the presence of several explanatory variables, which is generally the case in the analysis of the test results, it is imperative to avoid using the coefficient of determination R2 to estimate the descriptive quality of the model. It is necessary to use the adjusted (adjusted coefficient of determination R2). If the number of experiments is equal to the number of unknowns in the system, the coefficient R2 will always be equal to 1. It is to avoid this that the adjusted coefficient of determination R2 has been introduced. This coefficient is defined as the difference in 1 of the ratio between the mean square of the deviations of the residuals and the average square of the experimental deviations. The closer the adjusted R2 and R2 values are to 100%, the better the descriptive quality of the model (Khoder 2011). From Fig. 10 it is observed that R2 = 0.945205. The adjusted square R equals 0.863012; these two values are close to 1.

Effect on BOD5 Removal From Table 13, we find that the pH, the coagulant dose, and the interaction between the coagulant dose and that of the flocculant are significant for the removal of BOD5. Fig. 11 presents the graph of percentages of BOD5 observed as a function of the predicted percentages, and the graph of the percentage change in BOD5 during leachate treatment by SIWW as a function of the concentration of the coagulant and the pH of the medium, at 12 ml/l of Hocoloc DR3000 flocculant. It can be seen that the actual values were distributed near the straight line, indicating that the model matched the measured values. As a result, this plot showed a sufficient agreement between the observed values of eliminated BOD5 and the values obtained from the model. Analysis of the results obtained in Fig. 11 shows that the reduction of BOD5 is due to the increase in pH and the coagulant dosage. At a pH equal to 6, a coagulant Table 13 Effects tests for BOD5 removal

Source of variation X1(5,7) X2(32,48) X3(8,16) X1*X2 X1*X3 X2*X3 X12 X22 X32

Degree of freedom 1 1 1 1 1 1 1 1 1

Sum of squares 279.06621 189.02336 20.43218 1.06580 97.58045 4.32180 135.58152 104.51114 25.78362

Value F

Rapport t

Value p

27.3179 18.5036 2.0001 0.1043 9.5522 0.4231 13.2721 10.2306 2.5240

5.23 4.30 1.41 0.32 3.09 0.65 -4.68 -3.57 -1.59

0.0020* 0.0051* 0.2070 0.7577 0.0214* 0.5395 0.0108* 0.0186* 0.1632

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Fig. 11 BOD5 observed according to the predicted BOD5 and 3D surface profiler at 12 ml/l of flocculant

concentration equal to 40 ml/l and a flocculant concentration equal to 12 ml/l; the percentage of removal of BOD5 can reach 78.5%. A very good correlation between BOD5 observed as a function of the predicted BOD5 and 3D surface profiler at 12 ml/l of flocculant was obtained. The gross equation of the BOD5 reduction percentage is written as follows: %DBO5ØliminØ ¼ 76:51 þ 4:52X1 þ 3:72X2 þ 1:22X3 4:92X1 2 3:75X2 2 1:67X3 2 þ 0:36X1 X2

(2)

þ 3:49X1 X3 þ 0:73X2 X3

Effect on the Turbidity Removal Variance analysis allows us to see if the variables selected for modeling have a significant effect on the response as a whole. The results of this analysis are summarized in Table 14. We find that the linear coefficients X1 and X2 as well as the cubic coefficients (coagulant and pH interaction) are significant for a threshold of 5%, which is to say for a 95% confidence level. Indeed, it is found that the pH and the dose of coagulant are factors influencing the elimination of turbidity. The regression is significant at a 95% confidence level. The quality of the model fit for turbidity removal was evaluated by the determination coefficient R2 = 0.9696. The 96.96% change observed for turbidity removal was attributed to the selected variables (pH, coagulant, and flocculant assays), while the model did not explain 3.04% of the variations (Abouri et al. 2014). Another way to evaluate the quality of the model fit is to plot the experimental values against the predicted values for turbidity removal. Fig. 12 shows these plots. As can be seen, the model represents the experimental data on the studied range. The graph shows the best fit as it can also be observed by the regression coefficient. Fig. 12 also shows 3D response surface for the elimination

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Table 14 Effects tests for turbidity removal

Source of variation X1(5,7) X2(32,48) X3(8,16) X1*X2 X1*X3 X2*X3 X12 X22 X32

Degree of freedom 1 1 1 1 1 1 1 1 1

Sum of squares 107.29176 108.60617 7.84386 46.60951 26.10031 0.05951 444.55745 26.18743 0.91071

Value F

Rapport t

Value p

26.7815 27.1096 1.9579 11.6344 6.5150 0.0149 110.9678 6.5367 0.2273

5.18 5.21 1.40 3.41 -2.55 0.12 -8.24 2.52 0.48

0.0021* 0.0020* 0.2113 0.0143* 0.0433* 0.9070 10,000 mg/l), high BOD5/COD ratios (>0.7), acid pH (typically between 5 and 6), and ammonium concentrations from 500 to 1000 mg/l (Bietlot et al. 2011). The latter comes mainly from the hydrolysis and fermentation of the protein compounds contained in the waste. Changes in the composition and concentrations of dissolved elements in percolates are often due to the age of the landfill facility (C.E.T). Pollutant load changes in percolates depend on the degree of stabilization of the waste and the volume of water that has infiltrated through the surface. Contaminant levels generally reach their maximum in the first years of operation of the landfill (within 2 or 3 years). This trend is especially marked for organic compounds, organic pollution indicator parameters (COD, a5, and TOC), the microbiological population, and most inorganic ions (heavy metals, Cl, SO42). Certain specific interactions between the materials present in the percolates and in the waste can also result in the solution of certain pollutants. Many organic compounds, containing nitrogen, oxygen, or sulfur, can form soluble metal complexes as ligands (or chelates) and thereby increase the total concentration of metals in the percolates. By way of example, humic and fulvic acids

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(responsible for the brownish coloration of percolates) are considered as complex complexing ligands and play an important role in the mobilization of heavy metals in the long term (Cd, Cu, Zn, etc.). Some inorganic ligands, such as chlorides, can also complex heavy metals (Cu, Ni, etc.).

Conclusion The disposal of solid waste is one of the main environmental issues in Morocco. More than five million tons of solid wastes are generated across the country, with a growth rate of annual waste production reaching 3%. Waste production is accompanied by the production of young leachate rich in organic matter with a high level of COD. In this work, we studied the decontamination of leachate discharges by two treatment techniques, namely, the physicochemical treatment by coagulation flocculation using different reagents on the one hand and the biological treatment in aerobic and anaerobic environment. Indeed, the diagnosis of leachates used for the study (leachate produced by the solid waste of Mohammedia) has shown that the discharges are too heavily loaded with pollutants in the form of BOD5 and COD. Moreover, the concentration of ammonia in the leachate greatly exceeds 1 g/l, thus giving bad odors and increasing the concentration of NTK. As for detergents and phenols the diagnosis has shown high levels in these elements which requires an effective treatment to reduce the levels of these elements. The treatment techniques was tested at the laboratory level, in particular the effect of the pH on the leachate of municipal solid waste in the city of Mohammedia of different polluting loads, which has shown a significant reduction of the organic matter in the form of chemical oxygen demand (COD) and biological oxygen demand (BOD5). Little difference was observed in the variation of the conductivity as a function of pH. A comparative study was carried out on the flocculation coagulation by FeCl3 40% (commercial product) and a FeCl3-rich (SIWW) discharge produced at the Maghreb Steel industrial unit. The latter was valued for reducing the pollution of leachate discharges used for this study. The results obtained showed a very interesting performance as a function of pH, the optimum for the reduction of organic pollution in the form of COD and BOD5, and turbidity varies from one companion to another and the formation of sludge depends the composition of leachate. The commercial 40% FeCl3 coagulant removed 87% of the COD at a rate of 14.4 g/l FeCl3, while the SIWW coagulant removed only 45% of the COD. For sludge production and turbidity removal, little significant difference was observed for 40% FeCl3 and SIWW rejection loaded with FeCl3. The study of the elimination of the pollution of young leachate coming from the municipal solid waste of the city Mohammedia using the technique of continuous and discontinuous aeration was carried out in order to evaluate the effectiveness of

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a biological treatment (at low cost) for this type of effluent. A total of 90-60% removal of surfactant was obtained during discontinuous and continuous aeration, respectively. The phenol concentration decreased from 600 to 220 mg/l which represents 63% of the removal of phenol by aeration. This has been accompanied by a change in pH and sludge biodegradation. Discontinuous aeration for a period of 9 days achieves a COD and BOD removal efficiency of up to 44% and 39% which corresponds to the removal of 20 g/l of COD and 7 g/l of BOD5, respectively. The treatment of leachates of high organic load anaerobically for 43 days followed by intensive aeration at 3.5 L of air per minute in 10 l of leachate over a period of 30 days yielded variable results depending on the importance of the pollutant load and the nature of the inorganic constituents. According to the results obtained in this work, we can conclude that the leachate of municipal solid waste of the city of Mohammedia is loaded with easily biodegradable organic matter. Biological treatment by anaerobic process followed by an aerobic leachate process has yielded remarkable results in terms of pollution reduction. COD, BOD5, and turbidity removal percentages were 75%, 66%, and 60%, respectively, after 73 days with initial concentrations of 58,880 mg/l, 24,000 mg/l, and 2036 NTU for COD, BOD5, and turbidity, respectively. Devices used in the experiment for anaerobic treatment and aerobic treatment can be considered effective, easy, and less expensive. The use of oxygen diffusers in aerobic treatment is very interesting to obtain maximum performance, efficiency and incomparable quality of oxygenation. These diffusers significantly reduce the costs of using pure oxygen. The best way to oxygenate the leachate is to minimize the size of the air bubbles, the smaller the bubbles, the more effective the oxygen transfer.

Cross-References ▶ Advanced Treatment Technologies ▶ Biostimulation and Bioaugmentation: An Alternative Strategy for Bioremediation of Ground Water Contaminated Mixed Landfill Leachate and Sea Water in Low Income ASEAN Countries ▶ Hazardous Waste Management with Special Reference to Biological Treatment ▶ Management of Municipal Solid Waste in Morocco: The Size Effect in the Distribution of Combustible Components and Evaluation of the Fuel Fractions ▶ Technologies for Treatment of Colored Wastewater from Different Industries ▶ Wastewater Management to Environmental Materials Management Acknowledgments The work done in this study was funded by the Ministry of the Environment as part of a leachate treatment project at the Kenitra City Wastewater Facility, which we warmly thank.

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Application of Liquid ChromatographyMass Spectrometry for the Analysis of Endocrine Disrupting Chemical Transformation Products in Advanced Oxidation Processes and Their Reaction Mechanisms

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Jin-Chung Sin, Sze-Mun Lam, and Abdul Rahman Mohamed Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chromatography Coupled with Mass Spectrometry for Intermediate Products Detection . . . Advanced Oxidation Processes for EDCs Degradation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Photolysis for EDCs Degradation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Photocatalysis for EDCs Degradation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sonolysis for EDCs Degradation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Iron Catalyzed H2O2 Production for EDCs Degradation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion and Future Trends . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cross-References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

Advanced oxidation processes (AOPs) have gained a great deal of attention as they held great promise for the treatment of wastewater contaminated with noneasily removable organic pollutants. The assessment of transformation products allowed the overall efficiency of AOPs to be better understood since some transformation products possessed higher toxicity than the mother compounds. Liquid chromatography coupled to mass spectrometry has been heavily used as

J.-C. Sin (*) Department of Petrochemical Engineering, Faculty of Engineering and Green Technology, Universiti Tunku Abdul Rahman, Kampar, Perak, Malaysia e-mail: [email protected] S.-M. Lam Department of Environmental Engineering, Faculty of Engineering and Green Technology, Universiti Tunku Abdul Rahman, Kampar, Perak, Malaysia A. R. Mohamed School of Chemical Engineering, Universiti Sains Malayisia, Engineering Campus, Nibong Tebal, Pulau Pinang, Malaysia © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_90

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an analytical tool in the environmental field as it allowed the best performance in assigning the identity of the transformation products detected. This chapter reviews liquid chromatography-mass spectrometry for the identification of endocrine disrupting chemical (EDC) transformation products formed from several AOPs such as photolysis, photocatalysis, sonolysis, and Fenton treatment. EDCs are a group of special interest due to their ability to exert hormonal imbalance activity and consequently caused adverse health effects in an organism or its progeny. Data concerning to the mechanism of AOP-mediated reactions on the EDCs degradation are also discussed in this chapter. Finally, the future prospects of AOPs on EDCs degradation are summarized and discussed. Keywords

Advanced oxidation process · Endocrine disrupting chemical · Degradation intermediate · Liquid chromstography coupled to mass spectrometry

Introduction Endocrine disrupting chemicals (EDCs) are defined as exogenous substances that can alter the functions of the endocrine system and consequently led to adverse health effects in an organism or its progeny or subpopulations (Avasarala et al. 2011; Sin et al. 2012). A plethora of the EDC substances can be classified into two categories: (1) natural hormones that are naturally present in the environment and (2) hormones that are fabricated. Their general nomenclature and representative examples are shown in Table 1. The classification of EDC substances is depended on mode of their endocrine actions including (1) mimicking the action of endogenous hormones, (2) antagonizing hormone receptors, (3) disrupting the hormones synthesis, metabolism, transport, and excretion, and (4) altering natural hormone production pathways. In addition, some EDCs have demonstrated multiple modes of action, which can lead to deleterious effects on the earth ecosystems (Eertmans et al. 2003). Myriad examples of reproductive and developmental abnormalities related to EDCs exposure have been revealed over the years in a broad spectrum of wildlife including mollusks, fish, reptiles, birds, and mammals (Esplugas et al. 2007; Skinner et al. 2011; Mennigen et al. 2017). According to Caliman and Gavrilescu (2009), the total number of compounds suspected of interacting with the endocrine system was approximately 38,000. Nevertheless, there were more than 80,000 chemicals required to analyze and investigate their endocrine actions. The EDCs are typically detected in the environment at trace concentrations (ppm to ppb), which are still extremely toxic to aquatic biota (Sin et al. 2012). Sources by which the EDCs entered the environment mainly from sewage treatment plant (STP) effluents although other sources such as direct discharge, leakage from septic tanks, and run-off from agricultural lands (Laganà et al. 2004; Liu et al. 2009; Burkhardt-Holm 2010). Wastewater treatment plants operated with technologies such as nanofiltration, reverse osmosis, lagoon or pond stabilization, biological treatment with chlorination,

Class Steroids and natural hormones

Empirical formula C18H24O2

C20H24O2

C18H24O3

C18H20O2

C18H22O2

Representative EDC Estrone

17α-Ethinylestradiol

Estriol

Trendione

17α-Trenbolone

Table 1 Classification of various EDC compounds with representative examples

270.37

268.35

288.38

272.38

Molecular weight (g/mol) 270.37

(continued)

Chemical structure

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Phthalates

Class Polycyclic aromatic hydrocarbon (PAHs)

Table 1 (continued) Empirical formula C12H8O C20H12

C10H8 C14H10 C10H10O4

C10H10O4

C16H22O4

Representative EDC Dibenzofuran Benzopyrene

Naphthalene Phenanthrene Dimethyl phthalate

Diethyl phthalate

Di-n-butyl phthalate

278.34

194.18

194.18

178.23

128.17

252.31

Molecular weight (g/mol) 168.19

Chemical structure

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Pesticides

Parabens

Alkylphenols and Bisphenol A

C15H16O2

C14H22O

C8H8O3

C9H10O3

C14H12O3

C6Cl5OH

C12H10O

C12H18N2O

Bisphenol A

4-Tert-octylphenol

Methylparaben

Eythlparaben

Benzylparaben

Pentachlorophenol

o-phenylphenol

Isoproturon

206.28

170.21

266.34

228.24

166.17

152.15

206.32

228.29

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oxidation by permanganate (VII), combined ozone, and filtration in sewage treatment plants currently are used on the elimination of EDCs (Acero et al. 2010; Al-Rifai et al. 2011; Stalter et al. 2011; Qiang et al. 2013; Zhang et al. 2013). However, most reported studies have shown that application of advanced oxidation processes (AOPs) provided a high degree of EDCs degradation in a short period of reaction time (Lau et al. 2007; Lam et al. 2013; Zhang and Li 2014; Simsek 2017). Most importantly, AOP methods can convert the EDCs into low molecular weight intermediate products and eventually led to mineralization to CO2 and H2O. Furthermore, there is no tendency to produce secondary pollution in AOP methods. Examples of AOPs applied to the oxidation of organic pollutants are photolysis, photocatalysis, sonolysis, photo-Fenton, and Fenton oxidation. However, many intermediate products formed during the EDCs degradation could be more toxic than the parent compound been treated. Identifying the intermediate products generated in the course of the AOPs degradation of pollutants in water is thus gaining attention to understand the degradation pathways and the degradability of the intermediates produced. Recently, liquid chromatography coupled with mass spectrometry (LC-MS) has become a highly appropriated technique for by-products identification. Time-of-flight MS (TOF-MS) and hybrid quadrupole TOF (QTOF)-MS/MS systems in combination with ultra-high performance LC (UPLC) have also received preference in separating, monitoring, and identifying of intermediate products generated in wastewater treatment (Gu et al. 2011; Souissi et al. 2012; Zhu et al. 2012; Sirtori et al. 2014). Hitherto, characterization and identification of degradation products are still challenging tasks for researchers working in this area. On the basis of the above consideration, this chapter provided a comprehensible review dealing with the use of LC-MS techniques in analyzing the different degradation products generated during the AOP-mediated EDCs degradation and to reveal their possible degradation pathways.

Chromatography Coupled with Mass Spectrometry for Intermediate Products Detection During the EDCs degradation process, several degradation intermediates were formed and needed identification using a proper analytical method. Identification of these intermediates could give a further insight into the mechanism of EDCs degradation and can assist to obtain a total picture of the degradation pathway with their final mineralization products. The desired final mineralization products for a complete degradation process are CO2 and H2O. The most commonly used analytical method is the chromatography coupled with MS (Souissi et al. 2012; Khataee et al. 2016). Gas chromatography-mass spectrometry (GC-MS) has been well recognized as the “gold standard” for both identification and quantification of nonpolar and volatile pollutants due to its high selectivity and sensitivity (Khataee et al. 2016; Ba-Abbad et al. 2017; Huang et al. 2017). However, its application was limited to some important problems such as the loss of analytes

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during the time-consuming derivatization process and its incompatibility with polar, thermally unstable and nonvolatile compounds. In recent years, LC-MS has therefore become analytical method of choice for identifying of degradation intermediates in aqueous solution. The LC-MS spectra can give information about the mass of the molecule through the quasi-molecular ion (MH+). Therefore, the direct coupling of LC and MS has developed into a powerful method to analyze the evolution and breakdown of intermediate products in a degradation process. This coupling method was also beneficial to separate and detect polar compounds without a derivatization step (Medana et al. 2005). The development of ultra-high performance LC (UPLC) equipped with sub-2-μm-particle-size reversed-phase column also provided faster analysis duration, better resolution, reduction of matrix effects, and increase of sensitivity. The LC and UPLC have performed with the use of mobile phase mixtures of water-organic solvent (acetonitrile or methanol) and infrequently acidified to enhance the ionization efficiencies of the compounds (Sin et al. 2014; Al-Hamdi et al. 2016).

Advanced Oxidation Processes for EDCs Degradation The AOPs has appeared as a promising method in wastewater treatment due to the fact that they offered different possible processes for strong oxidizing hydroxyl (•OH) radicals generation to react with organic pollutants and caused the ring cleavage. Unlike many other radicals, •OH radicals are nonselective and therefore readily to transform a large variety of organic pollutants into nontoxic molecules to eliminate the environmental pollution (Peller et al. 2001; Patil et al. 2010; Sin et al. 2013; Liu et al. 2016; Shen et al. 2017). With adequate reaction time and optimum operating experimental conditions, the AOPs can mineralize the hazardous organic pollutants to CO2, which was the most stable final product for chemical oxidation. The common AOPs developed for destroying EDCs are photolysis, photocatalysis, sonolysis, and Fenton oxidation. Below is an overview on different AOPs for EDCs degradation with respect to the use of LC-MS for intermediate products analysis.

Photolysis for EDCs Degradation Photolytical process is dependent on the ability of the pollutant to absorb the emitted radiation from artificial or natural light. In this method, the EDCs can react photochemically by reaching an excited state via the direct absorption of radiant energy or by •OH radicals generated from dissociation of water molecules (Vallejo et al. 2015). Weidauer et al. (2016) investigated the photodegradation of benzotriazoles (BTs) under sunlight irradiation (290–800 nm) at neutral pH in aqueous solution for 24 h. With the identification of degradation intermediates via LC-QTOF-MS, a proposed pathway of the photolysis of BTs is shown in Fig. 1. The results in their investigation showed that the aniline (TP 1) and aminophenol (TP 2) were the initial intermediate

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Fig. 1 Sunlight photolysis degradation of benzotriazoles (Adapted from Weidauer et al. 2016)

products based on the photolytic elimination of molecular N2. A slight amount of NH4+ was also reported to form due to the mineralization of BTs. An acetate conjugate (TP 4) was also formed after the denitrogenation by a reaction with the acetate buffer. Finally, the aminophenol isomers were reported to convert into dihydroxyphenazine (TP 5), 1,6-phenazinediol (TP 6) and aminoquinone (TP 3). The formation of intermediate products was also found from the direct photolysis of bisphenol A (BPA) under a UV-C lamp (Kondrakov et al. 2014). Their study demonstrated two catechol derivatives (BPA catechol and 4-(2-hydroxypropan-2-yl)catechol were detected using LC-MS-TOF approach. They went further to describe the detected catechol derivatives have weaker estrogenic activity compared to BPA. Souissi et al. (2012) identified the degradation intermediates generated upon photolysis of estrone (E1) under simulated UV irradiation. Nine main degradation products for E1 were observed by LC-Q-TOF, which revealed one to three additional hydroxylation preferentially located on the aromatic ring of E1. In addition, they also observed that the phenolic structure of the photolysis by-products still maintained, which needed a further investigation to evaluate the estrogenic risk in the environment. Jiao et al. (2008) investigated the photolytic degradation of tetracycline (TC) under a 500 W medium mercury lamp (λ = 365 nm, light intensity = 0.53 mW/cm2). The degradation of TC improved at low initial TC concentration and high solution pH. The intermediate products from TC photolysis were identified using LC-ESI (electrospray ionization)-MS and a degradation pathway was proposed (Fig. 2). The degradation mechanism of TC was explained via the loss of N-methyl, amino, and hydroxyl groups. Only 15% decrease in the total organic carbon (TOC) of the degraded solution was also reported, revealing a majority of TC

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Fig. 2 Photolysis of tetracycline under UV-A irradiation (Adapted from Jiao et al. 2008)

converted into intermediate products without complete mineralization. More recently, the degradation of microcystin-LR (MC-LR) under UV-B photolysis has also been examined (Moon et al. 2017). Their results showed that high UVB light intensity benefited the degradation of MC-LR. The degradation rate constants were found to be 0.0020, 0.0043, and 0.0061 min1 for 0.47, 0.58, and 1.57 mW/cm2. Based on LC-MS/MS analysis, the degradation mechanism of MC-LR was determined and proceeded via two processes: (1) bond cleavage and (2) intramolecular electron arrangement by electron pair in the nitrogen atom.

Photocatalysis for EDCs Degradation The photocatalysis process for the degradation of organic pollutants involved the use of semiconducting materials as photocatalysts to generate active species such as positive hole (h+), superoxide anion (O2•–) and •OH radicals. The active species were produced based on the generation of electron (e) in the conduction band and h+ in the valence band when the photocatalysts subjected to light irradiation. The e can then reduce the adsorbed O2 to O2•– radicals, whereas the h+ can oxidize either the organic pollutants directly or adsorbed H2O molecules to •OH radicals and hydrogen cation (H+) (Weber et al. 2012; Sin et al. 2014; Chen et al. 2015; Lam et al. 2016). Therefore, enhancing the e–h+ separation is an important consideration in photocatalysis to generate large amount of active species for organic pollutants degradation. Norfloxacin (NOR) was subjected to BiOBr/Fe2O3 photocatalysis and the degradation by-products were analyzed using LC-(+ESI)-MS (Guo et al. 2017). It

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was postulated that the overall photocatalytic degradation process of the organic compounds was most likely to have been initiated by •OH radicals. Their studies proposed several pathways of the NOR degradation and indicated eleven intermediate compounds produced during the course of reaction (Fig. 3). Three different tentative pathways were postulated for the degradation of the NOR (piperazine ring transformation, decarboxylation, and defluorination). The report also showed that the NOR degradation started by cleavage of the piperazine ring. The primary reaction intermediates were further decarbonylated to give some of the amide group-bearing phenyl ring products with the loss of CO2 gas. Further defluorination of the compounds was the possible reason that led to simpler molecules with m/z 278 and 192. López-Muñoz et al. (2013) carried out the degradation of aqueous isoproturon using the TiO2 (Degussa P25) as photocatalysts. Structural information of the main products formed during the photocatalytic reaction has been identified using LC-ESI-TOF-MS. A detailed mechanism for the different by-products formation has been suggested (Fig. 4). The processes partook are: (1) •OH radicals attacked on the isopropyl chain, dimethylamine group, and the aromatic ring, generating monohydroxylation (m/z 223), di-hydroxylation (m/z 239), and their oxidized products (m/z 221 and m/z 237), (2) N-demethylation and/or demethylation on the alkyl chain (m/z 193, 191, 165, 163, 151, 149, and 94), and (3) substitution or oxidation of NH2 groups linked to the aromatic ring (m/z 121, 95, and 124). The chromatography-mass spectrometry was successfully employed in their study to identify intermediates that could not be detected by other techniques. The photocatalytic degradation and biotoxicity reduction of tetracycline (TC) was investigated in the presence of TiO2-supported on MCM-41 as a catalyst (Zhou et al. 2017). The intermediates and the final products of degradation were tentatively identified by the LC-(–ESI)-MS/MS technique. The results of LC-MS/MS demonstrated 10 major intermediates, which were in a tendency to increase first and subsequently declined with the degradation of the TC molecules. Several mass spectra of by-products with their m/z values were reported in their paper; however, no reaction mechanism towards the formation of various by-products was studied. Indomethacin (IDM) was subjected to visible light in the presence of N-doped carbon dots/g-C3N4 photocatalysts, and the degradation products were tested using LC-MS/MS (Wang et al. 2017). It was suggested that the overall photodegradation of the IDM was initiated by •OH radicals. LC-MS/MS studies proposed possible pathways of IDM degradation, and inferred numerous of by-products were produced during the photocatalytic reaction (Fig. 5). Four different mechanistic pathways were suggested for the degradation of the IDM (cleavage of the amide bond, decarboxylation of the acetic chain, addition reaction of the C2–C3 double bond, and hydroxylation reaction of the chloro benzene ring). Furthermore, frontier electron densities (FEDs) data of the IDM molecules were calculated to predict the reaction sites for the radicals attack. The C19, C17, and C20 positions indicated higher FED2HOMO + FED2LUMO values, which demonstrated the high possibility of •OH radicals substituted reactions took place at the chloro aromatic rings. The C17, C12, C5, C22, C16, and C2 atoms of IDM showed more positive point charge than others,

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Fig. 3 Photocatalytic degradation of norfloxacin using BiOBr/Fe2O3 under visible light irradiation (Adapted from Guo et al. 2017)

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Fig. 4 Photocatalytic degradation of isoproturon over TiO2-Degussa P25 under a 150 W medium pressure mercury lamp (Adapted from López-Muñoz et al. 2013)

Application of Liquid Chromatography-Mass Spectrometry for the. . .

Fig. 5 Visible light photocatalytic degradation of indomethacin in the presence of N-doped carbon dots/g-C3N4 (Adapted from Wang et al. 2017)

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which implied that C17, C12, C5, C22, C16, and C2 were attacked by O2• via nucleophilic addition reactions. The prediction of potential attack positions by reactive species based on theoretical calculations was in accordance with their LC-MS/MS findings. Mena et al. (2017) investigated the visible light photocatalytic ozonation to degrade N,N-diethyl-meta-toluamide (DEET) using WO3 catalysts. The molecular structures of the by-products were deduced by analyzing the samples with LC-(-ESI)-QTOF-MS analysis. The parent molecule DEET showed a clear mass signal corresponding to m/z 206. The degradation mechanism encountered steps of mono- and poly-hydroxylation and/or oxidation, de-alkylation, and subsequently rupturing of the benzene ring to lead the production of short-chain aliphatic organic acids and eventually mineralized to CO2. Their mineralization efficiencies were further confirmed by total organic carbon (TOC) removal. Salma et al. (2016) reported an investigation on the identification of the degradation intermediates generated by UV-C photolytic and TiO2 photodegradation in aqueous ciprofloxacin (CIP) at different pH values (pH 3, 5, 7 and 9). An UPLC-MS/MS method was used in their study together with high-resolution QTOF-MSA (Duo-Spray Ion Source working in negative ion mode). The results showed that there were (1) fluorine atom photo-substitution by an OH group at neutral and moderately basic conditions, (2) defluorination at neutral conditions, and (3) fluorine conservation at strongly acidic medium. They also added that high photon energy produced by UV-C irradiation significantly improved the photodegradation of CIP in the presence of TiO2. Investigation on the photodegradation mechanism of phenol and the efficiency of photocatalysis by commercial TiO2 has also been carried out (Dang et al. 2016). A complete degradation of phenol aqueous solution was observed after 24 h of UV-C irradiation. The intermediate products from the phenol degradation were identified using LC-MS. Two different mechanistic pathways were proposed for the phenol degradation (Fig. 6). Their first degradation pathway suggested that phenol was degraded via hydroxylation reaction to form catechol, benzoquinone, hydroxyhydroquinone, and hydroxybenzoquinone. With a further attack of the •OH radicals, an oxidative aromatic ring-opening reaction occurred and rendered the formation of simple hydrocarbons, followed by oxidation to CO2 and H2O. They also suggested that 2-phenoxylcyclohexa-2,5-dienone, [1,10 -biphenyl]-4-ol, and tectoquinone were formed by combination of two phenoxide ions in the second degradation pathway.

Sonolysis for EDCs Degradation During the sonochemical process, •OH radicals were generated over acoustic cavitation that induced the hemolytic scission of H2O molecules. The acoustic cavitation was comprised of the formation, growth, and collapse of microbubbles in water. The microbubbles formed via the acoustic cavitation can tend to increase and decrease its size continuously until the resonance size reached prior to its violent collapse. The collapse of these microbubbles can generate extremely high

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Fig. 6 Photocatalytic degradation of phenol in the presence of commercial TiO2 under UV-C light irradiation (Adapted from Dang et al. 2016)

temperature (> 5000  C) and pressure (> 1000 atm) inside the bubbles. Therefore, the EDC molecules can react with •OH radicals or be degraded by pyrolysis (Torres et al. 2007; Chowdhury and Viraraghavan 2009; Ayoub et al. 2010). Papadopoulos et al. (2016) identified the intermediates and degradation pathway of ethylparaben (EP) through LC-TOF-MS analysis. A proposed sonochemical degradation pathway of EP is shown in Fig. 7. The LC-TOF-MS analysis of EP at 8 h reaction time showed the formation of four degradation intermediates methylparaben (MP), 3,4-dihydroxybenzoic acid (3,4-DHB), 2,4-dihydroxybenzoic acid (2,4-DHB), and 4-hydroxybenzoic acid (4-HB) appearing at 6.9 min, 1.4 min, 1.6 min, and 2.5 min, respectively. All the above aromatic intermediates were proposed to further oxidize through ring rupturing reactions into aliphatic intermediates via the continuous attack of •OH and SO4•– radicals. In another study of the paraben, Sasi et al. (2015) studied the •OH radicalsmediated sonolytic degradation of methylparaben (MPB). The detected degradation intermediates from the LC-QTOF-MS analysis showed that the MPB degradation occurred via three major steps: (1) aromatic hydroxylation, (2) hydroxylation at the ester chain, and (3) hydrolysis reaction. It was also reported from this study that the aromatic intermediates were further oxidized through ring cleavage reactions into lower aliphatic acids, which finally decomposed into CO2 and H2O as evidenced by their chemical oxygen demand measurement.

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Fig. 7 Sonochemical degradation of ethylparaben (Adapted from Papadopoulos et al. 2016)

Sonolysis of persistent pharmaceutical carbamazepine (CBZ) was also performed, and LC-MS/MS was used in the structural characterization of the degradation intermediates (Tran et al. 2013). The analysis of degradation intermediates showed that the attack of •OH radicals on CBZ led to two different degradation pathways by the formation of 10,11-epoxy CBZ and 10,11-dihyroxy CBZ. In the first degradation pathway, 10,11-epoxy CBZ was further attacked by •OH radicals and then formed acridine via a deamination process. In the second pathway, 10,11-dihyroxy CBZ was also attacked by •OH radicals and produced anthranilic acid and salicylic acid. Then, the acidic intermediates were reported to convert into aniline and benzoic acid. It was proposed that these organic compounds were finally mineralized into CO2 and H2O. Other EDCs such as 17β-estradiol (Ifelebuegu et al. 2014), 17α-ethinylestradiol (Ifelebuegu et al. 2014), phenol (Entezari et al. 2003; Lesko et al. 2006), 2,4,6trichlorophenol (Park et al. 2011), pentachlorophenol (Park et al. 2011), tetracycline (Eslami et al. 2016), bisphenol A (Torres et al. 2007, 2008), alachlor (Bagal and Gogate 2012), and dichlorvos (Golash and Gogate 2012) have also been successfully degraded by sonochemical treatment. The results showed that this method can be served as an effective tertiary treatment option in wastewater applications.

Iron Catalyzed H2O2 Production for EDCs Degradation Among transition metals, iron has been known to be a very effective catalyst for •OH radicals generation via Fenton process. The generally accepted mechanism of the Fenton process suggested that the generation of •OH radicals is based on the electron transfer between H2O2 and a homogeneous catalyst (iron) (Eq. 1). The iron can subsequently be reacted with H2O2 to produce hydroperoxyl (HOO•) radicals (Eq. 2). The generated •OH radicals have higher oxidation potential than the

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HOO• radicals; thus, the degradation process was mainly due to the former (Wongniramaikul et al. 2007; Tay et al. 2011; Sirtori et al. 2014). H2 O2 þ Fe2þ ! Fe3þ þ • OH þ OH

(1)

H2 O2 þ Fe3þ ! Fe2þ þ HOO • þ Hþ

(2)

Marković et al. (2015) investigated the degradation of ibuprofen (IBP) in aqueous solution by Fe2+/H2O2 treatment. Under the experimental conditions (Fe2+, H2O2, and IBP concentrations = 25.2, 306, and 60 mg/L, respectively), a significant decrease of IBP concentration was observed within 1 min and reached 78% degradation efficiency. The LC-MS-TOF analysis of degradation products showed that the Fenton reaction gave four aromatic products (C9H10O2, C13H18O4, C10H12O2 and C13H18O3) and two aliphatic products (C4H6O4 and C7H12O2) due to the powerful oxidizing •OH radicals. The effect of Fenton reaction on the degradation of bisphenol A (BPA) in a novel electrical discharge plasma reactor was also reported (Dai et al. 2016). The structures of degradation intermediates and final products were identified using LC-QTOF-MS. Various degradation products such as bisphenol-o-quinone (BPA12QN), 5-hydroxybisphenol A (BPA2OH), 4-hydroxyacetophenone (BPAP2), 4-[2-(4-hydroxyphenyl)propan-2-yl]-2-nitrophenol (BPA2NO2), 4,40 -propane-2,2diylbis(2-nitrophenol) (BPA210NO2), 5-[2-(4-hydroxy-3-nitrophenyl)propan-2-yl]-3nitrobenzene-1,2-diol (BPA210NO26OH), 4-[2-(4-hydroxy-3-nitrophenyl)propan-2-yl] benzene-1,2-diol (BPA2NO210OH), and 5-[2-(4-hydroxy-3,5-dinitrophenyl)propan-2yl]benzene-1,2,3-triol (BPA26NO21014OH) were detected. Based on these degradation intermediates, a possible degradation pathway of BPA was proposed in Fig. 8. They concluded that the attack of •OH radicals on BPA hydroxyl group was the primary pathway for the by-products formation. They also added that all the identified degradation products have lower estrogenic activity than BPA and ultimately degraded into CO2 as well as H2O. Sirtori et al. (2014) identified the thiabendazole fungicide (TBZ) degradation products in water over the Fenton treatment by LC-QTOF-MS. In their investigation, twelve degradation products were identified and most of them can be eliminated after 15 min of reaction time. They suggested possible transformation pathway of TBZ degradation and revealed numerous hydroxylation reaction occurred in both benzimidazole and thiazole rings by the •OH radicals attack (Fig. 9). The effect of iron species has also been studied by analyzing the degradation of amoxicillin (AMX) aqueous solution under a solar simulator (Trovó et al. 2011). It was found that the degradation of AMX was favored over potassium ferrioxalate complex (FeOx) compared to FeSO4. A complete oxidation of the solution over FeOx was obtained after 5 min, while 15 min of reaction time was necessary in the presence of FeSO4. The intermediates generated during the treatment were also identified using LC-ESI-TOF-MS analysis, which allowed to suggest the degradation mechanism proceeded via the opening of the four-membered β-lactamic ring and further oxidized the methyl group to aldehyde and/or hydroxylation of benzoic

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Fig. 8 Degradation pathway of bisphenol A by the Fenton treatment (Adapted from Dai et al. 2016)

ring, producing intermediate products after bound cleavage between different atoms and further oxidized to low molecular weight aliphatic compounds. Using metoprolol (MET) as endocrine disrupting chemical, the efficiency of photo-Fenton treatment was studied under different irradiation sources: (1) UV-C (λ = 254 nm), (2) black blue lamps (λ = 365 nm), and (3) simulated sunlight (Romero et al. 2016). Their results showed that the photo-Fenton using black blue lamps gave the best degradation and mineralization efficiencies. Complete degradation of MET was obtained after 7 min of reaction time, and the corresponding total organic carbon (TOC) removal was reported to be 81.2% after 90 min of reaction time under the determined experimental conditions (Fe2+ = 10 mg/L and H2O2 = 150 mg/L). Using the intermediate products detected via the ESI-MS and a LC-MSD-TOF, their proposed degradation pathway of MET was presented where

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Fig. 9 Thiabendazole degradation pathway for the Fenton treatment (Adapted from Sirtori et al. 2014)

the essential degradation was possible owing to the attack of •OH radicals (Fig. 10). An overall toxicity reduction of treated MET solution was also observed in their photo-Fenton process.

Conclusion and Future Trends The increasing usage of EDCs worldwide has garnered great public concern because of the vulnerability of humans and wildlife to numerous initiations of hormone-like activities even in trace concentrations in surface waters. Researches have shown that AOPs are potential technologies that widely studied and applied for degradation of organic pollutants including EDCs. The most typically investigated AOP method for destructing EDCs in water was semiconductor photocatalysis compared to photolysis, sonolysis, and Fenton treatments. Some of research works were also found in particularly with hybrid processes of AOPs to obtain advantages of synergistic effects as viable solution to treat these organic compounds in aqueous solution. Although a wide range of AOP processes have been applied, •OH radical generation

Fig. 10 Degradation pathway of metoprolol by the Fenton treatment (Adapted from Romero et al. 2016)

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has been recognized as a main active species for EDCs degradation. At the same time, along with the EDCs degradation, it was crucial to develop a comprehensive database for a myriad of intermediates and by-products formed since they can possibly be more toxic and persistent than the parent compounds. LC-MS has shown attractiveness as a technique used for the determination of intermediate products of EDCs degradation owing to minimal sample preparation, ease of separation, relatively low operating cost, and compatibility of the LC-MS system with aqueous samples. In addition, the ionization techniques employed in the MS can deliver data related to the mass of the parent compound and intermediate products under analysis. Selected compounds can be further investigated using MS/MS fragmentation for structural indications. Despite substantial progress has been accomplished, there are several technical points that required to be further examined. A great number of the technical papers involved investigating the behavior of an individual component tested in much higher concentrations than identified in the aqueous environment, while the real EDCs occurred in mixtures of multicomponents rather than as an individual component. In addition to target chemical analysis, research works should take account to carry out the toxicity screening in extending the understanding of AOPs effectiveness for EDCs degradation as toxicity bioassays are able to test the toxicity of either target or unknown contaminants. For the toxicity of the degradation intermediates, the Microtox method using bacteria Vibrio fischeri as toxicity indicator and the inhibition of Escherichia coli respiration can be used. At this point, the significance of degradation intermediates in point of ecotoxicological should consider in future AOPs studies. As a final remark, it should also be noticed that recent development and validation of suitable analytical protocols seem to be key in reliably assuring the transformation products of EDCs that can deliver good support to the examination of the overall performance of the treatment processes.

Cross-References ▶ Advanced Treatment Technologies Acknowledgments This work was supported by the Universiti Tunku Abdul Rahman (UTARRF/ 2016–C2/S03 and UTARRF/2017-C1/L02) and Ministry of Higher Education of Malaysia (FRGS/ 1/2015/TK02/UTAR/02/2 and FRGS/1/2016/TK02/UTAR/02/1).

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Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . OMW Characterization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Physicochemical Characteristics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biological Characteristics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . OMW Toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . OMW Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Physicochemical Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biological Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Oxidation and Advanced Oxidation Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Combined Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . OMW Valorization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Land Application . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biogas Production . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Composting . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Extraction of Valuable Products . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Other Uses . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Situation in Morocco . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cross-References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

Olive oil production industry is an important traditional agro-industry in the Mediterranean area, with an annual production of three million tons, which accounted for 97% of the world production. Morocco is one of the Mediterranean countries concerned with the attractive developing production of olive oil, with an annual production capacity of 1.5 million tons of olives and the sixth largest

Y. Rharrabti (*) · M. EI Yamani Polydisciplinary Faculty of Taza, Taza, Morocco © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_91

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producer of olives after Spain, Italy, Greece, Turkey, and Tunisia. Despite the economic importance of olive mill industry, generation of huge quantities of olive mill wastewater (OMW) and its uncontrolled disposal create a substantial environmental problem. OMW are characterized by its high toxicity related to its low pH, high content of mineral salts, and high organic load. Several treatment methods have been proposed to search the best potential solutions, including physicochemical methods (e.g., evaporation, coagulation, flocculation, membrane processes), biological treatment (aerobic and anaerobic digestion), and oxidation processes (e.g., ozonation, wet air and Fenton oxidation), but the most common method applied has been the storage of OMW in lagoons, followed by evaporation during summer season. The complex composition of OMW and their poor biodegradability due mainly to the presence of phenolic compounds, lipids, and organic acids make simple treatment methods not sufficient to ensure their purification. Many biotechnological applications have been proposed for the valorization of these liquids that significantly reduce the environmental impact of olive mills (e.g., composting, biogas production, and recovery of valuable compounds). Keywords

OMW treatment · Thermal processes · Chemical processes · Membrane processes · Anaerobic treatment · Anaerobic treatment · Oxidation processes · Ozonation · Wet oxidation · Fenton oxidation · OMW valorization · Land application · Composting · Biogas production · Morocco

Introduction The olive oil industry is an agro-industrial sector of great economic importance. Mediterranean countries are the main producers with 97% of the total olive oil production. Worldwide olive oil production for the crop year 2015–2016 is assessed at 3,159,500 tons. 2,322,000 tons comes from European Union countries. The biggest olive oil-producing country is Spain (1,401,600 t), followed by Italy (474,600 t), Greece (320,000 t), Turkey (143,000 t), Tunisia (140,000 t), Morocco (130,000 t) then Portugal breaking the 100,000 t barrier with 109,100 t (Fig. 1) (IOOC 2016). Since ancient times, the oil was traditionally obtained by pressure. After the big increase in production, the extraction of olive oil was obtained by the continuous extraction system (Fig. 2) including a vertical and horizontal centrifugation which separates the olive mixture in three-phase: oil, pomace, and a black liquid effluent called olive mill wastewaters (OMW) or in two phases: oil and wet pomace. The two-phase system uses a small amount of cold water and therefore a lower dissolution of phenolic compounds remain in the oil, but this also make it bitter. Despite the economic importance of olive oil extraction industry especially in the Mediterranean basin, this area is affected by pollution coming from the great amounts generated by olive oil mills. This black liquid wastewater known as “OMW” (which comes from the olive fruit-vegetation water, the water used for

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Fig. 1 Olive oil producing countries (International Olive Oil Council 2016)

Fig. 2 Three-phase (a) and two-phase (b) centrifugation systems

washing and treatment and a portion of the olive pulp and residual oil) and its uncontrolled disposal create a substantial environmental problem. The volume of OMW varies from 40 to 60 L for pressing method, but it ranged from 80 to 100 L for triple phase centrifugation process per 100 kg of olives (Harwood 2000). Whereas, OMW released by dual phase decanter are of small amounts compared to the other systems, mainly due to the addition of very low quantities of water during the olives crushing.

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The studies conducted on the composition of OMW indicated that this effluent contains 83–92% of water as a major part and large amounts of organic molecules, particularly polyphenolic mixtures with different molecular weights and other organic molecules, including nitrogen compounds, sugars, organic acids, and pectins that increase their organic load (Amaral et al. 2008; Massadeh et al. 2008; Danellakis et al. 2011; Ntougias et al. 2013; Bouknana et al. 2014; Mseddi et al. 2015; Alaoui et al. 2016; El Yamani et al. 2017). The most common practice for the management of OMW includes the use of evaporation ponds and the subsequent discharge of solids in landfills and on soil. However, evaporation do not contribute to the reduction of OMW toxicity, which makes the problem is always persistent. Several other treatment options have been investigated to search the best potential solutions, and these can be divided into four general categories: physicochemical methods (e.g., coagulation-flocculation, adsorption, combustion, and different membranes processes), biological methods (e.g., aerobic and anaerobic treatment), and oxidation processes (e.g., ozonation, wet air, and Fenton oxidation). Combined processes haves also proposed to overcome the weaknesses of each method and to increase the processing efficiency (Adhoum and Monser 2004; Paraskeva et al. 2006; Kapellakis et al. 2008; Coskun et al. 2010; Sampaio et al. 2011; Di Lecce et al. 2014; Amor et al. 2015; Martins et al. 2015; Weber et al. 2015). The complex composition of OMW and their poor biodegradability due mainly to the presence of phenolic compounds, lipids, and organic acids make simple treatment methods not sufficient to ensure their purification. Moreover, treatment of OMW requires high capital and operating cost units with limited efficiency due to high polluting loads. Recently, several research carried out on OMW had focused on their valorization through numerous applications (e.g., composting, use as fertilizer, biogas production, recovery of valuable compounds) (Zenjari 2000; Capasso et al. 2002; Visioli et al. 2005; Sarris et al. 2013; Elkacmi et al. 2016; El-Abbassi et al. 2017). This study provides updated information on research works carried out on: (i) the composition of the OMW and their toxicity, (ii) the options proposed for their treatment, and (iii) the possible applications of valorization.

OMW Characterization Physicochemical Characteristics Several studies were conducted on OMW characterization. These effluents have a very complex and heterogeneous physicochemical composition, which varies depending on many factors such as the variety and maturity of the olives, period of production, climatic conditions, farming methods, region of origin, and especially the oil extraction technology (Ben Sassi et al. 2006; El Yamani et al. 2017), among other affecting factors such as type of olives, tillable soil, and use of pesticides or fertilizers.

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OMW are characterized generally by an intensive dark brown to black color, an acidic character, high pollution load due to their high organic matter content, which mainly consists of polysaccharides, sugars, polyphenols, polyalcohols, tannins, proteins, organic acids, and lipids. Moreover, OMW contain considerable amounts of suspended solids (Amaral et al. 2008; Massadeh et al. 2008; Danellakis et al. 2011; Ntougias et al. 2013; Bouknana et al. 2014; Mseddi et al. 2015; Alaoui et al. 2016; El Yamani et al. 2017). Besides its strong organic content, OMW contain significant quantities of mineral salts in which the most representative elements are potassium, phosphate, and sodium (Massadeh et al. 2008; Ntougias et al. 2013). Recent literature data concerning the average concentrations of key physicochemical constituents of OMW has been summarized in Table 1.

Biological Characteristics Several studies were also carried out on different samples of OMW in order to highlight the microbial diversity that could be developed in this type of effluent. OMW microbiota is under many affecting factors such as soil and fresh water environments, specific cultivar, cultivation, and harvesting practice (Kavroulakis et al. 2011; Tsiamis et al. 2012). Yeasts related to Geotrichum (G. candidum), Candida (C. membranifaciens, C. michaelii, C. inconspicua, and C. tropicalis), Pichia (P. fermentans and P. holstii), Rhodotorula (R. mucilaginosa), and Saccharomyces (S. cerevisiae) have been isolated from Italian OMW (Sinigaglia et al. 2010; Bleve et al. 2011). Pichia (P. guilliermondii–syn. Meyerozyma guilliermondii) and Candida (C. diddensiae and C. ernobii) spp. were also the main yeast biota in OMW from Moroccan olive mills (Ben Sassi et al. 2008). The fungal flora consists essentially of Aspergillus flavius, Aspergillus candidus, Penicillium negricans, and Alternaria sp. (Aissam 2003). Tsiamis et al. (2012) have reported that OMW bacterial diversity consisted of members of Firmicutes, Actinobacteria, Alphaproteobacteria, Betaproteobacteria, Gammaproteobacteria and Bacteroidetes. While, bacterial flora in OMW was dominated by fermentative members of Bacteria, such as lactic acid (Lactobacillus and Oenococcus spp.) and acetic acid (Acetobacter and Gluconacetobacter spp.) bacteria (Kavroulakis et al. 2011). Mouncif et al. (1993b) have reported the total absence of fecal bacteria in OMW. In contrast, the presence of this kind of bacteria was confirmed by Kavroulakis et al. (2011) and it was related to the family Prevotellaceae and the RuminococcusEubacterium Clostridium cluster.

OMW Toxicity OMW toxicity is mainly related to the phenolic compounds’ action, which are responsible for their black coloring and which have phytotoxic and antimicrobial properties (Capasso et al. 1992; Casa et al. 2003; Ouzounidou and Asfi 2012).

7.45–68.48

0.50–9.50

3.13–30.22 nd

nd nd 0.027–1.051 nd 2.47–62.30 0.022–0.302 0.011–0.162 0.169–2.210 nd nd nd nd

g O2/L

g/L g/L

g/L g/L g/L g/L g/L g/L g/L g/L g/L g/L g/L g/L

Chemical oxygen demand Biological oxygen demand (5 days) Total solids Total suspended solids Volatile solids Ash Total phenols Total sugars Oil and grease Kjeldahl nitogen Phosphorus Potassium Sodium Magnesium Calcium Chlorides

– mS/ cm g O2/L

PH Conductivity

5.10–5.80 nd

Units

Parameters

References Amaral et al. (2008)

nd nd 7.739–10.432 nd nd 0.398–1.036 0.158–0.403 2.053–5.492 0.130–0.384 0.038–0.63 0.276–0.757 0.486–1.111

nd 14.21–46.19

23.25–63.27

78.54–160.10

nd nd

Mssadeh et al. (2008)

Table 1 Physicochemical characteristics of OMW

62.18 nd 11.00 41.65 nd 0.71 0.328 0.730 0.152 0.114 0.284 nd

64.68 nd

nd

nd

Danellakis et al. (2011) 5.62 11.28

nd nd 0.41–1.30 0.03–1.00 0.47–3.00 nd nd nd nd

nd nd

41.90–54.76 nd

nd

nd

4.98–5.10 nd

Ntougias et al. (2013)

nd nd 0.20–1.80 3.52–10.48 0.80–27.40 nd nd nd nd nd nd 23.79–142.71

nd 19.00–27.40

8.50–25.00

52.00–120.00

4.50–5.32 13.00–42.00

Bouknana et al. (2014)

1.12 0.154 20.10 2.40 1.40 0.32 0.38 nd

nd nd 5.00

70.25 12.8

35.00

64.00

4.33 15.56

Mseddi et al. (2015)

0.6 nd nd nd nd nd nd nd

nd 4.96 2.10

23.9 14.8

45.00

64.00

4.5–5.0 14.00

Alaoui et al. (2016)

nd 1.31–1.78 nd nd nd nd nd nd nd nd 0.83–1.16

nd nd

nd

nd

El Yamani et al. (2017) 4.86–5.38 9.48–11.70

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Morever, DellaGreca et al. (2001) have proven the negative effects of these wastes on soil microbial populations, on aquatic ecosystems, and even in air quality. OMW phenolic fraction has been reported to exhibit antimicrobial activity against some strains, which is even greater than the respective activities induced by the individual phenolic compounds, indicating the synergistic action of OMW phenolics (Obied et al. 2007). Concerning plants, uncontrolled OMW application could inhibit germination of seeds (Casa et al. 2003; Mekki et al. 2006), plant growth, and photosynthetic pigments (Ouzounidou et al. 2008). Even though phenols are considered as the principal cause of OMW toxicity, long-chain fatty acids and volatile acids attributed to non-phenolic related toxicity (Ouzounidou et al. 2010). Capasso et al. (1992) demonstrated that OMW remained phytotoxic to vegetable marrow and tomato plants even after total extraction of the polyphenols. In addition, Hanif and El Hadrami (2009) have suggested that the low pH and the osmotic stress caused by the presence of high Na+ and Cl concentrations may play a role in OMW acute toxicity. While Karaouzas et al. (2011) have attributed the negative effect of OMW on the aquatic fauna of fluvial ecosystems to both high organic load and fecal contamination.

OMW Treatment Physicochemical Methods Thermal Processes Evaporation is the most widely used mean for OMW management. It allows a concentration of OMW achieved either by a manmade heat source or by a natural source of thermal energy (air, sun). The last way for OMW evaporation (natural evaporation) is the most practical way in the Mediterranean countries. This process consists of the storage of OMW in the large evaporation ponds of shallow depth where OMW remain there for several weeks or even months depending on climatic conditions. OMW is evaporated, and at the same time, a partial biological degradation of the organic matter takes place through a series of aerobic and anaerobic fermentation processes. This simple method avoids OMW discharging into sewers and rivers and is a relatively cheap solution. However, it requires larger area together with production of black foul smelling sludge difficult to remove and pollutant infiltration to ground water. Moreover, the remaining paste needs further treatment. To overcome natural evaporation problems, evaporator panels were used to facilitate OMW evaporation (Fiestas Ros de Ursinos and Borja 1992). This way has reduced the area of evaporation ponds and multiplied by 40–100 the amount of water per m3 occupied by the soil, but odor nuisances and high-energy cost are the main disadvantages. Others thermal treatment methods such as combustion and pyrolysis have been tested as a ways of OMW management and as means of recovering energy for co-fueling the olive oil extraction plant. Combustion and pyrolysis, both are

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destructive techniques with high-energy requirements and expensive equipment needed. Pretreatment of OMW and posttreatment of the emissions of toxic substances are also required.

Chemical Processes Several chemical treatment processes for OMW are found in the literatures. Coagulation/flocculation is one of the most effective and used methods and remains the least expensive compared to its performance. It consists of OMW treatment by the use of additional chemicals in order to destabilize the suspended and colloidal matter and form an insoluble solid that can be removed easily from the waste. Oil, suspended solids, COD, and BOD are decreased in this way. Several researchers tested different chemicals products; the most important are ferric and ferrous chloride, ferric sulfate, and aluminium sulfate. Mixture of these coagulants (flocculants) and acidifying of the waste with hydrochloric acid solution are also tested (Kestioglou et al. 2005). Lime treatment of OMW was also studied and corresponds to pH increase (at about 11–12) for optimal performance (Paraskeva et al. 2006). COD, phenols, and suspended solid removal efficiencies of 50–90% were obtained in chemical treatments. The major disadvantage of this process is that there is in fact a simple transfer of the pollution from the soluble state to the mud state; large quantities of sludge with high pollution load are produced leading to serious disposal problems. In addition, most organic compounds contained in OMW are difficult to precipitate. Recently, the chemical method that has attracted attention for OMW treatment is electrocoagulation. It is a technique for destabilizing suspended, emulsified, or dissolved pollutants in an electrocoagulation cell by introducing an electric current without adding a chemical coagulant. The coagulant is generated in the solution from the conductive metal plates, commonly known as “sacrificial electrodes” often made on aluminium or iron (Fig. 3) (Bani Salameh et al. 2015). Adhoum and Monser (2004) have found a reduction of 76% COD, 91% phenols, and 95% color after 25 min of treatment at 75 mA/cm2. Electrocoagulation process allowed removal of total solids and COD of about 82.5% and 47.5%, respectively at 45 mA/cm2 after 70 min by using coupled iron-aluminum electrodes (Bani Salameh et al. 2015). The main advantages of this process are its high effectiveness in removing contaminants, simplicity of equipment, and generation of a lower volume of sludge compared with the classic coagulation technology (Hanafi et al. 2009). Adsorption is another chemical process used to remove hazardous inorganics and organic compounds, especially phenols from OMW. It is a simple and relatively economical method widely used in the removal of pollutants. Granular activated carbon (GAC) is the most commonly used adsorbent for removing organic pollutants. Adsorption on GAC showed about 30% COD reduction and a requirement of 50 kg carbon m3 effluent (Kestioglou et al. 2005). Adsorption of the OMW onto activated clay reduced the COD by a further 71% and the phenol content by a maximum of 81% (Paraskeva and Diamadopoulos 2006). Al-Malah et al. (2000) have also used activated clay for OMW treatment, reduction of 81% phenols and 71% COD was obtained. Other inexpensive minerals (i.e., clays, zeolite, etc.) can be used for removal of color and phenol adsorption from OMW.

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Fig. 3 Experiment set up of electrocoagulation method: (1) power supply, (2) pump, (3) magnetic stirrer, (4) wastewater reservoir tank, (5) cathode, (6) anode, (7) wastewater reservoir tank (Bani Salameh et al. 2015)

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1 A

6

V

5

2

4

7

rpm 3

Membrane Processes Membrane processes is one of the most promising treatment processes for OMW. They have gained a main role to seek for a viable process to treat OMW streams due to their capability to eliminate almost all of the pollutants in the water without adding solvents. These processes are based on the use of filtration membranes that allow separation under the pressure gradient effect of dissolved substances according to particle size and electrical charge. Ultrafiltration is the widely considered membrane process, while microfiltration, nanofiltration, and reverse osmosis also have been investigated and suggested for OMW treatment both for organic matter reduction and for polyphenols recovery. In addition, a separation of fats that are rejected by the membrane from salts, sugars, and phenolic substances that pass to the permeate can also be performed, enabling the economic exploitation of these substances. Microfiltration and ultrafiltration are used mainly for primary treatment purposes while nanofiltration and reverse osmosis are used for final treatment (Coskun et al. 2010). Membranes fouling may occur very easily due to gelling substances contained in OMW. It is the major technical drawback for the implementation of membrane technologies, leading to a reduction of the membrane efficiency. In addition, membrane processes are not suitable for the treatment of strong OMW because of their limited efficiency and their high costs, which make their use especially recommended as pretreatment steps in processes that aim at the recovery of valuable,

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expensive components such as polyphenols and flavoring agents from OMW. For the reasons stated above, some measurements, pretreatment steps, and process combinations seem absolutely necessary before the use of these processes for OMW treatment, as reported by many studies. Coskun et al. (2010) investigated the OMW treatment (Turkey) by membranes techniques; OMW were previously centrifuged, then filtered via ultrafiltration membranes followed by nanofiltration, and finally reverse osmosis membranes. The maximum COD removal efficiencies obtained at 10 bars ranged from 59.4% to 79.2% for the nanofiltration membranes, whereas about 96.2% for the reverse osmosis membranes, respectively. Likewise, an integrated centrifugation-ultrafiltration process was proposed by Turano et al. (2002) for the OMW treatment in Italy, reductions of 90% COD and 80% of suspended solids concentration were achieved. Di Lecce et al. (2014) have studied the fractionation of OMW (Italy) using a two steps microfiltration and nanofiltration membrane process at pilot scale. Results revealed a rejection of the nanofiltration membrane towards COD, dry matter, phenolic compounds, and antioxidant activity greater than 97%, independent of the volume concentration factor. Paraskeva et al. (2007) have tested a combination of different membrane processes for the treatment and fractionation of OMW; ultrafiltration followed by nanofiltration and/or reverse osmosis. The recovery ratio was fixed between 80% and 90% of the initial OMW volume, at 15–35  C operating temperature and transmembrane pressure between 1.0 and 2.25 bar. Phenols present in the OMW were removed to an extent exceeding 95% of the initial value following the nanofiltration step. The concentrate obtained at this stage was very rich in phenols. Better efficiency of the OMW treatment was achieved applying reverse osmosis after ultrafiltration.

Biological Methods Anaerobic Treatment Anaerobic biological process consists of microbiological digestion of OMW in the absence of molecular oxygen driven mostly by bacteria that converts organic compounds into biogas (methane and carbon dioxide). It involves three major steps; hydrolysis of complex organic compounds to their monomers which are converted to organic acids during acidogenesis and methanogenesis, the most significant anaerobic stage that consist of conversion of the organic acids into biogas (methane and carbon dioxide) (Sabbah et al. 2004). The method is widely used and particularly advisable because of their advantages associated to the feasibility to treat wastewaters with high organic load, low energy requirements, low production of sludge, ability to restart easily after several months of shut down, and generation of energy in the form of biogas (Mantzavinos and Kalogerakis 2005). However, anaerobic processes are affected by temperature, retention time, pH, H2 partial pressure, and the chemical composition of the wastewater, therefore, the additions of both alkali substances to neutralize pH and

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substances that are sources of nitrogen such as urea or ammonia are necessary. Moreover, the presence of compounds, such as phenols and organic acids, toxic to methanogens in OMW appears to be a major limitation of the anaerobic digestion of OMW (Hamdi 1996). Two approaches are proposed to overcome this problem: (i) the dilution of OMW to reduce phenols and organic acids concentration, (ii) the removal of compounds that are toxic to methanogenic consortia, with a prior aerobic treatment. The latter appears to be the most suitable. Numerous studies using different anaerobic reactor types have been investigated. In a recent study, OMW were digested in its original composition (100% v/v) in an anaerobic hybrid. High concentrations (54–55 kg COD/m3), acid pH (5.0), and lack of alkalinity and nitrogen are some OMW adverse characteristics. Loads of 8 kg COD/m3/day provided 3.7–3.8 m3 biogas/m3/day (63–64% CH4) and 81–82% COD removal (Sampaio et al. 2011). Raposo et al. (2004) have reported that when using an anaerobic reactor with bentonite as support medium for hydraulic retention times of up to 25 days and an organic loading rates range of 0.86–5.38 kg COD/m3/day, COD reductions were up to 88.8%, and methane production was 0.31 m3 CH4/kg COD. Bertin et al. (2004) have tried to overcome the problem of inefficiency of the conventional contact bioreactors to remove OMW phenolic compounds by employing an anaerobic OMW-digesting microbial consortium passively immobilized in column reactors packed with granular activated carbon (GAC) or “Manville” silica beads (SB). Anaerobic reactors packed with GAC and with SB showed a marked improvement in the removal of organic matter compared with conventional anaerobic treatment. The GAC reactor achieved 78.4% COD depletion, 90% phenol reduction, and 0.08 m3 CH4/kgCOD yield, while the SB reactor achieved 48.3% COD reduction, 50.6% phenol reduction, and 0.18 m3 CH4/kg COD yield. Ubay and Ozturk (1997) were investigated the anaerobic treatability of OMW using a laboratory scale upflow anaerobic sludge blanket reactor (UASBR). CODs were varied from 15 to 22.6 g/L while retention times ranged from 0.83 to 2 days; soluble COD removal was around 70%. A methane conversion rate of 0.35 m3 per kg COD removed was achieved. Dilution, nutrient addition, and alkalinity adjustment were required.

Aerobic Treatment Aerobic treatment is the use of aerobic strains for the biodegradation of the organic content of the waste. It is a commonly used technology in wastewater treatment. For OMW, known for their high organic contain specifically phenolic compound, a prior dilution and an acclimatization period for the microorganisms are required for the method to be effective (Yesilada et al. 1998). Kapellakis et al. (2008), have also reported that the process can operate efficiently only at the low concentrations of the effluent; i.e., of the order of 1 g/L COD. A long hydraulic retention time and/or with high recycle ratios are recommended in case of high concentrations (Paraskeva and Diamadopoulos 2006). Therefore, aerobic processes are unsuitable for direct and efficient treatment of OMWW. They are often used as pretreatment to increase the efficiency of the anaerobic processes.

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Many microorganisms (fungi, yeast, bacteria) have been tested in aerobic treatment of OMW: Pleurotus ostreatus, Bacillus pumilus, Panus tigrinus, Yarrowia Lipolytica, Aspergillus niger, Azotobacter chroococcum, Azotobacter vinelandii, Aspergillus terreus, Phanerochaete chrysosporium et Pleurotus ostreatus, etc. (Hamdi et al. 1991a, b; Kissi et al. 2001; Fadil et al. 2003). The various studies carried out on the aerobic treatment of OMW have shown a considerable rates removal of COD and of phytotoxic compounds were achieved. Yesilada et al. (1998) have reported a removal of 70% COD and 93% phenols using Funalia troggi for OMW treatment, while Coriolus versicolor allowed a reduction of 63% COD and 90% of the phenols for an initial COD of 28.20 g O2/L. A reduction of 75–66% of total organic carbon and total phenols, respectively, and a discoloration of about 45% after 4 days of incubation, were observed using Leutinus edodes (Vinciguerra et al. 1995). Kissi et al. (2001) have tested OMW treatment by Phanerochaete chrysosporium; results were reduction rates of 73% COD and 83% polyphenols with 12 days of incubation for an initial COD of 50 g O2/L. A removal of 55%, 52.5%, and 62.8% of COD was obtained with aerobic treatment of OMW by Geotrichum sp., Aspergillus sp., and Candida tropicalis, respectively (Fadil et al. 2003). Tziotzios et al. (2007) have examined the capability of olive fruit bacteria to remove COD and phenolic compounds from OMW using flasks reactors at different dilutions (20%, 50%, and 100%). The maximum phenolic and dissolved COD removal reached up to 82–90% for the dilutions of 20%, 50%, and 100%, in 11, 23, and 30 days, respectively.

Oxidation and Advanced Oxidation Processes Ozonation Ozonation is a very interesting technology for industries as it can operates at ambient conditions. It has been successfully employed for the OMW treatment either alone or in conjunction with other process. The method can be seen as a part of advanced oxidation processes (AOPs) that could be used in great number of reactions with organic and inorganic compounds given that hydroxyl radical results from the decomposition of ozone, which is catalyzed by hydroxyl ion or initiated by the presence of traces of other substances (Bani Salameh 2015). In fact, Ozone is considered as a powerful oxidizing agent and effective disinfectant, able to degrade bio-refractory organic matter by attacking selectively the double bonds of unsaturated fatty acids and phenolic compounds in OMW, leading to low-molecular-weight molecules more amenable for other treatments especially those biological (Mantzavinos and Psillakis 2004). Ozone production requires a large amount of electrical energy and its use can be uneconomical in wastewaters with high organic load. Therefore, this treatment can be enhanced by the addition of hydrogen peroxide and/or UV radiation, and even with catalysts or photo-Fenton reaction. Several studies were investigated in order to evaluate ozone’s efficiency in treating OMW and especially removing toxic compounds. Siorou et al. (2015) have reported that ozonated OMW held for 0, 60, 120, 300, 420, 540 min in

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a glass bubble reactor, showed a drastic reduction of OMW total phenols (almost 50%) after 300 min of ozonation with a concomitant decrease of OMW toxicity. Abatement of 80% of phenolic content and removal of 12% COD at the steady state were reached by ozonation treatment of OMW in a continuous reactor at pH 9, 1 ml/min of liquid flow rate and 15 g O3/m3. An integration schemes were also tested, ultrafiltration followed by ozonation was able to reach 93 and 20% of total phenols and COD depletion, respectively (Martins et al. 2015). Tsintavi et al. (2013) have performed a batch ozonation experiments on OMW in a glass bubble reactor. The results were a 57–76% reduction in total phenols and a 5–18% decrease in total carbohydrates contained in OMW. The higher removal efficiency (91% for total phenols and 19% for COD) was showing with ozone oxidation at initial pH value of 12, in a study carried out by Bettazzi et al. (2007).

Wet Air Oxidation Wet air oxidation (WAO) consists of a direct and total oxidation of organic compounds by molecular oxygen in a liquid aqueous phase, under relatively elevated temperatures and pressures (220–320  C, 50–200 bars). This technique was successfully applied to the treatment of different kinds of wastewaters (García et al. 1990; Verenich et al. 2004). Nevertheless, it implies severe conditions for operating (WAO units with high nickel content alloys, much more expensive than common stainless steel, and large thickness, in order to support the high pressures involved) that involved high capital costs and safety issues related to these conditions. Therefore, various catalysts were introduced to mitigate the severe reaction conditions to more amenable values (125–220  C, 5–50 bars) (Gomes et al. 2007). The WAO process has been successfully performed in the OMW treatment, and it has been demonstrated to be a feasible to effectively reduce the levels of contamination of these effluents. In fact, the oxidation reactions turn phenolic compounds into less toxic end products, such as carboxylic acids, carbon dioxide, or other harmless small-molecular-weight products (Weber et al. 2015). García et al. (1990) have achieved reductions of the total phenolic content of OMW close to 99% using pure oxygen (partial pressure 35 bars) at temperatures around 250  C. The WAO process catalyzed by hydrogen peroxide coupled with bioxidation for OMW was investigated by Chakchouk et al. (1994). Oxidation was conducted between 180  C and 200  C. WAO treatment allows a complete decoloration of the solution, a 77% COD reduction; the remaining COD is mainly low–molecular-weight carboxylic acids (chiefly acetic) easy for a biological post-treatment. In a comparative study, the oxidation using air as the oxygen source of OMW diluted with synthetic urban wastewater (1:10) has been carried out in the liquid phase at high temperatures and pressures. A positive effect has shown of the previous neutralization of the wastewater if compared to the oxidation conducted at the original pH of the effluent (pH = 5.3). Removal of toxic phenolic-type compounds is accomplished under relatively mild conditions of temperature and pressure (453 K and 7.0 MPa total pressure). In terms of DCO depletion and final biodegradability characteristics of the effluent, the use of free radical promoters, for instance, hydrogen peroxide, resulted in a significant enhancement of the process.

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While the use of copper oxide or platinum supported catalysts has showed not only an improvement in the COD removal rate but also a high degree of the mineralization of the wastewater contaminant load (Rivas et al. 2001a). Recently, Gomes et al. (2007) have studied the suitability of catalytic wet air oxidation for the OMW treatment, in a high-pressure reactor under an oxygen partial pressure of 6.9 bars. The complete total organic carbon and color removal was obtained using carbon supported platinum (1 wt.% Pt) after 8 h of reaction at 200  C.

Fenton Advanced Oxidation Fenton process is an advanced oxidation technology based on hydroxyl radicals generation by the decomposition of hydrogen peroxide when reacting with iron ions (Fe2+ and/or Fe3+) acting as homogenous catalyst at acidic pH (2–4) and ambient conditions. The iron (III) ions generated during the oxidation stage promote the removal of other pollutants by coagulation and sedimentation (Bautista et al. 2008). The Fenton process has variants, such as Fenton-like, photo-Fenton, and electroFenton processes for enhancement. Generated Fe3+ can be also reduced by reaction with exceeding H2O2 to form again ferrous ion and more radicals. This second process is called Fenton-like; it is slower than Fenton reaction and allows Fe2+ regeneration in an effective cyclic mechanism. Photo-Fenton’s process is an improvement of the classical Fenton’s reagent through the addition of ultraviolet radiation or visible light. When, in electro-Fenton process, pollutants are destroyed by oxidation at the anode surface and/or using the Fenton’s reagent in the bulk (Amor et al. 2015; Hadjltaief et al. 2015; Madani et al. 2015; Pariente et al. 2015). The main advantage of Fenton process is that the reagent components are safe to handle and environmentally benign. The process presents some other advantages: the reaction takes place at atmospheric pressure and at ambient temperature, hence no energy is required to activate the hydrogen peroxide; the cost-saving due to the use of metal iron compared to iron salts, the faster recycling of ferric iron at the iron surface and the production of harmful byproducts, is too low compared with other advanced oxidation processes. Nevertheless, the generation of large amount of iron sludge created at the end of the reaction that needs further treatment and the strict control of the pH (2.5–3.5) to guarantee a high catalytic performance are the major weaknesses of this process (Bautista et al. 2008). Fenton oxidation processes are very effective techniques in the removal of various organic pollutants from wastewater such as OMW and can be used as an effective pretreatment step to reduce toxicity. They have been extensively studied for the OMW treatment. Amor et al. (2015) were investigated the pretreatment of OMW with a classical Fenton method. They have been able to achieve reaching reductions of 17.6 and 82.5% of COD and total polyphenols, respectively, at pH 3.5 after 8 h reaction, producing an effluent suitable for anaerobic treatment. In another study done by Madani et al. (2015), the pH, iron-salt, and hydrogen peroxide dosage were also found to significantly affect the efficiencies of Fenton process, and acidic pH conditions were the most effective. The process showed high efficiency of COD (83%), total phenols (98.6%), color (77%), and aromaticity (67%) removal from the OMW.

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Different results were found in similar studies; the removal of about 65% of COD was reached when the Fe(II) concentration is 0.5 mol/L and the reaction time is longer than 4 h (Rivas et al. 2001b). Complete removal of phenols and COD reductions in the range 40–60% with a 2 h reaction time at a dose of 2–3 g/L ferrous sulfate and 3 mL of peroxide (60% w/w) (Vlyssides et al. 2004). The removal of 40% COD after 2 h of treatment remained unchanged thereafter, at lower Fe(II) and peroxide concentrations (0.03 and 0.25 mol/L respectively), COD removals were 40% after 2 h of treatment and remained unchanged thereafter (Ahmadi et al. 2005). Pariente et al. (2015) have tested an intensified Fenton catalytic process with a Santa Barbara Amorphous-15 silica-supported iron oxide as catalyst (Fig. 4), nearly 99% of phenol can be removed at 160  C in an acidic environment. In other researches, experiments with photo-Fenton and electro-Fenton processes were performed. The photo-Fenton reaction was less pH dependent than the Fenton process; the phenol removal rate was 60% at pH 2 and 70% at pH 5 (Mofrad et al. 2015). Hadjltaief et al. (2015) highlighted the role of UV wavelength and indicated that the use of both UV-C (λ = 254 nm) and UV-A (λ = 365 nm) can reach 100% phenol degradation, but UV-C was more efficient than UV-A. Electrochemical oxidation with a titanium-tantalum-platinum-iridium anode has also been used for OMW treatment. It was found that the process was able to remove almost entirely the content of phenols and the appearance of color, while, COD removals were up to 40% (Gotsi et al. 2005).

Combined Processes Combining ozonation and aerobic treatment, Benitez et al. (1999) reported a total COD reduction of 82.5%, a percentage higher than either of the two technologies

Fig. 4 Flowsheet of experimental installation for intensified-Fenton catalytic runs (Pariente et al. 2015)

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could achieve alone, indicating that ozonation increased the biodegradability of the OMW. When aerobic treatment preceded ozonation, COD reductions of up to 81% were reported. A combination of biological and UV/O3 process for the OMW treatment seems to be a serious alternative in the reduction of the COD. In particular, biodegradation of UV/O3 pretreated OMW found to have the highest removal levels; the percent of COD removal reaches about 91% (Lafi et al. 2009). Ultrafiltration followed by ozonation was able to reach 93% and 20% of total phenols and COD depletion, respectively. Moreover, this sequence led to an effluent with a BOD5/ COD ratio of about 0.55 which means that it likely can be posteriorly refined in a municipal wastewater treatment plant (Martins et al. 2015). In a previous study, Canepa et al. (1988) have combined ultrafiltration and reverse osmosis in a pilot plant scheme and observed 93–99% COD reduction. Khoufi et al. (2006) have investigated an electrochemical pretreatment step of OMW using the electro-Fenton reaction followed by an anaerobic bio-treatment. The electro-Fenton process removed 65.8% of the total polyphenolic compounds and subsequently decreased the OMW toxicity from 100% to 66.9%, which resulted in improving the performance of the anaerobic digestion. The anaerobic process applied as post-treatment reached a loading rate of 10 g COD/L/day without any apparent toxicity. Furthermore, in the combined process, a high overall reduction in COD, suspended solids, polyphenols, and lipid content was achieved by the two successive stages. Amor et al. (2015) in a study of OMW treatment applied Fenton’s reagent followed by anaerobic digestion. This combined process has presented a significant improvement on organic load removal, reaching COD degradations from 64% to 88%. In another research, adsorption, biological treatment, and photo-Fenton caused decreasing phenolic contents from OMW of 48.69%, 59.40%, and 95%, respectively. However, after the sequential treatment of the three methods was performed, higher reduction percentages in phenolic (total 99%) and organic contents (90%) were observed (Aytar et al. 2012). The different literature methods for OMW treatment are summarized in Fig. 5.

OMW Valorization Land Application OMW are rich in organic content and contain a significant amount of nutrients (sodium, phosphate and potassium) (Massadeh et al. 2008; Ntougias et al. 2013). OMW spreading on soil can enhance its fertility and promote microbiological activity (Sierra et al. 2007), especially for soil that have low levels of organic matter, microbial activity, biomass, and nutrient availability. Many researchers have applied OMW on soil. Positive effects were confirmed and were related to high nutrients concentration, especially K, and its potential for mobilizing soil ions (Paredes et al. 1999). Cox et al. (1997) suggested OMW use to attenuate leaching of toxic organic chemical such as herbicides (clopyralid and metamitron) in cultivated lands, because of its ability to reduce the mobility of certain organic

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Fig. 5 Treatment options for OMW

compounds in soil. Saadi et al. (2007) have reported that in addition to nutritional value of OMW, its potential herbicidal activity and ability to induce suppression of soil-borne plant pathogens are of extra value. Kotsou et al. (2004) have also confirmed the OMW suppressiveness against the plant pathogen Rhizoctonia solani. Whereas, OMW application on soil may lead to some negative effects, associated with its high mineral salt content, acidity, lipids accumulation, organic acids, and the presence of phytotoxic compounds, especially polyphenols (Cegarra et al. 1996; Paredes et al. 1999). This is the reason for a limitation of any effective use of OMW as a fertilizer, therefore, pretreatment before spreading is of great importance in order to avoid or reduce the negative effects on crops, soils, and the environment.

Biogas Production OMW are heavily polluted, because of their high organic load and phenols content, and hence, it can be used as a renewable energy source for biogas production through microbial treatment. Anaerobic digestion has been studied extensively by many researchers to produce energy (biogas) and to purify the OMWand thus re-use it especially in irrigation (Ergüder

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et al. 2000; Sarris et al. 2013). Although anaerobic digestion of OMW is feasible for the treatment, the presence of phenolic compounds known by their high antibacterial activity can slow down the process, hinder removal of part of the COD, and detract from its economic viability (Hernandez and Edyvean 2008). Therefore, pretreatment of OMW is a prerequisite and has been able to improve the performance and behavior of the anaerobic purification of this wastewater (Mantzavinos and Kalogerakis 2005). Ergüder et al. (2000) have reported that treatment of 1 L of OMW by anaerobic methods allowed production of 57.19 ( 1.5) L of methane gas. Mouncif et al. (1993a) also studied the production of biogas by anaerobic digestion of OMW, the rate of biogas produced was reached 207 L/kg of digested organic matter. A maximum ethanol concentration of 52 g/L was obtained in a batch reactor, using media enriched with glucose, with sugar concentrations up to 115 g/L (Sarris et al. 2013). Marques (2001), studied the anaerobic digestion of OMW (83%) with piggery effluent in an up-flow anaerobic filter. The production of biogas was 1–3 m3/m3/day (65–75% CH4). The co-digestion of various organic wastes has also been studied. Athansoulia et al. (2012) have investigate the use of OMW (30%) as a co-substrate with waste activated sludge (70%) to improve biogas production (methane). The biogas production rate reached 0.73, 0.63, 0.56 and 0.46 L biogas/L reactor/day for hydraulic retention times of 12.3, 14, 16.4, and 19.7 day. Dareioti et al. (2009) have examined the biodegradation of a mixture containing 55% OMW, 40% liquid cow manure, and 5% cheese whey in a two-stage continuous stirred tank reactor anaerobic process. The methane production rate at steady state conditions reached 1.35  0.11 l CH4/L reactor/day. Fezzani and Cheikh (2008) studied the mesophilic anaerobic co-digestion of OMW with olive mill solid waste in a batch digester and results showed an increase in biogas production and COD removal efficiency from 11.17  2.5 to 30.5  2.5 L/L digester and from 44.5  3 to 83.4  2%, respectively.

Composting Various disposal approaches have been proposed for appropriate OMW management, leading to a number of possible recycling and valorization methods (Federici et al. 2009). Composting is one of the main technologies of OMW recycling and it is a of particular interest for its operational simplicity and capacity to transform OMW into a high-quality amendment, rich in stabilized organic matter and nutrients for plants (Altieri and Esposito 2010). It has been considered an appropriate low-cost technology for organic waste recycling and organic fertilizer production (Ruggieri et al. 2009). Composting is an aerobic decomposition process that degrades organic matter over a period of weeks into a granular humus-like product which can be used as a fertilizer or soil conditioner (Tomati et al. 1996). It proceeds through three phases: an initial activation phase, followed by a thermophilic phase characterized by a rapid temperature increase, and a final mesophilic phase where the organic mixture cools down to air temperature (Ryckeboer et al. 2003).

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Piperidou et al. (2000) have developed a biofertilization system, based on the ability of Azotobacter vinelandii to grow on OMW and hence their conversion into an organic liquid fertilizer. Different substrates were tested with OMW to produce stable compost rich of nutrients. Tomati et al. (1995) found that a fertilizer with a high level of humification and no phytotoxic effects was obtained by composting OMW with wheat straw. Zenjari (2000) has studied a composting process by mixing OMW with cellulosic waste (straw), leading to a stable compost rich in humic substances, with a reduction of 50% and 95% of COD and total phenols, respectively. The nutritional value of OMW compost has been evaluated and confirmed by several studies. Cegarra et al. (1996) applied OMW composts to cultivate horticultural and other crops and found that yields obtained with compost fertilization were similar to, and sometimes higher than, those obtained with mineral fertilizers. Tomati et al. (1996) also observed an enhancement of activities in the plant-soil system after the addition of OMW composts.

Extraction of Valuable Products OMW, known by their highly polluting organic load, are also promising source of bioactive compounds and substances of high value and great interest. Therefore, the development of recovery processes for these valuable substances has great importance and are more studied. Phenolic compounds are one of the most valuable substances that OMW can provide. Excellent biological properties such as antioxidant, free radical scavenging, antimicrobial, and anticarcinogenic activities of the OMW biophenols are documented (Obied et al. 2005). Visioli et al. (2005) have indicated that phenolic compounds, as natural dietary antioxidants, may protect the organism against oxidative damage caused by oxidant agents (active oxygen, free radical, etc) that are involved in the etiology of chronic diseases such as cancer and atherosclerosis. Many recovery processes of these compounds have been studied: reverse osmosis, integrated centrifugation-ultrafiltration system, photochemical degradation of phenols, electrocoagulation, Bioreactor (Turano et al. 2002; Adhoum and Monser 2004). Another group of interesting compounds that could be obtained from OMW is soluble sugars, including the sugar alcohol or polyol called mannitol. Mannitol is used as an exicipient in pharmacy and as anticaking and free-flow agent, lubricant, stabiliser and thickener, and low-calorie sweetener in the food industry. Fernández-Bolaños et al. (2004, 2006) were able to recover mannitol with a high degree of purity. As another form of OMW valorization, Elkacmi et al. (2016) have used a fractional crystallization technique, on one hand, to extract oleic fatty acid as a product with a very important commercial value and, on the other hand, to produce soap and glycerin. Other compounds of great interest and benefits could be extracted from OMW: Pectins and oligosaccharides (Fernández-Bolaños et al. 2004), Polymerin (Capasso et al. 2002), etc.

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Other Uses Other ways of OMW valorization were investigated and proposed, which could be less common but also of high interest. Several studies were reviewed that OMW can be also used as a favorable medium for the enzymes production, which can found applications in the pharmaceutical, detergent dairy, and other industries. Lipases (Abrunhosa et al. 2013), laccases (Apohan and Yesilada 2011), Mn-dependent peroxidases, and pectinases (Crognale et al. 2006) seem to be the main enzymes obtained through a microbial (fungial) treatment. OMW represent a nutrient substrate for the production of single cell proteins thanks to their richness in organic matter. The production of single cell protein through aerobic yeast culture was studied by many authors because of the facility of their cultivation and the high protein content obtained (Gharsallah 1993). It has also demonstrated the ability of different strains to produce exopolysaccharides (Aguilera et al. 2008), rhamnolipids (Mercade et al. 1993) using OMW as sole nutrient and energy sources. Bioplastics (polyhydroxyalkanoates) have also been produced using OMW as a starting medium by using different Azotobacter strains (Cerrone et al. 2010). Recently, many researchers tried to take advantage from the phytotoxic and antimicrobial properties of OMW by using it in agriculture as biopesticide for crops protection (El-Abbassi et al. 2017). The different literature options of OMW valorization are summarized in Fig. 6.

Situation in Morocco Morocco is one of the Mediterranean countries concerned with the attractive developing production of olive oil. The olive oil sector contributes 5% to the national agricultural GDP. Covering an area of 922,000 hectares, the national farms have a

Fig. 6 Valorization options of OMW

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total production of around 1,500,000 tons of olives. The country also produces 130,000 tons of olive oil, and ranked the sixth largest producer of olives after Spain, Italy, Greece, Turkey, and Tunisia. In terms of exports, 17,000 tons of olive oil can be found on international markets (IOOC 2016). The main regions of production cover almost the entire country, except for the Atlantic coastal strip. The main variety produced is the “Moroccan Picholine” up to 96% of plantations. The olive oil processing sector is formed from both traditional, semi, and modern units. The traditional segment comprises 16,000 olive oil mills with the capacity to crush some 170,000 t of olives per year. Oil extraction rates are low, barely exceeding 14%. The modern and semi-modern segment is made up of 334 facilities equipped with an aggregate capacity of roughly 530,000 t of olives per year, and with substantial room for increase (IOOC 2012). Olive oil extraction using the traditional press and semi and modern methods (three phases and two-phase decanter) results in the production of large amount of highly contaminated OMW. The annual production of olive mill wastewater (OMW) in Morocco exceeds 250,000 m3 (Hanafi et al. 2010). Many researches are also carried out on Moroccan OMW. They confirmed the founding in studies performed in other countries. OMW is one of the strongest industrial effluents, characterized by a low pH, high organic load (explained by high values of DCO and DBO), great amount of total solids and a strong concentration of recalcitrant compounds such as phenols (the most toxic compounds), lignins and tannins which give it a characteristic dark colour (Ben Sassi et al. 2006; Bouknana et al. 2014; El Yamani et al. 2017). Concerning the microbial diversity in Moroccan OMW, Yeast and fungi populations appear to be high in OMW than bacteria. These microorganisms support the high salinity and acidic pH and are more resistant than bacteria to phenolic substances. The fungal flora consists essentially of Aspergillus flavius, Aspergillus candidus, Penicillium negricans, and Alternaria sp., while, no fecal bacteria was highlighted (Mouncif et al. 1993b; Aissam 2003). Thermal treatment, through evaporation in ponds (lagoons), represents the most practiced treatment of OMW in Morocco (Fig. 7), although several methods have been tested, but their application is limited. The options of valorization remain also

Fig. 7 Evaporation pond of OMW (Taza province, Morocco)

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at laboratory scale. Hanafi et al. (2009) have the electrocoagulation method for OMW treatment. A removal capacity of 80–85% COD, 75–80% phenolic compounds, 96–99% color, 7–8 kg/m3 suspended solids, and a reduction of more than 70% for orthophosphates, ammonium, zinc, and iron after 15 min of electrolysis at 250 A/m2 were reported. Electrocoagulation was also tested as pretreatment to enhance the fungal treatment. The electrocoagulation was able to decrease the COD and phenol content of the 25% OMW by more than 80% and the color by more than 90%. The strain Aspergillus niger van Tieghem was more efficient to reduce COD, phenol content, and color when the OMW was prior diluted or pretreated (Hanafi et al. 2010). Ben Sassi et al. (2010) have also studied the biological process using yeast isolated from OMW. Result was a reduction of about 50% of phenolic compounds. Achak et al. (2014) have used the absorption onto wheat bran for the removal of phenolic compounds from OMW in another study. The finding was a phenolic compounds adsorption rate of 67% at 50 g/L of absorbent dose with an increase of adsorption capacity at high alkanity. Valorization options of OMW were also studied by several authors. El-Abbasi et al. (2012) have revealed the considerable antioxidant capacity of the OMW, which can be considered as an inexpensive potential source of high added value powerful natural antioxidants comparable to some synthetic antioxidants commonly used in the food industry. El Yamani et al. (2017) also reported a high antibacterial activity of phenolic compounds extracted from OMW. Belaqziz et al. (2016) have investigated the effect of direct amendment of OMW on the fertility of soil, described as poor in the area of Marrakech (semi-arid region) in Morocco. Results of amendment with untreated OMW for two consecutive years (10 L/m2/ year) were increase of nutritive elements by 81% for nitrogen, 66% for phosphorus and 88% for potassium in soil. The accumulation of phenolic compounds and the increase of total peroxidase activity in plants have provided evidence of their protective role against the physiological stress induced by OMW. Recently, El-Abbassi et al. (2017) have reported that the OMW can be used to suppress the growth of the main bacterial, fungal phytopathogens, and weed species without any negative effects on crop growth. Nevertheless, some measures should be respected when using OMW as biopesticide especially in regards of dose and timing of use.

Conclusion Technologically, the problem of OMW could be more or less solved after many successful attempts at treatment and valorization. While realistically, it is still far from being solved, mainly because of practical and economic reasons related to nature of OMW, dispersion of the olive mills, and the seasonality of the process (Paraskeva and Diamadopoulos 2006). Hence, the olive oil industry needs even more studies on the development of olive oil extraction techniques or on the options for removal of OMW toxicity and their valuation.

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Cross-References ▶ Biofilm-based Systems for Industrial Wastewater Treatment ▶ Major Environmental Issues and Problems of South Asia, Particularly Bangladesh ▶ New Techniques for Treatment and Recovery of Valuable Products from Olive Mill Wastewater ▶ Technologies for Treatment of Colored Wastewater from Different Industries ▶ Wastewater Management to Environmental Materials Management

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Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Materials and Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Results and Discussion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Temperature . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . pH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Electrical Conductivity (EC) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Total Dissolved Solids (TDS) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Total Alkalinity (TA) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Dissolved Oxygen (DO) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Total Hardness (TH) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Anions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Study of Trace Elements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Iron . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Lead . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Manganese . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Arsenic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Zinc . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Mercury (Hg) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . An Overview of Activated Carbon . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Materials and Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Preparation of Activated Carbon . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Results and Discussion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cross-References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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D. Kibami (*) Department of Chemistry, Kohima science college (Autonomous) Jotsoma, Kohima, Nagaland, India © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_92

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Abstract

With an increasing level of pollution leading to quality deterioration, water quality assessment and remediation is becoming one of the thrust areas of research internationally. In acknowledgment of this, a preliminary investigation was done under the topic “Studies on Water Quality of Mokokchung District, Nagaland, India and Removal of Trace Elements Using Activated Carbon Prepared from Locally Available Bio-Waste” which has now come to form the subject matter of this project. Two aspects of study were taken up which include physicochemical analysis of water quality of Mokokchung district and also studies on the use of activated carbon for removal of trace elements. Accordingly, the studies on the quality of water of Mokokchung district have been presented as one full chapter. Again, the work on synthesis of activation carbon using two different adsorbents has been discussed. Studies of their application in removal of trace elements like Pb has been done using batch adsorption technique which includes effects of various experimental parameters such as initial concentration, contact time, and pH. Adsorption data fitted well with all the adsorption isotherm models. However, Freundlich isotherm displayed a better fitting model; this indicates the applicability of multilayer coverage of the Pb (II) on the surface of adsorbent. The adsorption kinetics was studied using four simplified models, and it was found to follow the pseudo-second-order kinetic model. The adsorption mechanism was found to be chemisorption and the rate-limiting step was mainly surface adsorption. Keywords

Pollution · Water quality · Mokokchung district · Activated carbon · Batch adsorption · Adsorption isotherm · Freundlich isotherm · Adsorption kinetics · Chemisorption

Introduction The quality of drinking water is a powerful environmental determinant of health (World Health Organization 2011). According to the World Health Organization, drinking water must be free of chemicals and microbial contaminations which are risk to human health. Good drinking water quality is essential for the wellbeing of all people (World Health Organization 2011; APHA 1985). It has been reported that increasing population and its necessities have led to the deterioration of surface and subsurface water (USEPA 1979; Rao and Mamatha 2004). It is therefore vital to regularly monitor the quality of water by studying the physicochemical properties of water which are important not only in the assessment of the degree of pollution but also in the choice of the best source and the treatment needed (World Health Organization 1984).

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In the state of Nagaland, the predominant sources of water are surface water from rivers, streams, ponds, and natural springs and subsurface water occurring as ground water. Information of water quality in Nagaland is very scanty, as there is very little documentation on the state of rivers/water bodies, and thus monitoring is a very recent phenomenon and so far has been taken up on a very limited basis (http:// npcbngl.nic.in/laboratory.htm). This study was considered important since Mokokchung is a fast developing town of Nagaland, where rapid urbanization is leading to rampant pollution of water sources, and hence, it was thought that a preliminary study of surface and ground water quality would be of value for developing management strategies for maintaining potable water quality. Mokokchung with Latitude of 26.12 N and 26.45 N and a Longitude of 94.18 E and 94.50 E is situated in the north-eastern region of Nagaland state. It is located 1325 meters above sea level and receives an annual rainfall of around 200 cm on an average (http://nagaland.nic.in). The present study was planned by selecting (one sample from each colony/ward) different ground and surface water sources around Mokokchung town (Fig. 1).

Fig. 1 Map of Mokokchung area

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Materials and Methods The water samples were collected from the various sources in separate container; prewashed (with HNO3 and thoroughly rinsed with deionized water) sterilized polyethylene bottles of 2 liters capacity were used for storing the water samples. Parameters like water temperature, pH, electrical conductivity, total dissolved solid (TDS), alkalinity, dissolved oxygen (DO) total hardness, calcium, magnesium chloride, sodium, potassium, phosphate, nitrate, and sulfate, fluoride were analyzed as per standard procedures (World Health Organization 1984). Each sample was analyzed in duplicate and the mean result reported. All The reagents used for the analysis were AR grade, and double distilled water was used for preparation of solutions (Table 1).

Results and Discussion The physicochemical parameters of the water samples collected from 12 different places. In and around Mokokchung town is describing as below (Tables 2 and 3).

Temperature Temperature is a significant factor which influences the abiotic and biotic components of the environment. It helps in controlling the solubility of gases. The Table 1 Determination of water quality parameters Sl.No. 1. 2. 3. 4. 5. 6. 7. 8. 9. 10. 11. 12. 13. 14. 15. 16.

Parameters Temperature Hydrogen ion concentration (pH) Electrical conductivity (EC) Total dissolved solids (TDS) Alkalinity (ALK) as CaCO3 Total hardness (TH)as CaCO3 Calcium (Ca) Magnesium (Mg) Sodium (Na) Potassium (K) Chloride (CI) Nitrate (NO3) Sulfate (SO4) Phosphate Fluoride Dissolved oxygen (DO)

Methods of determination Thermometer pH metry Conductometry Evaporation method Titrimetry EDTA -Titrimetry EDTA -Titrimetry EDTA –Titrimetry Flamephotometry Flamephotometry (argentometric method) Titrimetry Spectrophotometry Spectrophotometry Spectrophotometry Fluoridemeter Titrimetry

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Table 2 Results of physicochemical analysis of water Sl. No. 1. 2. 3. 4. 5. 6. 7. 8. 9. 10. 11. 12.

Water source Tap water Stream Bore well Stream Pond Pond Bore well Pond Bore well Stream Pond Bore well

pH 7.69

Electrical conductivity μScm1 322

TDS mg/L 287

Total alkalinity (TA)mg/L 111.5

Alisunkum 22.3 Tondentsunyoung 22.6

6.93 6.50

55.1 332

49.9 297

109 106

Alongmen Medical Majakong Salangtem

22.5 22.4 22.6 23.4

7.52 7.01 6.55 5.66

262 204 214 794

234 185 180 728

107.5 112.5 111 114.5

Dilong Aongza

24.6 23.7

5.05 6.35

547 390

495 351

115.5 109.5

Alempang Arkong Yimyu

24.1 24.3 22.3

5.51 5.10 7.16

484 375 215

530 342 189

117.5 117 113.5

Ward/colony FAC campus

Temperature  C 22.2

Table 3 Indian standard specifications for drinking water IS: 10,500 Sl. No. 1. 2. 3. 4. 5. 6. 7. 8. 9. 10. 11. 12. 13.

Parameters Temperature ( C) pH EC (μScm1) TDS (mg/L) Total hardness as CaCO3 (mg/L) Total alkalinity (mg/L) Calcium (mg/L) Magnesium (mg/L) DO (mg/L) Chloride \(mg/L) Sulfate (mg/L) Nitrate (mg/L) Fluoride (mg/L)

IS:10500,1991 Requirement (desirable limit) – 6.5–8.5 500 500 300

Permissible limit in absence of alternate source – No relaxation 1000 2000 600

200 75 30 3 250 200 45 1

600 200 100 10 1000 400 100 1.5

temperature of the water samples analyzed did not have much variation and was between 22.2  C and 24.6  C which indicated that the water samples were in the natural state with less contamination from the effluents (Patil and Patil 2010).

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pH pH is an important factor which expresses the acidic or alkaline nature of a solution. According to Fakayode, S.O. (Fakayode 2005), the pH of a water body is very important in determination of water quality since it affects other chemical reactions such as solubility and metal toxicity (Kataria 1995). The pH values vary from 5.10 to 7.62 where water samples collected from Salangtem, Dilong, Aongza, Alempang, and Arkong (sample no.7, 8, 9, 10, and 11) were found below the prescribed limit. This lower limit indicates the corrosive nature of the water samples collected which is not fit for drinking purpose and its prolong use may cause serious health problems (Golterman et al. 1978; Maiti 1982).

Electrical Conductivity (EC) Conductivity is the ability of water to carry an electrical current. It is a rapid method to measure the total dissolved ions and is directly related to the amount of total dissolved salts (USEPA 1979). An observation of the results reveals the nonuniformity in the variation of EC values in all samples. EC values in water samples ranged from 55.1 to 794 μ mhos/cm. Most of the water samples collected lies in the range of excellent and good category and can be used for both drinking as well for irrigation purpose (John 1997; Anil Kumar De 1989).

Total Dissolved Solids (TDS) Total dissolved solids indicate the nature of salinity in water (USEPA 1978). It also gives an idea about suitability of the water for various uses. Dissolved solids tend to increase with increasing pollution of water. TDS above 500 mg/L is not considered desirable for drinking. The TDS values for study area have varied from 49.9 to 728 mg/L. Most of the water samples lie below the permissible limit. High TDS values for Salangtem and Dilong (sample 7 & 10) could be due to presence of high content of Ca2+, CI-.\, and Na+ ions in the water (Anil Kumar De 1989; Keller and Pitblade 1986).

Total Alkalinity (TA) Alkalinity of water is a measure of its capacity to neutralize strong acids and is due to the presence of bicarbonate, carbonate, and hydroxide compound of calcium, sodium, and potassium (Gilbert Masters 2013). The observed values of alkalinity ranged between 106 and 117.5 mg/L in water. According to United States environmental protection agency (USEPA 1979), the maximum permissible limit of total alkalinity is 120 mg/L. Here the observed values are within the permissible limit.

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Dissolved Oxygen (DO) Dissolved oxygen refers to the amount of oxygen dissolved in the bodies of water. It is a relative measure of the amount of oxygen that is dissolved or carried in a given medium. Insufficient oxygen often caused by the decomposition of organic matter tends to suppress the presence of aerobic organism. The levels of dissolved oxygen varied from 6.90 to 15.20 mg/L, which is within the desirable limit as Prescribed by ISI (Indian Standard Institute 1993). The high value of DO for the water samples may be due to wave action, pollution load, organic matter, and photosynthetic activity (NEERI 1979; Suess 1982; Kannan et al. 1993).

Total Hardness (TH) Total hardness is determined by the concentration of multivalent cations (usually Mg2+ and Ca2+) in water (Somasekhara Rao and Someswara Rao 1995). It is usually expressed as the equivalent of CaCO3 concentration. Hardness is commonly understood as a property of water, which prevents the lather formation with soap. Total hardness for the samples under study was found to vary between soft and moderately soft category with their hardness values in the range of 55–150 mg/L (Table 4).

Cations The cations analyzed in the present study includes calcium, magnesium, sodium, and potassium. For the study area, the concentration of calcium has varied from 8.01 to 29.65 mg/L which is well below the permissible limit (IS 1050092 recommended its desirable upper limit at 75 mg/L), whereas the concentration of magnesium varies from 2.43 to 15.10 mg/L in the study area (upper limit IS10500-91 for Magnesium is 30 mg/L). For the study area, water is in the range of soft and moderately soft category which is due to the low concentration of calcium and magnesium ions present in the water. Potassium content in the water samples varied from 1.0 mg/L to 11.0 mg/L which is well below the recommended limit, whereas the concentration of sodium ranges from 3.0 to 53.0 mg/L in the study area. Sodium in drinking water normally presents no health risks, as about 99% of the daily salt intake is from food and only about one percent from water (U.S. Environmental Protection Agency 1991; Kumar and Sinha 2010; WHO regional office for Europe 1979).

Anions For the present study, the anions analyzed were chloride, sulfate, phosphate, nitrate, and fluoride. The concentration of chloride in all the samples in the study area ranges from 11.36 to 63.90 mg/L (permissible value for drinking water standard as per IS10500-91 is 250 mg/L), which indicated low level of pollution by sewage disposals

Sl.No. 1. 2. 3. 4. 5. 6. 7. 8. 9. 10. 11. 12.

Dissolved oxygen (DO) mg/L 7.00 7.80 9.00 13.00 6.90 12.00 12.50 14.30 15.20 8.60 14.50 12.35

Total hardness (TH) mg/L 125 100 145 130 55 60 85 95 125 150 70 60

Table 4 Results of physicochemical analysis water samples Ca mg/L 15.23 8.01 29.65 20.04 18.43 16.03 26.45 16.83 17.86 20.84 8.21 9.61

Mg mg/L 2.92 4.87 15.10 6.33 5.84 4.38 5.36 12.18 9.49 2.90 6.81 2.43

CI mg/L 14.21 21.31 42.63 35.52 32.66 11.36 63.90 61.06 14.20 56.80 51.12 39.76

NO3 mg/L 0.194 0.113 0.010 0.016 0.080 0.191 0.005 0.021 0.028 0.015 0.141 0.098 SO4 mg/L 7.83 7.79 14.01 17.46 16.60 9.86 12.76 14.45 16.54 9.67 8.47 12.67

K mg/L 11.00 1.00 1.00 5.00 3.00 3.00 3.00 5.00 2.00 3.00 2.00 4.00

Na mg/L 8.00 3.00 10.00 17.00 20.00 24.00 53.00 47.00 22.00 20.00 39.00 9.00

F mg/L 1.76 0.43 1.24 0.41 0.57 1.90 0.15 0.18 1.35 1.95 0.19 0.83

PO4 mg/L 0.75 0.41 0.30 0.80 0.70 0.84 0.42 0.55 0.40 0.65 0.59 0.45

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(the concentration of chloride in water serves as an indicator of pollution by sewage National Research Council 1972). Whereas the concentration of sulfate, phosphate, and nitrate were found to be 7.79 to 17.46 mg/L, 0.30 mg/L to 0.84 mg/L, and 0.005 to 0.194 mg/L, respectively. These values were well below the permissible limits, and their presence could only be from the natural sources (Bernard and Ayeni 2012; Delisle and Schmidt 1977). The concentration of fluoride in all the samples in the study area has varied from 0.15 to 1.95 mg/L. Fluoride content in FAC campus, Majakong, and Alempang were beyond the permissible limits, which may be due to weathering of minerals, rock dissolution where fluorine is leached out and dissolved in ground water (Galagon and Lamson 1953; USEPA 1980; Katoky et al. 2008; Gupta et al. 1986, 1999, 2011; Haloi and Sharma 2011; Sarala Kumar and Rao 1993; Somani et al. 1972).

Study of Trace Elements Trace elements, especially the so-called heavy metals, are a serious pollutant in our natural environment due to their toxicity, persistence, and bioaccumulation problems which is a threat to human life and the environment (Nassef et al. 2006). Trace element pollution is one of the major problems coming with increasing of human activities, especially near water resources and rivers (Fang and Hong 1999). Their presence in the wastewater of several industrial processes, such as electroplating, metal finishing, metallurgical work, tanning, chemical manufacturing, mining, and battery manufacturing, has brought about more environmental concerns due to their toxicity even at low concentrations. (Rajappa et al. 2010; Weaver et al. 1961; Fatoki and Hill 1994). Many studies have focused on the determination of trace pollutants particularly heavy metals like Cu, Zn, Pb, and Hg in different sites highly polluted by domestic, agricultural, and industrial wastes worldwide (ATSDR 2000; Griffin 1960; Trivedy and Goel 1986; The report of the scientific review committee Ottawa 1990). The objectives of the present investigation were to find out the status of the trace elements in water in the study area. For this study, the presence of different trace element in the water was planned to be studied. The following trace elements, i.e., iron, zinc, lead, manganese, arsenic, and mercury, were chosen. They are classified as highly toxic and moderately toxic or less toxic in nature. The results of the analysis of trace elements in the study area are given in Table 5. The significance of the presence of individual trace element is discussed below.

Iron Iron is one of the most abundant metals in the earth’s crust. It is an essential element in human nutrition. The common sources of iron in groundwater are weathering of iron bearing minerals and rocks, industrial effluent, acid-mine drainage, sewage, and landfill leachate. In drinking water, Fe (II) salts are unstable and are precipitated as insoluble Fe (III) hydroxide, which settles out as a rust-colored silt. Iron may also be present in drinking water as a result of the use of iron coagulants or the corrosion of steel and cast iron pipes during water distribution. Iron also promotes

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Table 5 Results of physicochemical analysis of water samples Sl.No. 1. 2. 3. 4. 5. 6. 7. 8. 9. 10. 11. 12.

Ward/colony FAC campus Alisunkum Tondentsunyoung Alongmen Medical Majakong Salangtem Dilong Aongza Alempang Arkong Yimyu

Fe mg/L 3.200 0.091 0.145 0.295 0.193 0.154 0.098 0.082 0.156 0.144 0.246 0.275

Pb mg/L 3.00 0.062 0.023 0.034 0.104 0.101 0.001 0.023 0.024 0.048 0.035 0.029

Mn mg/L 0.00 0.063 0.035 0.022 0.505 0.176 4.801 0.195 0.486 0.367 0.474 0.234

As mg/L 0.00 0.019 0.027 0.011 0.044 0.014 0.000 0.017 0.046 0.048 0.014 0.045

Zn mg/L 0.00 1.542 0.065 1.108 1.274 0.921 0.00 1.347 1.234 1.643 1.972 0.450

Hg mg/L 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00

undesirable bacterial growth (“iron bacteria”) within a waterworks and distribution system, resulting in the deposition of a slimy coating on the piping (Weaver et al. 1961; Florence 1977). Iron in water can cause yellow, red, or brown stains on laundry, dishes, and plumbing fixtures such as sinks. Normally it causes slight toxicity, but excessive intake can cause siderosis and damage to organs through excessive iron storage (Indian council of Medical Research 1975). It is observed that the concentration of iron has varied from 0.091 to 3.200 mg/L in the study area. The water sample from FAC campus (sample 1) showed a high Fe content, whereas the other water samples collected in the area under study had Fe concentration below the desirable limit. One of the reasons for high concentration of Fe lying in these water bodies might be due to the presence of iron-rich soil layers and sediments. Another reason could be due to the presence of wastes that get to the surface as effluent discharges (Rajappa et al. 2010).

Lead Lead occurs naturally in the environment. However, most lead concentration that is found in the environment is a result of human activities. The presence of lead is primarily from household plumbing systems containing lead in pipes, solder, fittings, or the service connections to homes. This is more likely to happen when the water is slightly acidic. Other sources of lead include organo-lead compounds, which is used extensively as antiknock and lubricating agents in petrol, although their use for these purposes in many countries is being phased out (Fatoki and Hill 1994). The World Health Organization has established 0.05 mg/L as a tentative limit for lead in drinking water. The present study has showed that the concentration of lead varied from 0.001 to 3.00 mg/L in water samples. The water samples collected from all the area under study, except in FAC campus (sample 1), Medical (sample 5), and Majakong (sample 6), generally have values below the permissible limit. The high concentration of lead in this water samples could be due to extensive use of lead piping in the household plumbing system (Scoullos et al. 2001).

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Manganese Manganese occurs naturally in many surface water and groundwater sources. In surface waters, manganese occurs in both dissolved and suspended forms, depending on factors like pH, anions present, and oxidation–reduction potential (Griffin 1960). The manganese ion imparts an undesirable taste to beverages and stains plumbing fixtures and laundry (Hanaa et al. 2000). The stains caused by manganese are harder than those of iron to remove. The maximum permissible limit of manganese in water is 0.5 mg/L. The careful observation of the results indicates that the distribution of manganese concentration is not uniform in the study area. Some of the water samples do not contain manganese, whereas in two samples (no.5 and 7) the concentration of manganese was found to be well beyond the permissible limit.

Arsenic Arsenic is a natural element which occurs in many environmental compartments such as rock, soil, water, and air as well as in plant and animal tissues. Whatever it is natural or industrial origin, it is often responsible for the contamination of water supplies. Depending on the physicochemical condition of the environment, some arsenic compounds can be highly soluble, resulting in a high level of bioavailability. The presence of these compounds has therefore been declared as a major risk to human health in various parts of the world. Drinking well water with low to moderately elevated levels of arsenic over a long period of time may lead to chronic health effects (Klavins et al. 2000; Hanaa et al. 2000). Chronic health effects, such as cancer, develop over a number of years and can be difficult to detect, especially in the early stages. Higher levels of arsenic can also lead to more immediate or acute health effects that usually have more noticeable symptoms. For arsenic in drinking water, a provisional guideline value of 0.05 mg/l is recommended (ISI-IS: 2296-1982). In the present study, the concentration of arsenic varied from 0.00 to 0.048 mg/L which is within the permissible limit. The result reveals that the distribution of arsenic concentration is not uniform in the study area.

Zinc Zinc is an essential trace element found in virtually all food and potable water in the form of salts or organic complexes. WHO has established 5 mg/L as the highest desirable level and 15 mg/L as the maximum permissible limit for zinc in water for domestic use. Zinc is required and is a beneficial element in human metabolism. A deficiency of zinc in the diet of the children leads to growth retardation (Majumdar 2003). Zinc may be toxic to aquatic organisms, but the degree of toxicity varies greatly depending on water quality characteristics as well as the species being considered. In the present study, the concentration of zinc has been observed to be varying from 0.00 to 1.97 mg/L in the study area. The results indicate that the concentration of zinc is below the maximum permissible limit in all the samples.

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Mercury (Hg) Mercury occurrence is generally industrial in origin. It is used in the electrolytic production of chlorine, in electrical appliances, in dental amalgams, and as a raw material for various mercury compounds. It is a very toxic element, especially when methylation of inorganic mercury occurs in fresh water and in seawater (Michael Scoullos et al. 2001), although almost all mercury in uncontaminated drinking water is thought to be in the form of Hg2+. Thus, it is unlikely that there is any direct risk of the intake of organic mercury compounds. However, there is a possibility that methyl mercury will be converted into inorganic mercury. Food is the main source of mercury in nonoccupationally exposed populations; the mean dietary intake of mercury in various countries ranges from 2 to 20 mg/day per person. In cases of mercury poisoning of any type, the kidney is the organ with the highest bioaccumulation of mercury. Mercury was not detected in any of the water samples in the area understudy.

Conclusion The analysis of the water quality parameters from different places around Mokokchung town indicates that electrical conductivity, TDS, total alkalinity, total hardness, calcium, magnesium, chloride, sodium, potassium, dissolved oxygen, and sulfate are well within the permissible limits. However, some of the water samples collected showed low pH value indicating the corrosive nature of this water samples, which may be due to the presence of toxic metals such as Pb and Cu. Low degree of effluents sources of nitrate and phosphate was observed as concentration of nitrate and phosphate in water was minimal. Fluoride content in water samples collected from FAC campus, Majakong, and Alempang was above the permissible limit. It was found that the concentrations of some of the elements were found to be above the permissible limits, whereas in some areas elements such as Zn, As, Hg, and Pb could not be detected as it was below the detection limit. The study is done for one set of samples. However, in order to understand the proper seasonal variation, a detailed study for 3 years might be required. The results reported here are an initial report of water quality of Mokokchung district (Kibami et al. 2013).

An Overview of Activated Carbon This section contains a brief account of activated carbons. Considering that the uniqueness of activated carbons is because of their role as adsorbents, this term as well as the term porosity have also been briefly explained. This section also contains a concise discussion on metal ions as water pollutants, with special emphasis on trace elements which are found in water, such as manganese and lead. The methodology that has been developed in this study for removal of Pb, based on biosorption has also been explained.

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Introduction The presence of lead in drinking water is hazardous to the environment and public health because of its toxic effects (Bahadir et al. 2007). The chief sources of lead in water are from the effluents of processing industries (Nadeem et al. 2005; Bhattacharjee et al. 2003; Sekar et al. 2004; Rivail Da Silva et al. 1996). Lead produces a variety of adverse effects on cellular function and primarily affects the nervous system, gastrointestinal tract, and hematopoietic system (Tong 1998). Chronic lead poisoning occurs when small amounts of lead are taken in over a longer period (Volesky 1990). According to the US Environmental Protection Agency (EPA), the permissible level for lead in drinking water is 0.05 mg L1 (Bhattacharjee et al. 2003) and that of Bureau of Indian Standards (BIS) is 0.1 mg/l (BIS 1981). Therefore, a very low concentration of lead in water is very toxic (Bhattacharjee et al. 2003). Removal of lead in water thus assumes a very important part in the process of water treatment. At present, various methods like chemical precipitation, electrochemical reduction, ion exchange, reverse osmosis, membrane separation, and adsorption have been developed to remove Pb2+ from wastewater (Connell et al. 2008; Acharya et al. 2009; Ricordel et al. 2001; Saeed et al. 2005; Doyurum and Celik 2006). Adsorption of activated carbon has been shown to be very effective for the removal of toxic metal ions from aqueous solutions. However the high cost of activated carbon limits its use especially in developing countries. This has led for the search of activated carbons from cheap and renewable sources by various researchers over the years (Ferro-Garcia et al. 1990; Sohail and Qadeer 1997; Netzer and Hughes 1984). In our study, activated carbon prepared from common bamboo and common buckwheat plant was used owing to its easy availability in the area. Moreover, it requires little processing and was thus inexpensive.

Materials and Methods Preparation of Activated Carbon Preparation of Activated Carbon by Indigenous Method For this study, the initial carbonization is done at bamboo mission Dimapur, Nagaland, India. In the bamboo mission, indigenous method of preparation of bamboo charcoal (taken from Bamusa vulgaris/BVC) is done for commercial purpose. In this process, the whole bamboo culms are cut into a uniform size of 1 m each so that it fits into the kiln and stacking is easier. The bamboo is placed through the door at the bottom and is stacked horizontally in the kiln (Fig. 2). The door is then closed with bricks and plastered with mud on the outside for better insulation and to prevent leakage. The feed is fired through the opening at the top of the kiln, and once the feed is ignited, the opening is closed. During the initial stages of firing, the openings in the wall of the kiln are kept open to create the required draft. Initially black smoke will be emitted from the opening at upper end after which it will change to dense

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Fig. 2 Bamboo charcoal kiln at bamboo mission Dimapur

white fumes. Once the black smoke changes, openings are closed one by one starting from the top to bottom. Carbonization is achieved by maintaining the temperature at 400–500  C by regulating the openings both in horizontal and vertical directions. All the openings are closed after 2 days so that air is not allowed to enter the kiln to prevent the charcoal from catching fire. Cooling is done for a day to reduce the temperature to 100  C. The biomass is removed from the opening at the bottom of the kiln and is washed, dried, and crushed and further subjected to chemical activation with 0.1 N HNO3 and 0.1 N H3PO4, and for which the carbons were washed with double-distilled water to remove the excess acid and dried at 150  C for 12 hours. All the activated carbons are chemically activated with 0.1 N solution HNO3 and 0.1 N H3PO4. The powdered activated carbon obtained after HNO3 and H3PO4 treatment has a particle size in the range of 40–50 μm mesh. Another form of activated carbon in powder form is prepared by the pyrolysis of Fagopyrum esculentum Moench (FEMC) (common Buckwheat). The biomass were collected, washed, dried, and crushed before carbonizing in a Muffle furnace electrically heated at 600  C for 4 hours. The activated carbon prepared was cooled to room temperature and washed with deionized water until the effluent was clear in color. Finally the synthesized carbon was dried in oven at 110  C for 12 hour. The synthesized carbon is chemically activated with 0.1 N solution HNO3 and H3PO4, respectively, under similar conditions to modify the chemical structure. The surface modification of carbons was also done by subjecting to liquid phase oxidation (Somasekhara Rao 1993).

Preparation of Lead (II) Solution Stock solution of Pb(II) was prepared (1000 mg/l) by dissolving required amount of Pb(NO3)2 in acidified double distilled water. The stock solution was diluted with distilled water to obtain desired concentration ranging from 5 to 45 mg/l. All the chemicals used were of analytical reagent grade. The lead (II) ions is analyzed by atomic absorption spectroscopy at 283.3 nm, using graphite furnace Analytikjena Vario-6. The calibration is carried out versus an aqueous standards curve.

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Adsorption Studies of Lead on Activated Carbons by Batch Method In order to understand the adsorption behavior of lead, various experimental parameters have been investigated using batch adsorption experiments. The effect of initial concentration is studied by varying lead concentrations between 5 and 45 mg/L. The effect of contact time was studied by varying the agitating time (range: 2–240 min) at fixed optimum initial concentration of lead (20 mg/L) with optimum dose of adsorbents (0.25 g/L), and also the effect of pH was studied ranging from 2 to 5.5. The percentage removal of the lead and the amount of lead adsorbed were calculated by the following equations. ðCi  Cf Þ Ci:   Ci  Cf V Amount adsorbed ðqeÞ ¼ M Percentage removal ¼ 100

where Ci and Cf are the initial and final equilibrium solution concentrations of the lead (mg/L), V is the volume of the solution (L), and M is the mass of the activated carbon (g). The data obtained have been analyzed for adsorption isotherms models and intraparticle diffusion model.

Results and Discussion Effect of Initial Concentration The effect of initial concentration on the adsorption of lead was studied by varying the initial lead concentration between 5 and 45 mg/L in 100 ml of lead solution and adding 1 g of adsorbent with a contact time of 120 minutes. Effect of initial varying concentration of Pb (II) ions showed that the percentage adsorption decreases with increase in initial concentration of the adsorbate. This decrease in Pb (II) ions uptake capacity with increase in initial metal concentration may be due to the formation of clusters of carbon particles resulting in decreased surface area (Montanher et al. 2005; Chen and Wu 2004). The variation of percent removal of lead with increasing initial concentration and variation of the amount of lead adsorbed with increasing initial concentration are shown in Figs. 3 and 4. The data reveal that under identical experimental conditions, the order of adsorption capacity of the various adsorbents is as BVC (HNO3) > BVC (H3PO4) > FEMC (HNO3) > FEMC (H3PO4). Adsorption Isotherm The relationship between the amount of a substance adsorbed per unit mass of adsorbent at constant temperature and its concentration in the equilibrium solution is called the adsorption isotherm (Feng et al. 2004). Various adsorption isotherm models are employed in this study to describe the experimental adsorption isotherm (Langmuir 1918; Freundlich and Helle 1939; Temkin and Pyzhev 1940).

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Fig. 3 Variation of percent removal of lead with increasing Initial concentration

BVC(HNO3) BVC(H3PO4) FEMC (HNO3) FEMC(H3PO4)

100 99

% Removal

98 97 96 95 94 0

Fig. 4 Variation of the amount of lead adsorbed with increasing initial concentration

10

5

40

50

40

50

BVC(HNO3) BVC(H3PO4) FEMC(HNO3) FEMC(H3PO4)

4

qe (mg/g)

20 30 Initial Conc. (mg/L)

3

2

1

0 0

10

20 30 Ci (mg/L)

(a) Langmuir adsorption model: The Langmuir adsorption model is the most common model used to quantify the amount of adsorbate adsorbed on an adsorbent as a function of partial pressure or concentration at a given temperature. It suggests the formation of monolayer adsorption and also the surface is energetically homogeneous (Feng et al. 2004). The Langmuir equation that is valid for monolayer adsorption onto a surface is given below

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qe¼

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K L Ce 1 þ aL Ce

The linear form of the Langmuir equation can be represented by Ce aL 1 ¼ Ce þ KL qe K L where Ce is the equilibrium metal ion concentration (mg/l), qe the amount of lead adsorbed at equilibrium (mg/g), and aL/KL (mg/g) and 1/KL (L/mg) are Langmuir constants related to adsorption capacity and energy of adsorption, respectively. When Ce/qe was plotted against Ce, a straight line with a slope of aL/KL was obtained as shown in Fig. 5. The R2 value of 0.8799–0.9962 is shown in Table 6. This showed that the models have good correlation coefficient in all the cases. The essential characteristics of the Langmuir equation can be expressed in terms of a dimensionless constant called separation factor (RL, also called equilibrium parameter) which is defined by the following equation (Özacar and Sengil 2004; Mall et al. 2006; Crini et al. 2007)

0.18

0.40

0.16

0.35

0.14

Ce/qe (g/L)

Ce/qe (g/L)

0.20

0.12 BVC(HNO3)

0.10 0.08

0.30 0.25 FEMC(HNO3)

0.20

0.06

0.15

0.04 0.0

0.2

0.4 Ce (mg/L)

0.6

0.8

0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6 1.8 Ce (mg/L)

0.24 0.65

0.22

0.60 0.55

0.18 0.16

BVC(H3PO4)

0.14 0.12

Ce/qe (g/L)

Ce/qe (g/L)

0.20

0.50 0.45

FEMC(H3PO4)

0.40 0.35 0.30

0.10

0.25

0.08

0.20 0.0

0.2

0.4 0.6 Ce (mg/L)

0.8

1.0

0.0

0.5

1.0 1.5 Ce (mg/L)

2.0

Fig. 5 Langmuir adsorption isotherms for the removal lead by different adsorbents

2.5

3.0

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Table 6 Comparison of Freundlich, Langmuir, and Temkin adsorption isotherm models Frendulich isotherm

R2

Log K

1/n

BVC (HNO3) BVC (H3PO4) FEMC(HNO3) FEMC(H3PO4) Langmuir isotherm

0.9991 0.9913 0.9932 0.9984 R2

0.5453 0.4380 0.4868 0.3619 Qmax

0.3166 0.3427 0.5213 0.6589 b

BVC (HNO3) BVC (H3PO4) FEMC(HNO3) FEMC(H3PO4) Temkin isotherm

0.9781 0.9962 0.9333 0.8791 R2

0.1408 0.3019 0.5649 0.5771 BT

BVC (HNO3) BVC (H3PO4) FEMC(HNO3) FEMC(H3PO4)

0.9629 0.9844 0.9342 0.9451

0.3529 0.4312 0.3245 0.3538

0.0588 0.0795 0.0943 0.1306 AT (L/g) 1.2776 1.8338 1.4681 1.3905

RL ¼

qe(expt.) (mg/g) 1.9840 1.9628 0.9823 0.9753 qe(expt.) (mg/g) 1.9840 1.9628 0.9823 0.9753 qe(expt.) (mg/ g) 1.9840 1.9628 0.9823 0.9753

qe(cal) (mg/g) 1.9647 1.9542 0.9235 0.9155 qe(cal) (mg/g) 1.9659 1.9392 0.9102 0.9040 qe(cal) (mg/ g) 1.9245 1.9142 0.9056 0.8956

χ2 0.0001877 0.0000376 0.0035195 0.0036664 χ2 0.0001651 0.0002837 0.0052923 0.0052122 χ2 0.0017844 0.0026411 0.0059281 0.0065125

1 1 þ a L Ci

where Ci (mg L1) is the initial adsorbate concentration and aL (L mg-1) is the Langmuir constant related to the energy of adsorption. The value of RL indicates the shape of the isotherms to be either unfavorable (RL > 1), linear (RL = 1), favorable (0 < RL < 1), or irreversible (RL = 0); the results from Table 2 show RL < 1 in all the cases; this indicates a favorable isotherm for the adsorbents used in this study. (b) Freundlich Adsorption Model: Freundlich isotherm is an empirical equation describing the heterogeneous adsorption and assumes that different sites with several adsorption energies are involved (Sekar et al. 2004; Crini et al. 2007). The linear form of the Freundlich equation is shown below. log qe ¼ log k þ

1 log Ce n

where Ce is the equilibrium metal ion concentration (mg/L), qe the amount of lead adsorbed at equilibrium (mg/g), K and 1/n are the Freundlich constants incorporating all the factors effecting adsorption capacity, an indication of favorability of metal adsorption onto adsorbent (Freundlich and Helle 1939). The slope 1/n gives adsorption capacity and intercept log K gives adsorption intensity from the straight portion of the linear plot obtained by plotting log qe versus log Ce which gives a straight line

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with slope of 1/n with value of 0.3166–0.3658 (Fig. 6); Freundlich constants KF and n were also calculated and are listed in Table 6. (c) Temkin Isotherm Model: The Temkin isotherm assumes that the heat of adsorption of all the molecules in a layer decreases linearly due to adsorbent-adsorbate interactions and that adsorption is characterized by a uniform distribution of binding energies, up to some maximum binding energy (Areco and Afonso Areco and Afonso 2010). The isotherm assumes that the fall in the heat of adsorption is linear rather than logarithmic as stated in Freundlich expression (Teles de Vasconcelos and Gonzalez Beca 1993). Unlike the Langmuir and Freundlich equation, the Temkin isotherm takes into account the interactions between adsorbents and metal ions to be adsorbed and is based on the assumption that the free energy of sorption is a function of the surface coverage (Deng and Ting 2005). The Temkin isotherm is applied in the following form RT ln ðAT :Ce Þ bT

0.8

0.8

0.6

0.6

0.4

0.4

0.2 BVC (HNO3) 0.0

Log qe (mg/L)

Log qe (mg/L)

qe ¼

0.2 BVC(H3PO4) 0.0 -0.2

-0.2

-0.4

-0.4

-1.4 -1.2 -1.0 -0.8 -0.6 -0.4 -0.2 0.0 Log Ce (mg/L)

0.8

0.8

0.6

0.6

0.4

0.4

0.2 0.0

FEMC (HNO3)

Log qe (mg/L)

Log qe (mg/L)

-1.8 -1.6 -1.4 -1.2 -1.0 -0.8 -0.6 -0.4 -0.2 0.0 Log Ce (mg/L)

0.2 0.0

-0.2

-0.2

-0.4 -1.4 -1.2 -1.0 -0.8 -0.6 -0.4 -0.2 0.0 0.2 0.4

-0.4

Log Ce(mg/L)

FEMC(H3PO4)

-1.0 -0.8 -0.6 -0.4 -0.2 0.0 0.2 0.4 0.6 Log Ce (mg/L)

Fig. 6 Freundlich adsorption isotherms for the removal of lead by different adsorbents

0.2

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D. Kibami

The linear form of Temkin equation is qe ¼

RT RT ln AT þ ln Ce bT bT

qe ¼ β lnα þ β ln Ce where β¼ bRT ; α = AT, T is the absolute temperature in Kelvin, R is the universal gas T constant, 8.314 J/mol K, bT is the Temkin constant related to heat of sorption (J/mg), and AT the equilibrium binding constant corresponding to the maximum binding energy (L/g). The Temkin constants AT and bT together with R2 values are shown in Table 6. A plot of qe vs. ln Ce yielded a linear line as shown in Fig. 7. (d) Validity of Adsorption Isotherm Models

5

5

4

4

3

3

2

BVC (HNO3)

qe (mg/L)

qe (mg/L)

Chi-Square Analysis To identify the suitable isotherm for sorption of lead (II) ions onto carbon adsorbents, the chi-square analysis was carried out. The mathematical statement for chi-square analysis is

1

2

1

0

0 -1.8 -1.6 -1.4 -1.2 -1.0 -0.8 -0.6 -0.4 -0.2 0.0 ln Ce (mg/L)

-1.4 -1.2 -1.0 -0.8 -0.6 -0.4 -0.2 0.0 ln Ce (mg/L)

4.5

4.5

4.0

4.0

3.5

3.5

3.0

3.0 qe (mg/L)

qe (mg/L)

BVC (H3PO4)

2.5 2.0 1.5

FEMC (HNO3)

2.5 2.0

FEMC (H3PO4)

1.5

1.0

1.0

0.5

0.5

0.0 -1.4 -1.2 -1.0 -0.8 -0.6 -0.4 -0.2 0.0 0.2 0.4

0.0

ln Ce (mg/L)

0.2

-1.0 -0.8 -0.6 -0.4 -0.2 0.0 ln Ce (mg/L)

Fig. 7 Temkin adsorption isotherms for the removal of lead by different adsorbents

0.2

0.4

0.6

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Studies on Water Quality of Mokokchung District, Nagaland, India, and. . .

χ2 ¼

1707

 2 X qeðexpt:Þqeðcal:Þ qeðexpt:Þ

where qe(expt.) and qe(cal.) are the experimental sorption capacity of lead (II) ions (mg g-1) at equilibrium time and the corresponding value that is obtained from the sorption model. If data from the model are similar to the experimental data, χ2 will be a small number, while if they differ; χ 2 will be a bigger number (Meeenakshi and Viswanathan 2007). χ2 values are low for Freundlich isotherm when compared to other isotherms for all the four sorbents. This is also in accordance with their corresponding R2 values.

Effect of Contact Time The effect of contact time on the removal of lead was studied by varying the agitating time (range: 2–240 minutes) at fixed optimum initial concentration of 100 ml lead (20 mg/L) with a dose of adsorbents (1 g/L). Removal efficiency and adsorption capacity of the adsorbents increase sharply in the initial stage and then gradually remain steady with the increase of agitation time (Fig. 8). The initial fast adsorption may be attributed to large uncovered surface area of adsorbents. With further increase in the agitation time, the availability of the uncovered surface area gradually diminishes. The uptake was slower at the final phase due to the internal surface adsorption attaining plateau region at equilibrium (Aroua et al. 2008; Lo et al. Lo et al. 1999). The equilibrium was attained after shaking for 120 min. After equilibrium was attained, the percentage of sorption did not change with further increase in time. The decrease in the extent of removal of lead after 120 min of contact time in some cases may be due to the desorption process (Itodo et al. 2009). Fig. 8 Effect of Contact time on percent removal of lead

100 99 98 % Removal

97 96 95

BVC (HNO3) BVC (H3PO4) FEMC(HNO3) FEMC(H3PO4)

94 93 92 0

50

100 150 Time (min)

200

250

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D. Kibami

Adsorption Kinetics One of the most important factors in designing an adsorption system is predicting the rate at which adsorption takes place, referred as the “‘kinetics of sorption.” Four simplified kinetic models were adopted to examine the mechanism of the adsorption process. (a) Pseudo-First-Order Model: Langergren and Svenska used rate equation to describe the adsorption of adsorbate from the liquid phase (Langergren and Svenska 1898); the linear form of pseudo-first-order rate expression is given as logðqe  qt Þ ¼ logqe 

k1 t 2:303

where qe and qt (mg g1) are the adsorption capacities of lead(II) ions at equilibrium and at time t (min), respectively, and k1 is the rate constant of pseudo-first-order kinetics. The correlation coefficient (R55) and pseudo-first-order constant (k1) for both adsorbents are summarized in Table 7. The values of R2 vary between 0.9682

Table 7 Comparison of kinetic models showing calculated qe(cal.) and experimental qe(expt.) values Pseudo-first-order reaction Adsorbent qe (expt.)(mg g1) BVC(HNO3) 1.9840 BVC(H3PO4) 1.9628 FEMC(HNO3) 1.9452 FEMC(H3PO4) 1.9322 Pseudo-second-order reaction Adsorbent qe (expt.)(mg g1) BVC(HNO3) 1.9840 BVC(H3PO4) 1.9628 FEMC(HNO3) 1.9452 FEMC(H3PO4) 1.9322 Intraparticle diffusion Adsorbent qe (expt.)(mg g1) BVC(HNO3) BVC(H3PO4) FEMC(HNO3) FEMC(H3PO4)

1.9840 1.9628 1.9452 1.9322

k1  101 0.3721 0.3573 0.2570 0.1967 K2 1  10 3.5404 5.0845 2.3164 2.8778 Kip 1  10 0.5730 0.7310 0.4840 0.5440

qe(Cal.)(mg g1)

SSE

R2

1.9602 1.9222 1.8290 1.6484

0.0001445 0.0004279 0.0035685 0.0215741

0.9850 0.9682 0.9898 0.9858

qe(Cal.)(mg g1)

SSE

R2

1.9711 1.9514 1.9373 1.9258

0.0000425 0.0000337 0.0000165 0.0000109

0.9999 0.9995 0.9994 0.9991

qe(Cal.)(mg g1)

SSE

R2

1.9069 1.8670 1.8850 1.8628

0.0015163 0.0023823 0.0009578 0.0012901

0.8767 0.9666 0.8317 0.8429

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and 0.9898, and the experimental qe values did not agree with the calculated values obtained from the linear plots (Fig. 9). (b) Pseudo-Second-Order Model: The pseudo-second-order rate expression, which has been applied for analyzing chemisorptions kinetics from liquid solutions (Ho 2004; Shrihari et al. 2005), is linearly expressed as: t 1 1 ¼ þ t 2 qt k2 qe qe where k2 is the rate constant for pseudo-second-order adsorption (g mg1 min1) and k2qe55 (mg g1 min1) is the initial adsorption rate. The linear plot of t/qt versus t, as shown in Fig. 10, yielded R2 values that were greater than 0.999 for all adsorbents. It also showed a good agreement between the experimental and the calculated qe values (Table 7). Hence, it can be concluded that the adsorption for both adsorbents was predominated by pseudo-second-order kinetic model. (c) Intraparticle Diffusion Model:

-1.6

-1.4 -1.6

BVC(HNO3)

-1.8

-2.0

Log (qe-qt)

Log (qe-qt)

-1.8

-2.2 -2.4

BVC(H3PO4)

-2.0 -2.2 -2.4 -2.6 -2.8

-2.6

-3.0 -2.8

-3.2 0

10

20

30

40

50

60

0

10

20

time (t) mins

30

40

50

60

time (t) mins

-1.2 -1.2 -1.4

-1.4 FEMC(HNO3)

-1.6 Log (qe-qt)

Log (qe-qt)

-1.6 -1.8 -2.0

FEMC(H3PO4)

-1.8 -2.0 -2.2 -2.4 -2.6

-2.2

-2.8 -3.0

-2.4 0

20

40 time (t) mins

60

80

0

20

40

60

time (t) mins

Fig. 9 Pseudo-first-order kinetic model fit for Pb(II) Ions adsorption on the adsorbents

80

1710

D. Kibami

Fig. 10 Pseudo-second-order kinetic model fit for Pb (II) Ions adsorption on the adsorbents

70 60 50

t/qt

40 30 BVC(HNO3) BVC(H3PO4)

20

FEMC(HNO3)

10

FEMC(H3PO4)

0 0

20

40 60 80 Time (t) mins

100

120

Intraparticle diffusion is a transport process involving the movement of species from the bulk of the solution to the solid phase. In a well-stirred batch adsorption system, the intraparticle diffusion model has been used to describe the adsorption process occurring on a porous adsorbent (Weber and Morris 1962; Özacar and Sengil 2005). According to Weber and Morris (1962), if the rate limiting step is intraparticle diffusion, a plot of solute adsorbed against the square root of the contact time should yield a straight line passing through the origin (Gerente et al. 2007). Also the rate constant for intraparticle diffusion is obtained from the slope of the curve. Weber and Morris (1962) theorized that the rate of intraparticle diffusion varies proportionally with the half power of time and is expressed as:   qt ¼ Kid t1=2 þ C where qt = adsorbate uptake at time t, (mg/g) and Kid = the rate constant of intraparticle transport, (mg/g-t 1/2), which can be evaluated from the slope of the linear plot of qt versus t 1/2. The values of intercept, c, are related to the boundary layer thickness, i.e., the larger the value of the intercept, the greater is the boundary layer effect (Shrihari et al. 2005; Weber and Morris 1962). Figure 11 shows the multilinearity in intraparticle diffusion plots for Pb(II) adsorption with different adsorbents. This indicates that intraparticle diffusion was not the only involved for Pb(II) adsorption, but there were some other processes involve in the rate controlling step. This suggests that adsorption occurred in three phases: the initial steeper section represents surface or film diffusion, the second linear section represents a gradual adsorption stage where intraparticle diffusion is rate-limiting, and the third section is final equilibrium stage. Thus, there were three processes controlling the adsorption rate.

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2.00 1.98 1.96

qt (mg/g)

1.94 1.92 1.90 BVC (HNO3) BVC (H3PO4) FEMC(HNO3) FEMC(H3PO4)

1.88 1.86 1.84 0

2

4

6

8 10 t (1/2) min

12

14

16

Fig. 11 Intraparticle diffusion plots for the removal of lead by adsorption on various adsorbents

(d) Validity of the Kinetic Models: The best-fit among the kinetic models was assessed by the squared sum of errors (SSE) values. It is assumed that the model that gives the lowest SSE values is the best model for the particular system (Adhikary et al. 1989; Ho et al. 2000; John Freund 1981). The SSE values were calculated by the equation.

SSE ¼

 2 X qeðexpt:Þqeðcal:Þ qe 2 ðexpt:Þ

where qe(expt.) and qe(cal.) are the experimental sorption capacity of Lead (II) ions (mg g-1) at equilibrium time and the corresponding value that is obtained from the kinetic models. SSE values are least for pseudo-second-order kinetic model when compared to other models; this is also in accordance with their corresponding R2 values which suggest kinetic adsorption follows pseudo-second order.

Effect of pH pH is one of the key parameters which controls the adsorption efficiency by influencing the surface charges (Kannan and Rengasamy 2005). It is an important environmental factor influencing not only site dissociation, but also the solution chemistry of the heavy metals: hydrolysis, complexation by organic and inorganic ligands, redox reactions, and precipitation. It strongly influences the speciation and adsorption availability of heavy metals (Esposito et al. 2002). The effects of initial

1712

D. Kibami 99 98 97

% Removal

96 95 94 93

BVC (HNO3) BVC (H3PO4) FEMC (HNO3) FEMC(H3PO4)

92 91 1.5

2.0

2.5

3.0

3.5

4.0

4.5

5.0

5.5

6.0

pH

Fig. 12 Effect of pH on percent removal of Lead

pH on lead solution were investigated by varying the pH from 2 to 5.5 and using initial lead concentration of 20 mg/L in 100 ml of lead solution with 1 g of adsorbent. The lead adsorption usually increases as the pH is increased attaining optimum capacity at pH: 5.5 (Shukla et al. 2002; Jalali et al. 2002). Above pH 5.5, Pb(II) starts precipitating as Pb(OH)2 and hence studies above this range are not conducted (Patnukao et al. 2008) (Fig. 12).

Conclusions Adsorption studies of lead on activated carbons was studied by batch method where effect of initial concentration and pH were the parameters considered for determining the adsorption efficiency of the carbon samples. The results of the percentage removal of lead increased with the increase of contact time and pH. On the contrary, the percentage of removal decreased with the increase in initial concentration of the standard lead solution. The equilibrium data were analyzed using Langmuir, Frendulich and Temkin adsorption isotherms. Frendulich isotherm displayed a better fitting model than the other two models with a higher correlation coefficient of 0.99915 to 0.99138 indicating a multilayer adsorption. Three adsorption kinetic models were studied; the pseudo-second-order kinetic model accurately described the adsorption kinetics. The results show intraparticle diffusion was not the only rate limiting factor in adsorption of Pb(II) ions. The adsorption mechanism was found to

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be chemisorption and the rate-limiting step was mainly surface adsorption. Results from this study suggest that the both the adsorbents under study are very effective adsorbent for Lead (II) ions, as anticipated.

Cross-References ▶ Assessment of Some Aspects of Provisioning Sewerage Systems: A Case Study of Urban Agglomerations in Ganga River Basin ▶ Water Quality Assessment of an Unexplored Tropical Freshwater System in Thiruvananthapuram, India: A Multivariate Statistical Approach Acknowledgments The authors acknowledge the staff of SAIF, NEHU Shillong for providing necessary laboratory facilities for analyzing lead samples.

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Özacar M, Sengil IA (2004) Equilibrium data and process design for adsorption of disperse dyes onto alunite. Environ Geol 45:762–768 Özacar M, Sengil IA (2005) A kinetic study of metal complex dye sorption onto pine sawdust. Process Biochem 40:565–572 Patil VT, Patil PR (2010) Physicochemical analysis of selected groundwater samples of Amalner town in Jalgaon district, Maharashtra, India. Electron J Chem 7(1):111–116 Patnukao P, Kongsuwan A, Pavasant P (2008) Batch studies of adsorption of copper and lead onactivated carbon from Eucalyptus camaldulensis dehn bark. J Environ Sci 20:1028–1034 Rajappa B, Manjappa S, Puttaiah ET (2010) Monitoring of heavy metal concentration in groundwater of Hakinaka Taluk, India, contemporary engineering. Sciences 3(4):183–190 Rao SM, Mamatha P (2004) Water quality in sustainable water management. Curr Sci 87(7):942–947 Ricordel S, Taha S, Cisse I, Dorange G (2001) Heavy metals removal by adsorption onto peanut husks carbon: characterization, kinetic study and modeling. Sep Purif Technol 24:389–401 Rivail Da Silva M, Lamotte M, Donard OFX, Soriano-Sierra EJ, Robert M (1996) Metal contamination in surface sediments of mangroves, lagoons and southern bay in Florianopolis Island. J Environ Technol 17:1035–1046 Saeed A, Iqbal M, Akhtar MW (2005) Removal and recovery of lead(II) from single and multimetal (Cd, Cu, Ni, Zn) solutions by crop milling waste (black gram husk). J Hazard Mater 117:65–73 Sarala Kumar D, Rao PR (1993) Endemic fluorosis in the village Ralla Anantha puram in Andhra Pradesh. An epidemiological study. Fluoride 26(3):77–180 Scoullos MJ, Vonkeman GH, Thornton HI, Makuch Z (2001) Handbook for sustainable heavy metals policy and regulation. Mercury – Cadmium – Lead. Springer-Science & Business Media. Dordrecht, The Netherlands Sekar M, Sakthi V, Rengaraj S (2004) Kinetics and equilibrium adsorption study of lead(II) onto activatedcarbon prepared from coconut shell. J Colloid Interf Sci 279:307–313 Shrihari V, Madhan S, Das A (2005) Kinetics of phenol sorption by raw agrowastes. Appl Sci 6(1):47–50 Shukla A, Zhang YH, Dubey P, Margrave JL, Shukla SS (2002) The role of sawdust in the removal of unwanted materials from water. J Hazard Mater 23:95–137 Sohail A, Qadeer R (1997) Kinetic study of lead ion adsorption on activated carbon. Adsorp Sci Technol 15:815 Somani LL, Gandhi AP, Paliwal KV (1972) Note on the toxicity of fluorine in well waters of Nagpur and Jaipur district of Western Rajasthan. Indian J Agric Sci 8:752–754 Somasekhara Rao K (1993) Correlations among water quality parameters of ground waters of Nuzvid town and Nuzivid mandal. Indian J Environ Prot 13(4):261–266 Somasekhara Rao K, Someswara Rao B (1995) Correlations among water quality parameters of ground waters of Musunur mandal, Krishna District. Indian J Environ Prot 14(7):528–532 Suess MJ (1982) Examination of water for pollution control, vol 1 & 2, 1st edn. Pergamon press, Oxford Teles de Vasconcelos LA, Gonzalez Beca CG (1993) Adsorption equilibria between pine bark and several ions in aqueous solution Cd(II), Cr (III) and Hg(II). Eur Water Pollut Control 3:29–39 Temkin MJ, Pyzhev V (1940) Kinetics of ammonia synthesis on promoted iron catalysts. Acta Physiochim USSR 12:217–222 The Report of the Scientific Review Committee. Ottawa (1990) Department of National Health and Welfare (Canada). Nutrition recommendations Tong S (1998) Lead exposure and cognitive development: persistence and a dynamic pattern. J Paediatr Child Health 34(2):14–118 Trivedy RK, Goel PK (1986) Chemical and biological methods for water pollution studies. Environmental publications, Karad U.S. Environmental Protection Agency (EPA) (1991) Guidance for water quality-based decisions: the TMDL process. Washington, DC. Doc. No. EPA 440/4-91-001

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Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Advancements in Photocatalysts . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Synthesis of Various Bismuth Compounds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Photocatalyst Modification and Doping . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Supported Photocatalysts . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Multivariate Analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Energy Consumption and Economic Analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Electrical Energy Determination . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Total Operating Cost . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Photocatalytic Disinfection in Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Water Quality . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Turbidity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Oil and Grease . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Trace Contaminants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion and Future Prospects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cross-References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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S.-M. Lam (*) Department of Environmental Engineering, Faculty of Engineering and Green Technology, Universiti Tunku Abdul Rahman, Kampar, Perak, Malaysia e-mail: [email protected] J.-C. Sin Department of Petrochemical Engineering, Faculty of Engineering and Green Technology, Universiti Tunku Abdul Rahman, Kampar, Perak, Malaysia A. R. Mohamed School of Chemical Engineering, Universiti Sains Malayisia, Engineering Campus, Nibong Tebal, Pulau Pinang, Malaysia © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_93

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Abstract

Semiconductor photocatalytic process has been garnered a great interest in the topic of green technology as it aims at the total elimination or at least the minimization of waste production and the sustainable processes implementation in the water or wastewater industry. The viable of this green technology has been extensively shown to destruct a myriad of toxic organic pollutants and microbes in water. This chapter reviews the recent advancement of engineeredphotocatalysts, modification strategy of photocatalysts, process optimization, as well as energy consumption and economy analysis on the photodegradation processes for water treatment. A number of potential photocatalysts are discussed, in particular the bismuth-based photocatalysts. This chapter also describes the utilization of multivariable approach to evaluate the optimum parameters synthesis and process variables so as to enhance the photodegradation efficiencies. The energy consumption and economy analysis that elucidated the light utilization efficiency in the photocatalytic water treatment are presented. A brief discussion on the photocatalytic disinfection of microbes in the water treatment as an alternative waste treatment process is outlined. The water quality that significantly influenced on the photodegradation performances is also detailed. This present chapter will provide a useful scientific and technical information to researchers and engineers who partake in this area. Keywords

Organic pollutant · Semiconductor · Photocatalysis · Solar · Environmental control

Introduction Environmental pollution as well as the lack of natural energy resources has garnered significant interest to the crucial need for chemical technologies and ecologically clean materials. Photocatalysis is one of the most promising solutions not only to circumvent the environmental pollutions but also as an efficient route to harness solar energy as an alternative natural sources due to the depletion of conventional energy sources. This process is also technically defined as an increase in the rate of a thermodynamically allowed (ΔG < 0) reaction in the presence of photocatalysts with the increase originating from the creation of some new reaction pathways containing photogenerated species and a reduction of the activation energy (Lam et al. 2012). The catalysts in photocatalytic reactions are normally semiconductor materials, which have a band gap that separated the top of the electron-filled valence band from the vacant conduction band. Upon irradiation of semiconductor materials with light energy equivalent to or greater than their band gap energies (Eg), the electron (e) excited from the valence band to the conduction band, thereby creating an e and a hole (h+) in the conduction band and valence

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band, respectively. A band model is often used for schematic illustration of the electronic structures of semiconducting materials (Fig. 1). Both the photogenerated e and h+ diffused to the surface of semiconductors to undergo reactions either oxidation or reduction with the absorbed compounds. The combination of e with oxygen formed superoxide anion (O2•–) radicals, followed by reacting with H+ (producing hydroperoxyl (HO2•) radicals) and then further decomposition to form hydroxyl (•OH) radicals. The •OH radicals are powerful oxidants to degrade the organic pollutants in the photocatalytic system. At the same time, the adsorbed water molecules on the surface of semiconductors are oxidized by h+ in the valence band to form •OH radicals. It was also noted that the photogenerated e can easily recombine with h+ after their generation in the absence of e or h+ scavengers. In this regard, the presence of species scavengers is crucial for inhibiting the charge carrier recombination and for enhancing the efficiency of photocatalytic reaction (Paola et al. 2012; Dong et al. 2015). To date, the most extensively used photocatalyst in the research of water and wastewater treatment is Degussa P25 TiO2 catalyst. This catalyst is typically utilized as a standard reference for comparisons of UV photoactivity under different treatment conditions. On the contrary, exploration of a visible light active photocatalyst is highly warranted and the light absorption of suitable energy is determined by the band gap of the semiconductors. Among the semiconductors, bismuth photocatalysts is a new type of promising visible light-responsive photocatalyst for the organic pollutants. Degradation owing to its appropriate band gap and excellent chemical stability (Sivakumar et al. 2014; Meng and Zhang 2017; Lam et al. 2017). Their photoactivities can be enhanced by size and morphological changes on the catalytic surfaces which lead to the reduction of the recombination of e–h+ pairs. On the contrary, doping metal oxides have a positive impact on the photocatalytic effect, which can be attained by allowing charge carrier transportation between the photocatalyst, and coupling with an additional phase could delay the recombination of charge carriers. Therefore, the preparation and utilization of bismuth photocatalysts have attracted a great interest (Tseng and Lin 2014; Henriquez et al. 2017; Wang et al. 2017; Wu et al. 2017). Fig. 1 Schematic of the mechanism of photocatalytic conversion of organic pollutants

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Different bismuth phosphate (BiPO4) nanostructures such as particle-like, rodlike, needle-like, and rice-like were selectively synthesized to study the effect of BiPO4 morphologies on the degradation of methyl orange (MO). The highest activity of BiPO4 nanoparticles could be ascribed to its intrinsic distortion of PO4 tetrahedron and the smaller band gap structure (Li et al. 2011). Various bismuth oxyhalides BiOXs (X = Cl, Br, I) (Wang et al. 2015) and bismuth ferrite oxides (BiFeO3) (Gao et al. 2016) were also found to be capable of photo-destruction of either in visible light or non-absorbing visible light water pollutants. This review has surmounted the synthesis methods of bismuth compounds related to morphologies and their modified strategies for photocatalytic efficiencies over various toxic and persistence organic pollutants. The future challenges and opportunities on the way to implement photocatalytic materials have been outlined to assist on the development of energy research and finding ways to approach for the major problems.

Advancements in Photocatalysts Synthesis of Various Bismuth Compounds Synthesis of BiOX (X = F, Cl, Br, I) Bismuth oxyhalides BiOX (X = F, Cl, Br, I) are bismuth-based semiconductors that have demonstrated promising applications in pigments, pharmaceuticals, catalysts, and gas sensors as a result of their outstanding optical and electrical properties (Michel et al. 2011; Chen et al. 2013; Zou et al. 2015; Ganose et al. 2016; Zhang et al. 2014). All BiOX crystallized in the tetragonal matlockite structure have layered structure characterized by [Bi2O2] slabs interleaved by double slabs of halogen atoms. Electrical structures of BiOX were simulated based on density functional theory (DFT) and the findings are shown in Fig. 2 (Huang and Zhu 2008). Both the valence and conduction bands of BiOX composed of X np (n = 2–5 for X = F, Cl, Br, and I, respectively), O 2p, and Bi 6p orbitals. The theoretical measured values of band gap for BiOX (X = F, Cl, Br, and I) were 2.79, 2.34, 1.99, and 1.38 eV, respectively. Comparatively, the experimentally calculated band gap values were 3.64 (Su et al. 2010), 3.22, 2.64, and 1.77 eV (Zhang et al. 2008), respectively. Both theoretically and experimentally measured results showed that the band gap of BiOX decreased with an increase in the atomic number of halogen. Their differences in the band gap values can be due to the limitations existed within the generalized-gradient approximation (GGA) technique. A number of approaches have been successfully used for the synthesis of BiOX with different morphologies and structures (Lin et al. 2007; Zhang et al. 2008; Yu et al. 2012; Li et al. 2013; Sin et al. 2017). Hydrolysis can be easily done with the use of simple equipment. It was based on the reaction between soluble bismuth salts and oxyhalides or H2O. For example, Yu et al. (2012) obtained the pure plate-like BiOX (BiOCl and BiOBr) nanostructures as efficient photocatalysts quickly and directly from the hydrolysis of BiX3 (BiCl3 and BiBr3). The BiX3 used in their study can serve as both bismuth and halogen sources for the synthesis of BiOX. Using the

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Fig. 2 Band structures of BiOX: (a) BiOF, (b) BiOCl, (c) BiOBr, and (d) BiOI (Adapted from Huang and Zhu 2008)

hydrolysis method, Li et al. (2015a) fabricated sheet-like, honeycombed-like, and flower-like BiOBr by a reaction between BiBr3 with water, ethanol, and isopropyl alcohol, respectively, at room temperature. The results in their HRTEM analysis also found that the crystalline facets were varied from (102), (101) to (110), respectively (Fig. 3). They went further to explain the viscosities of the solvents played important roles in the growth of BiOBr crystals, which led to different morphologies and crystalline facets of BiOBr. Under solar light irradiation, sheet-like BiOBr showed superior photocatalytic activities and stability toward the degradation of methyl orange. They concluded that the water-based hydrolysis can be carried out simply, environmental friendly, and energy efficiently, which leads to an ideal technique to fabricate BiOBr photocatalysts for the practical applications. Hydrothermal/solvothermal processes were also widely used for the preparation of BiOX due to their numerous advantages such as low production cost, short processing time, easily scalable with high yield, environmentally benign, and relatively low temperature synthesis. These processes are usually carried out in an autoclave (a steel pressure vessel) under controlled temperature and pressure. The operating temperature is held above the boiling point of the solvent to selfgenerate saturated vapor pressure. Zhang et al. (2008) synthesized hierarchical BiOX (X = Cl, Br, and I) microspheres via a solvothermal route by autoclaving the mixture of Bi(NO3)3∙5H2O, KCl/NaBr/KI, and ethylene glycol at 160  C for 12 h. Their findings showed that the synthesized samples exhibited tetragonal phase spherical-like BiOBr with grain sizes in the range of 1–10 μm. They also added

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Fig. 3 TEM and HRTEM images of (a, b) sheet-like BiOBr, (c, d) honeycombed-like BiOBr, and (e, f) flower-like BiOBr prepared via a hydrolysis method (Adapted from Li et al. 2015a)

Fig. 4 Schematic illustration of formation process of hierarchical BiOX (X = Cl, Br, I) microspheres (Adapted from Zhang et al. 2008)

that the synthesized samples consisted of a large quantity of BiOBr nanoplates and assembled to construct the hierarchical microspheres with the help of ethylene glycol. A possible growth mechanism of these hierarchical microspheres has been proposed and divided into three steps: (1) the formation of BiOX nanoparticles and their growth into nanoplates at the early stages, (2) ethylene glycol-induced selfassembly of primary nanoplates to form loose microspheres, and (3) the formation of regular hierarchical microspheres through a dissolution-recrystallization process (Fig. 4). Their photocatalytic results showed that all the BiOX products exhibited higher degradation efficiencies of organic dye as compared to that of TiO2 under UV-visible and visible light irradiation. Solid state method can also be utilized to synthesize BiOX. Nevertheless, the requirement of high reaction temperature led to the synthesized products polluted by metal ions from the reactants and not being easy purified. Such observation has been reported by Lin et al. (2007) that some Na+ was present in bismuth oxyhalide products prepared via a solid state method. Hence, the solid state method has barely been utilized recently.

Synthesis of BiFeO3 BiFeO3 (BFO) has attracted a great deal attention owing to its possible applications as nanogenerators, spintronics, sensors, piezoelectric devices, and ferroelectric diodes (Bharathkumar et al. 2015; Papadas et al. 2015; Lam et al. 2017). In addition, band gap of BFO material (~2.2–2.8 eV) is lying in the visible light region of solar spectrum which extends its horizons to the photovoltaics and photocatalysis areas.

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The BFO exhibits multiferroic and ferroelectricity behaviors and has a rhombohedrally distorted perovskite structure with a space group R3c at room temperature (Lam et al. 2017). The BFO material shows a Neel temperature of 370  C and a curie temperature of 830  C. A myriad of methods have been established for the synthesis of nanometer-sized and micrometer-BFO crystallites including conventional solid-state reaction, coprecipitation, sol–gel process, and hydrothermal route in literatures (Valant et al. 2007; Wei and Xue 2008; Wang et al. 2010a; Safi and Shokrollahi 2012). Using a solvothermal method, Huo et al. (2010) revealed that BFO microsphere displayed eight times higher photoactivity for the visible light degradation of MB than that observed for TiO2 (Degussa P25) because of its ability to degrade the organic pollutants. In another study, Bharathkumar et al. (2015) found that BFO mesh was able to degrade up to ~98% of the MB dye after 4 h exposure to sunlight irradiation. They reported that the BFO mesh photocatalysis enhancement was attributed to the BFO band-bending process and enhanced interaction between the photocatalyst and dye molecules. This band bending at different band edge positions created an additional pathway for the transportation of photogenerated e–h+ at the photocatalyst-dye interface region, leading to the reduction of charge carrier recombination and hence enhanced photocatalytic activity. For degradation of organic dye solution, dye molecules can absorb visible light and in some cases their visible light absorption was intense than that of BFO, making it difficult to determine the photoactivity originated from the photocatalysts. On the contrary, for the organics without absorbing visible light including 4-chlorophenol and phenol, the photoactivities of the catalyst were typically studied in the presence of irradiation above 400 nm (Huo et al. 2011; An et al. 2013; Papadas et al. 2015). Under reaction conditions (4-chlorophenol concentration = 1.0  104 M; volume of 4-chlorophenol = 30 mL; reaction time = 4 h), aerosol-spraying prepared BiFeO3 (BFO-2-500) can degrade 4-chlorophenol under the exposure of visible light and that the photoactivity was superior to those of solid state preparation of BiFeO3 (BFO-SSR) and grinded BiFeO3 (BFO-2-500(G)) (Fig. 5a) (Huo et al. 2011). Huo and his co-researchers also observed BFO-2-500 (BFO-2-500) can effectively decompose RhB under visible light irradiation as compared to those of BFO-SSR and BFO-2-500(G) (Fig. 5b) (Huo et al. 2011). Nevertheless, in general, it is still inconvenient to prepare BFO compounds without impurity phases due to the low stability of Bi atom in perovskite BFO and it is not easy to achieve the structure-controllable BFO with high surface area (Xu et al. 2016). Additionally, the fabrication of BFO sometimes is based on complex solution processes and involves toxic precursors (Sakamoto et al. 2008). Thus, it is indispensable to employ an environmental-friendly method to produce the pure BFO photocatalyst. Stability and reusability of the catalyst need to be comprehensively studied before being devoted into practical applications. Papadas et al. (2015) showed that BFO 3D mesoporous network not only can be reused for three catalytic runs but can also be recovered easily from the 4-nitrophenol mixture using an external magnet (Fig. 6). The recovering process of catalyst from the reaction medium is

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a BFO-2-500

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Fig. 5 (a) Degradation of RhB solution in the presence of solid state prepared BFO (BFO-SSR) and grinded BFO (BFO-2-500(G)), aerosol-spraying prepared BFO (BFO-2-500) nanoparticles under visible light irradiation (λ > 420 nm) and (b) photo- degradation of 4-chlorophenol over solid state preparation of BFO-SSR and grinded BFO-2-500(G), BFO-2-500 nanoparticles under visible light irradiation for 4 h (Adapted from Huo et al. 2011)

another important factor for their practical applications. According to the abovementioned discussion, BFO owned magnetically attractive property and thus it was found to be relatively easier to filter from the mixtures. Nevertheless, the photocatalytic performance of durable use and lifetime of the BFO still presence of data shortage which required further investigation. Also, the safety of BFO photocatalyst has no report yet. In general, the Bi, Fe, and O elements are bonded in a form of perovskite-type material with thermal and chemical treatments, and hence, the photocatalysts should be focused on safety issue of human and environment.

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Fig. 6 Catalytically recyclable degradation of 4nitrophenol by BFO 3D mesoporous network with inserted isolation of the catalyst using an external magnetic field (Adapted from Papadas et al. 2015)

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Time (s)

Fig. 7 Crystal structures of (a) HBIP, (b) nMBIP, (c) mMBIP (Adapted from Pan et al. 2011)

Synthesis of BiPO4 Bismuth orthophosphate (BiPO4) demonstrated many advantages as photocatalysts such as excellent photocatalytic activity, exceptional optical and electronic properties, high mineralization ability, stable chemical structure, and fast rate of settlement separation. It is recognized to have three crystal structures: hexagonal phase (HBIP, space group: P3121), monoclinic phase (nMBIP, space group: P21/n), and monoclinic phase (mMBIP, space group: P21/m) as indicated in Fig. 7. The structure of BiPO4 can be simply described by one bismuth atom surrounded by eight oxygen atoms as well as one phosphate atom surrounded by four oxygen atoms. Among these crystal structures, HBIP with a band gap of 4.6 eV exhibited the lowest photocatalytic activity. Both the nMBIP and mMBIP demonstrated higher photocatalytic activity and their band gap were determined to be 3.8 eV and 4.2 eV, respectively (Li et al. 2011; Pan et al. 2011; Pan and Zhu 2015). BiPO4 photocatalyst was firstly prepared by Pan and Zhu (2010) via a hydrothermal process between Bi(NO3)3.6H2O and Na3PO4•12H2O for the degradation of

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organic dyes. Their results showed that the synthesized BiPO4 was a promising UV-responsive photocatalyst and its photocatalytic activity was about twice higher than that of the well-known Degussa P25 (Fig. 8a). Both the high position of valence band and the high e–h+ pair separation efficiency led to the high photocatalytic activity of BiPO4. They also added that the inductive effect of PO43 played an important role in assisting the e–h+ pair separation. Subsequently owing to its low cost, nontoxicity, and high photoefficiency, BiPO4 has been widely documented for the environmental purification. A facile and rapid microwave irradiation method has also been used for the synthesis of BiPO4 nanostructures as active photocatalysts for methyl orange degradation under UV-visible irradiation (Li et al. 2011). By varying the solvents, different morphologies of BiPO4 nanostructures such as rice-like, nanoparticles, nanorods, nanoneedles have been synthesized. Their proposed formation mechanism of BiPO4 nanostructures was a diffusion-controlled process in organic solvents, which greatly ascribed to variation of diffusion rate and concentration of monomer in different solvents (Fig. 8b). In addition to degradation of organic pollutants from aqueous solution, BiPO4 has been reported as photocatalysts for the conversion of gas phase benzene into CO2. Nevertheless, BiPO4 has poor visible light absorption due to its wide band gap and thus limiting its practical widespread application.

Fig. 8 (a) Photocatalytic kinetic constants on the degradation of methyl orange, rhodamine B, and 4-chlorophenol over BiPO4 and P25 (Adapted from Pan and Zhu 2010). (b) Schematic illustration of formation mechanism of BiPO4 nanostructures synthesized in different solvents (Adapted from Li et al. 2011)

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Photocatalyst Modification and Doping To absorb larger fractions of the solar spectrum for photocatalysts, numerous material engineering solutions have been developed such as morphology control, doping, heterojunction structures, and combining the carbon materials (Li and Yan 2009; Li et al. 2013; Park et al. 2013; Li et al. 2014a; Fan et al. 2015; Meng and Zhang 2017). The basis in using these material engineering strategies is to balance both the half-reaction rates of the photocatalysis by modifying the catalyst structure and composition as well as adding electron acceptors. By using these methods, either the charge carrier separation of photocatalyst or solar light harvesting can be effectively improved, which thus enhanced the photocatalytic performance. A lot of research works have been performed in order to gain optimum morphologies and crystal structures that are competent of improving the visible light photoactivity of photocatalyst powders. Rhombohedral BFO powders synthesized by a PVP-assisted hydrothermal route at 180  C for 72 h in the presence of different alkaline conditions obtained different morphologies: spindle-like structures (0.5 M NaOH), cube-like particles (2 M NaOH), and plate-like structures (4 M NaOH) (Huang et al. 2014). The formation of the different BFO structures was owing to the crystal facets matching as shown in their XRD patterns and HRTEM images. The photoactivity of the BFO plates with (104) facets exposed was the best among the three photocatalysts for visible light degradation of MO. Using microwave synthesis, Li et al. (2012) observed 3D flower-like BiOBr nanostructures possessed large surface area and exhibited excellent removal capacity and fast adsorption rate for Cr(VI) ions in a wide pH range. Owing to the great difference in intrinsic features of distinct elements, searching proper dopants for a certain photocatalyst also played a key role in enhancing the photocatalytic performance. Yin et al. (2011) have systematically studied the properties of metal-doped BiVO4 by using first-principles density functional theory (DFT), which indicated that monoclinic BiVO4 with minimized photo-generated carrier recombination could be achieved by doping Br, Ca, Na, or K at O-rich growth conditions or Mo, W at O-poor growth conditions. Remarkable enhanced photoactivity of W, Mo, Co, or Cu doped BiVO4 has also been found by experiments (Yao et al. 2008; Parmer et al. 2012; Park et al. 2013). In addition, research works on nonmetal doping bismuth-related semiconductors have garnered much attention. Zhang and Zhang (2010) revealed self-doping BiOIs with I/Bi molar ratio varying from 1 to 3. Their experimental results as well as DFT calculations revealed that iodine self-doping could efficiently alter the BiOI optical absorption property. Their photocatalytic activities were tested by degradation of MO and removal of nitric oxide. 1.5 of I/Bi ratio was found to be the best. Improved photoactivity has also been demonstrated in F-doped BiVO4 and F-doped Bi2WO6 (Fu et al. 2008; Jiang et al. 2013). In the photocatalytic reaction, heterojunction was established when the band gaps of two semiconductor materials were overlapped and thus e–h+ can be transferred in direct migration from one level to another with the existence of noble metallic (e.g., Au, Pd, Pt, and Ag) or conducting interface. Li et al. (2015b)

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reported that Ag@AgCl anchored on (040) crystal facet BiVO4 (Ag@AgCl/ BiVO4) was a visible light-responsive photocatalyst in which the metallic Ag species not only act as the solid-state electron mediator but can also absorb the photons from incident light and present the surface plasmon resonance (SPR) effect (as illustrated in Fig. 9a). This process could decrease the recombination of charge carriers and thus improved the photoactivity in the visible light degradation of RhB (Fig. 9b). As shown in Fig. 9c, under visible light irradiation, BiVO4 and metallic Ag could be excited. Metallic Ag can act as a bridge transporting e from its conduction band minimum (CBM) to CBM of AgCl, while photogenerated e in CBM in BiVO4 will migrate to the metallic Ag to facilitate the e–h+ separation in AgCl and BiVO4 alone. Novel Ag/AgBr/BiOBr hybrid with superior visible light–driven photocatalytic activities in sterilization of pathogenic organism and degradation of organic dye has also been reported by Cheng et al. (2011). Besides applying a noble metal heterojunction, the utilization of a semiconductor heterojunction can also act as an efficient photocatalyst system with promising visible light activity. Ren et al. (2011) fabricated a heterostructured photocatalyst containing the same Bi, Mo, and O elements (Bi3.46Mo0.36O6/

Fig. 9 (a) UV–vis diffuse reflectance spectra of the BiVO4, Ag/BiVO4, Ag@AgCl/BiVO4, and AgCl materials, (b) photocatalytic degradation of RhB in the presence of different photocatalysts under visible-light irradiation, and (c) schematic diagram illustration of Z-scheme mechanism photodegradation of RhB dye over Ag@AgCl/BiVO4 heterostructure photocatalyst under visible light irradiation (Adapted from Li et al. 2015b)

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Bi2MoO6). The as prepared samples could efficiently mineralize organic substances into CO2 under exposure of visible light, which was owing to high e–h+ pair separation between the heterojunction interfaces. Consistent reports including BiOI/BiOBr, Bi2O3/Bi2MoO6, BiVO4/Bi2O3, Bi2O3/Bi2WO6, Ag2O/ Bi2MoO6, Ag3PO4/Bi2MoO6, Bi2S3/BiOI, and AgI/BiOI have also been found in the literatures (Li and Yan 2009; Cao et al. 2011; Gui et al. 2011; Cao et al. 2012a, b; Xu et al. 2013; Meng and Zhang 2017). As for carbon materials, they have a great potential in a large variety of environmental clean-up applications owing to their outstanding characteristics of corrosion resistance, strong thermal stability, large surface area, and good electronic properties. Earlier reports have stated that the carbon material–semiconductor material junctions can hinder the e–h+ recombination as well as enhance the pollutants adsorption capacity on the composite, leading to the improvement of photoactivity. It was reported that the photoactivity can be improved when BiFeO3 coupled with graphene (GR) (Li et al. 2013, 2014a) and graphitic carbon nitride (g-C3N4) (Fan et al. 2015). Other bismuth-related materials such as BiOBr (Zhang et al. 2012), Bi2O3 (Wang et al. 2010b), and Bi2O2CO3 (Madhusudan et al. 2012) have also been successfully composited with graphene.

Supported Photocatalysts In catalyst-based photocatalysis technologies, agglomeration of nano-sized photocatalysts is one of the main reasons to use a support material (Lam et al. 2010). Many types of porous materials such as zeolite, carbon, silica, and others have been then investigated as photocatalyst supports in the photodegradation of numerous organic pollutants (Pozzo et al. 1997). These support photocatalysts are essential to be in micro or as mesoporous catalyst to facilitate easily particle separation from the fluids (liquid or gaseous) and enable them to be packed inside a vessel with fluids moving through the reactor system. Nevertheless, it is required to strongly anchor the nano-sized photocatalyst to the surface, for instance, functional groups bonding to the surface or defect sites on the surface and photocatalyst itself (Stevensa et al. 2005; Sebastian et al. 2010). The agglomeration can still take place if the nano-sized photocatalysts is not strongly deposited on the surface of the supports. Another merit of support photocatalysts are concerned on the porosity and surface area with respect to the adsorption of the organic pollutant on the photocatalyst surface. It seems clear that a large surface area provided by accessible pores is crucial for concentrating the target pollutant around the active photocatalyst. Consequently, the direct oxidation of h+ could be the main oxidation pathway since the adsorption of organic pollutants on the semiconductor surface was the prerequisite step for direct charge carrier transfer. Alternatively, direct oxidation by •OH radicals required the adsorption of water or OH ions on the catalyst surface to produce the •OH radicals.

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Multivariate Analysis Despite catalyst-based photocatalysis is an emerging technology utilized for the degradation of persistent compounds in the last few decades, some reports revealing on the statistical design of experiment (DOE) application toward the photocatalytic process are available (Lam et al. 2012; Callao 2014). Multivariate analysis is one of these statistical DOE approaches, which helps to account for interaction effects between the studied variables and determine more accurately the combination of levels that produces the optimum of the process. By comparing with the traditional one-factor-at-a-time (OFAT) or univariate approach, one variable involved in the photodegradation is varied, while holding the others variables constant. Moreover, the result of this univariate analysis indicates insufficient optimization toward response(s). In this regard, the OFAT approach is costly in sense of time and reagents and not that efficient. There is now increasing exploration of process optimization using experimental field, where it can be analyzed using commercial statistical software such as SAS, Minitab, and Design Expert. The first step during the development of this analysis is the selection of appropriate response(s) or output(s) since it should take into account the sources of error, ways of minimizing it, and of course the ability to follow the change in response(s) in course of time. For instance, the possible response(s) could be percent of degradation and decolorization rate or even degree of mineralization of the sample. On the contrary, variables that influence the response could be photocatalyst synthesis factors such as dopant content, heating temperature, photocatalyst dose, surfactant amount, or operation factors such as initial concentration of target compound, pH, light irradiance, and electron acceptors. The key of DOE is based on the selection of the experimental points at which the response could be examined. The selection of appropriate DOE have a great impact on the constructing of response surface and hence its accuracy on the prediction. Abdullah et al. (2012) proposed the use of central composite design (CCD)-DOE approach together with analysis of variance, statistical regression, and response surface analysis to study the combined effects of three key operation variables that affect the visible light photocatalytic removal using BiVO4. They utilized 3 experimental variables to analyze the 51 possible variables combination. It was reported that the interaction between the photocatalyst dose and solution pH had a positive synergistic effect on the overall removal rate. A response surface model was created to correlate the removal rate dependency on the three different variables according to the statistical regression as shown in Eq. (1): Y ¼ b0 þ

X

bi X i þ

XX ij

bij Xi Xj þ

XX X

bijk Xi Xj Xk þ e

(1)

jk

where Y is the predicted response; i, j, and k take values from the number of parameters; b0 is the constant; bi represents linear coefficient; and bij and bijk are quadratic and cubic coefficients where i = j and i = j = k, respectively. These refer to interactions of first or second order, respectively, Xi, Xj, and Xk are the levels of

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independent parameters and e is the random error. This model (Eq. 1) is empirical and independent for each photocatalytic system. The following verification works are required to measure the accuracy and applicability of such model for the prediction of photoreaction rate under the variation of its variables. Other DOE approach to photocatalytic system optimization has also been applied over the years such as factorial design, Box–Behnken design, and Taguchi design (Mera et al. 2016, 2017; Xiao et al. 2017).

Energy Consumption and Economic Analysis Although several studies are available in the literature on electrical energy determination of a variety of photocatalysis processes, it is essential to evaluate the electrical energy consumption and economic analysis of the photocatalysis under experimental conditions. Typically, the electrical energy consumption of photocatalysis governs by a number of experimental factors (type of pollutant being treated, configuration of the reactor, and type of light source used, etc.) and thus, electrical energy determination becomes crucial for the processes investigation. A number of factors including economics, effluent quality, cost, etc. also show a vital role in selecting a wastewater treatment technology. In literatures, the study of energy consumption and economic analysis of bismuth-based photocatalysts was scarcely reported. Hence, the following section is focusing on the photocatalysts which are typically found in the literatures.

Electrical Energy Determination Since the photocatalytic visible light degradation process is an electrical energy intensive process, the electrical energy denotes a major fraction of the operating costs. The figures-of-merit based on electrical energy per order (EEo) can be informative and useful as they allow for a rapid determination of the electrical energy cost. The EEo is defined as the number of kilowatt hours of electrical energy required to reduce the concentration of pollutant by 1 order of magnitude in a unit volume of contaminated water. The EEo (kWh/m3/order) can be obtained from the following Eq. (2). EEo ¼

P  t  1000 Ci V  60  log Cf

(2)

where P is the input power (kW), t is the irradiation time (min), V is the volume of water (L ) in the reactor, and Ci and Cf are the initial and final pollutant concentrations, respectively. This equation is implicitly related to the Langmuir-Hinshelwood (L-H) model first-order kinetics as shown in Eq. (3).

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Ci In Cf

¼ kapp t

(3)

By combining the Eq. (3) and the L-H first-order kinetics, the EEo can be expressed as shown in Eq. (4). EEo ¼

3:84  P V  kapp

(4)

Typical EEo for a range of organic pollutants is provided in Table 1. EEo makes the scale-up and comparison of relative treatment performance very simple. For example, from Table 1, 4-nitrophenol has an EEo of 72.3 kWh/m3∙order, while Acid Red 27 has EEo of 12.6 kWh/m3∙order (Behnajady et al. 2011). Acid Red 27 requires 5.7 times more energy for the same level of degradation as 4-nitrophenol. Thus, under similar influent and effluent conditions, the treatment system for Acid Red27 would require five times the UV power of that treating 4-nitrophenol in the photocatalytic system.

Total Operating Cost Cost evaluation is one of the most pivotal characteristics in water treatment. The overall costs can be indicated by the sum of the capital cost, operating cost, and maintenance cost. For conventional water and wastewater treatment systems, these costs strongly depend on the nature, concentration of pollutant, and the configurations of the reactor used. Thus, the cost evaluation was performed on the basis of the EEo determined for the entire experimental runs. The total operating cost for the photodegradation of organic pollutants is demonstrated in Eqs. (5) and (6). Totaloperationcost ðUSD=kgÞ Totalenergy consumed per mg

of pollutant removal ðkWhÞ USD  Unit cost of power  106 kWh ¼ pollutant removal ðmgÞ

(5)

Totalenergy consumed per mg of pollutant removal ðkWhÞ ¼

Powerinput ðkWÞ  Reactiontime ðminÞ 1000  60

(6)

It must be highlighted again from Table 1 that the reported studies have been conducted on photocatalytic degradation of organic pollutants using different experimental conditions with respect to, for instance, pH, temperature, and so on could strongly influence the obtained total operating cost values (Pujara et al. 2007; Benotti et al. 2009; Mehrjouei et al. 2013; Mahamuni and Adewuyi 2010).

log((TOC)i/(TOC)f)

Slurry

Thin film

Slurry

8.17

10

10 5 20 20 80 Degussa TiO2

TiO2-coated bead TiO2 nanoparticles

Slurry Slurry Thin film

27 3 15

Slurry

Degussa TiO2 CdS TiO2/ZnO/ chitosan Y-TiO2-ZSM

Slurry

10

Thin film

25

Slurry

Thin film

10

Catalyst Si-doped TiO2 TiO2-coated beads H3PW12O40/ TiO2 CdS nanosphere CdS/ZnS

44

Reactor type Slurry

Concentration (mg/L) 50

UV (one unit) UV (two units) UV (three units)

Forgacs and Cserhati (2004) Lachheb et al. (2002) Repo et al. (2013) Zhu et al. (2012)

1.17  103

18.4 85

0.13a 0.42a

39.5 79 118.5

1.48  104

1.56  1010

Pujara et al. (2007)

Behnajady et al. (2011)

Okte and Yılmaz (2008) Li et al. (2014b)

Lin et al. (2008)

6.99  103

1.21  102 1.58  102 3.59  1010

Lu et al. (2012)

Li et al. (2014b)

References Xie et al. (2010)

1.92  1010

1.12  104

EEO (kWh/m3 order) 8.00  1010

34.8 9.23 12.6 72.3 39.9

15

UV

2.59

10.60

0.053 0.0212 11.55

0.5

0.033

18.00

3.42

kapp (min1) 4.32

0.12 0.62 0.46 0.4 0.03a

10

120

UV UV

125 3 300

500

300

300

10

Power input (w) 150

UV LED UV-visible

Visible

Visible

Visible

UV

Light source UV-visible

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Phenol-4sulfonic acid

Malachite green Acid Red 27 4-Nitrophenol

Methyl orange

Methylene blue

Organics Rhodamine

Table 1 Evaluate of the electrical costs of the organic pollutants degradation 69 1735

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Similarly, conducting the real wastewater treatments can also alter these values, depending on the pollutant contents in the wastewater.

Photocatalytic Disinfection in Water The effectiveness of the photocatalysis technology in antimicrobial was first demonstrated by Matsunaga et al. (1985), who reported that the Lactobacillus acidophilus, Saccharomyces cerevisiae, and Escherichia coli were completely sterilized under metal halide lamp irradiation for 60–120 min. By far, most of the investigations were focused on the Gram-negative bacterial model Escherichia coli. For example, enhanced photocatalytic bacteriostatic activity toward Escherichia coli has been reported by Wang et al. (2016) via the visible-light driven BiOI/ BiOBr hierarchical microspheres. Their studies on the determination of cell structure destruction and the release of K+ proved the rupture of cell membranes during the photocatalysis. They also added that the h+ is the major species on the photocatalytic disinfection of Escherichia coli. Recently, Ahmad et al. (2016) reported that the death of Escherichia coli cells over BiOBr was due to the •OH radicals that formed during photocatalysis. The attack of •OH radicals led to the oxidation of functional groups in the cell membrane and interrupted the membrane permeability. This can lead to the leakage of intracellular substances, demolition of microbial structure, and ultimately the inactivation of Escherichia coli. In addition to Escherichia coli, a wide range of microorganisms including fungi, viruses, and many types of bacteria were also successfully inactivated by photocatalytic treatment.

Water Quality The constituents in wastewater can be derived from domestic, municipal, and industrial sources such as bath water, food waste, human excreta, personal and household maintenance products along with a broad range of other inorganic and organic compounds in trace amounts. Given a vast diversity of constituents that may be observed in the wastewater, it is common practice to characterize the water quality parameters in the wastewater. The following discussion is introduced on the water source of different qualities which is scarcely of concern in the catalystbased photocatalysis.

Turbidity Turbidity is a measure of water clarity. This parameter is an integrated measurement of the extent to which light is either absorbed or scattered by the suspended and dissolved particulates in the water. The turbidity may be caused by a wide variety of causes. For instance, the turbidity may be caused by glacier-fed rivers and lakes,

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rivers descend from mountain areas and plains, rivers under flood conditions, and rivers moving toward the ocean. In particular, as rivers moving toward the ocean, they pass through the urban areas where the industrial and domestic wastewaters treated or untreated may be added. The industrial wastes may include a huge quantity of organic and/or inorganic compounds that form turbidity. Certain domestic wastes may also include considerable numbers of organic and inorganic matters that produce turbidity. The presence of such particulate matters is greatly unfavorable to the catalystbased photocatalysis since they can influence the optical clarity and further obstruct the light penetration in the water (Chin et al. 2004; Tang and Chen 2004). This will lead to disparity in the anticipated utilization of light penetration path and its intensity as well as catalyst dose. Furthermore, if excessive turbidity is not removed from water, it can impede both the mineralization and disinfection efficiency of the photodegradation of organic pollutants as those pollutants might flee from the water treatment due to the shielding effects that attenuate the light penetration (Lam et al. 2012). Since photocatalysis processes are retrofitted to advanced water treatment steps, prior removal of turbidity could be attained by some traditional treatment methods including screening, settling, filtration, coagulation, and flocculation. Nevertheless, high cost of treated wastewater makes most of the people in the country sides decrease toward promptly accessible to the low-quality sources, leading them to contract the numerous waterborne diseases (Jadhav and Mahajan 2013). A series of guidelines have assigned appropriated turbidity levels for some common uses in the industrial field. According to the World Health Organization (2011), the turbidity in drinking water must be between 1 and 5 nephelometric (NTU) (Lapena 2000). The US Environmental Protection Agency has set more stringent turbidity standards for drinking water beginning in January 2002. Turbidity in treated drinking water must never exceed 1 NTU and must not exceed 0.3 NTU in 95% of dairy samples in any 1 month.

Oil and Grease Oil and grease are one of the important water quality parameters owing to their poor solubility in water and their tendency to separate from aqueous phase. Despite this property given a merit in facilitating the oil and grease separation by utilizing flotation devices, it can cause destruction in wastewater treatment units, complication in waste dissemination via pipelines, and disposal problems into receiving waters. Particularly, wastes from meat-processing industries where saturated fats from the slaughtering of chicken and cattle are partook and the restaurants have resulted in serious incline in the carrying capacity of wastewater. On the other hand, not all the grease and oil from crude and refined petroleum industries are discharged by primary clarifier units. Considerable amounts still remain in the settling wastewater in a finely divided emulsified form. Such conditions have stimulated many researchers to surmount for oil and grease

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removal in the wastewater prior to their discharge into receiving waters. For instance, Ghasemi et al. (2016) reported that even at high strength of petroleum refinery wastewater, the photodegradation on the TiO2 surface was efficiently accomplished. However, they noted that the efficiency of photodegradation was enhanced when TiO2 immobilized on Fe-ZSM-5 zeolite was used instead of unloaded TiO2 particles. The effect of various factors such as catalyst dosage, COD, pH, and reaction time on the photodegradation of petroleum waste water on TiO2 surface was investigated (Aljuboury et al. 2015). In another example, the process of refining crude palm oil consumes large amounts of water in the palm oil mill. Tabassum et al. (2015) reported that palm oil mill will generate about 2.5 m3 of palm oil mill effluent (POME) in every tonne of crude palm oil produced and this is equivalent to about 10.5 m3 of POME for every tonne of fresh fruit bunch processed. Consequently, the discharge of untreated POME into waterway would cause severe wastewater-containing oil and grease pollution. Ng et al. (2017) used TiO2 and ZnO photocatalysts to photocatalytic polishing of POME in Malaysia. The results showed that TiO2 exhibited higher photoactivity toward POME than ZnO photocatalysts. In quite similar studies, WO3 nanocrystals have also been demonstrated to efficiently degrade POME under visible light irradiation (Cheng et al. 2017). According to the Clean Water Act (CWA) of the United States of America, European Union (EU), and Malaysia Sewage and Industrial Effluent Discharge Standards Regulation, the grease and oil concentration in discharge water should not be more than 15 ppm, 5 ppm and 10 ppm, respectively (Abdullah 1995; Chakrabarty et al. 2008). Hence, oil emulsified in water in the range of 100–1000 ppm is considered as a major water pollutant.

Trace Contaminants Natural sources from both plant and animal as well as from the synthetic organic chemistry are the sources of trace inorganic and organic compounds in water supplies. In surface water, decomposition of plant materials from algae and fungi bring chemicals of wide diversity that cause taste and odors. Some of these chemicals are sulfur-bearing compounds including butyl mercaptan, isopropyl mercaptan, and methyl mercaptan. The exudation of natural plants can also discharge highly refractive colored substances. Earlier report from Chirstman and Minear (1971) has revealed that except for tannin, the exact nature of these substances still remains unknown, but they are polymeric, have high molecular weight, and are phenolic in character. These compounds are typically referred to as humic materials owing to their similarity to humus in soil-organic matter. Major concern related to trace contaminants in water bodies, nevertheless, originates from a plethora of synthetic chemicals that gain access to the surface waters and some ground supplies via the discharge of wastewaters from municipal and industrial complexes. These synthetic organic chemicals such as pharmaceuticals, pesticides, personal care products, and hormones that are typically found from spills

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that result from accidents that occur during transportation by land and water, from uncontrolled use or discharge of these materials to the environment. At an appropriate level of exposure, certain trace organic contaminants (TrOCs) have potential to cause chronic disorder in endocrine system, which affect the physiological and reproduction development of wildlife and humans (Sin et al. 2012). During the wastewater treatment, their behaviors are dependent on chemical properties. For instance, hydrophilic and nonbiodegradable substances are unaffected by wastewater treatment and hence perseveres in their original forms in the wastewater. The existence of TrOCs in the effluent could cause the dissemination of these contaminants to receiving water bodies. In light of the known risks to TrOCs in water quality and ecosystem health, there is also significant interest in the identification of trace inorganic contaminants (TrICs) from the effluent. For instance, Cu2+ and Zn2+ ions impart a metallic taste to water. In the surface water, Cu2+ ion is commonly used as the sulfate salt to control growth of algae in water supply reservoirs but it is toxic to aquatic plant at concentration below 1.0 mg/L (US EPA 2000). The toxicity of Zn2+ ions is very low (~5 mg/L) (US EPA 2000). The Zn2+ ions pass through the water environment from electroplating industries effluents, mining operations, and corrosion of galvanized piping. On the other hand, Al3+ ions in the form of alum are typically utilized as coagulant in water treatment. The presence of Al3+ ions was set at 0.02–0.5 mg/L to prevent its post-precipitation and discoloration of water in the distribution system (US EPA 2000). The Ag+ ions are not important pollutants of natural water, but soluble Ag salts are good disinfectants. Even a small ingestion of soluble silver can cause a darkening of the skin and eyes (argyria). Few discussions have been performed on the effects of different inorganic anions or cations on both catalysts photodegradation and photomineralization reactions. A general summary from these works revealed that Na+, Mn2+, Cu2+, Cl, HCO3, SO42, Mg2+, and Fe2+ ions at certain levels may influence photomineralization reaction efficiency, while Ca2+, Al3+, and Zn2+ ions may have negligible effects (Guillard et al. 2005; Habibi 2010; Lam et al. 2012; Santiago et al. 2014; Borthakur et al. 2017; Dugandzic et al. 2017). Dugandzic et al. (2017) observed that the effect of anions on the reaction was Cl > SO42 > F > NO3, while the effect of cations was Na+ > Ca2+ > Al3+ because of the small adsorption of nicosulfuron on the TiO2 catalyst surface. They attributed to this fact that the inhibitory effect of inorganic ions can be attributed rather to the scavenger role of anions than to the competitive adsorption. The presence of anions such as Cl, Br, NO3, SO42, CH3COO as well as cations such as Na+, NH4+, Mg2+, and Ca2+ ions has been investigated on the zwitterionic nature of the nanocomposite and their influence on the photodegradation efficiency toward degradation of azo dyes. Borthakur et al. (2017) found that the catalytic property of the nanocomposite was greatly influenced by its surface property and the different inorganic ions present in the system. The Na2+, Ca2+, and Al3+ ions were observed to have no discernible effects on the TiO2 degradation and mineralization of imazalil at pH 7 (Santiago et al. 2014). However, at acid pH the presence of Al3+ solution hindered both imazalil degradation and mineralization. Hence, the presence of inorganic ions in water subjected to

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catalyst-based photocatalytic treatment is an important factor in determining its successful implementation.

Conclusion and Future Prospects Photocatalytic water treatment process is a potentially promising technology for organic pollutant oxidation. The applicability of the photocatalytic technology for water treatment is constrained by several key technical issues ranging from catalyst development to process optimization has to be addressed. These included a myriad way by which the photocatalyst, particularly bismuth-based photocatalyst, can be modified for a high photoactivity. Modification may partake at its bulk, at the surface of photocatalyst or forming composite structures using any type of materials such as metals, oxides, semiconductors, carbonaceous materials, as well as support on porous materials. In order to promote the feasibility of photocatalytic water treatment, the process optimization of the systems has also highlighted. In general, photocatalysis studies were used an one-factor-at-a-time (OFAT) or univariate approach to investigate the sole variables effect on the treatment process. However, the account for interaction effects between the studied variables in the photocatalytic system will increase if it is incorporated with multivariate analysis approach. Additionally, the electrical energy per order (EEO) analysis can provide information regarding on treatment efficiency with potential overall saving in energy and chemical use. A broad range of microorganisms including Escherichia coli, fungi, viruses, and many types of bacteria were successfully inactivated by photocatalytic treatment. The water quality on the process performances in terms of the mineralization and disinfection was also elucidated. The potential of other constituents present in water to react with the parent organic compounds needs to be continuously assessed and monitored to meet the increasingly strict regulations. Despite substantial progress has been accomplished, there are some vital technical points that require to be further examined. The first challenge is lack of comprehensive understanding of the visible light photocatalytic enhancement mechanisms by bismuth photocatalytic materials. Fundamentally understanding the mechanisms is pivotal as the visible light activity and stability of bismuth compounds relied on further understanding of the mechanisms and their relationship with the catalyst active-site structures and composition. The use of novel material characterizations and computational techniques are advisable in the future research works. The practices used are not only intended at design and tailor-made catalyst model development at molecular and electronic levels but can also be utilized to analyze the virtual importance of some synthetic factors as well as under multi-interaction conditions. A direct and systematic comparison of bismuth photocatalytic materials with other visible light catalyst materials is also indispensable to provide potential application for the next-generation solar photocatalyst systems. A demonstrated ability to use bismuth compounds at a solar pilot scale for effluent purification processes would certainly benefit the commercial sector both in terms of

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environment and economy. Additionally, most of the literature studies concentrated only on the degradation rate and efficiency of target organic dyes disregarding the toxicity issues of the degradation intermediates. This aspect should not be overlooked while reporting any future work. Undeniably, there are still many challenges and prospects for bismuth compounds and they are still anticipated to be developed as potential photocatalysts to circumvent various environmental and energy-related issues.

Cross-References ▶ Wastewater Management to Environmental Materials Management Acknowledgments This work was supported by the Universiti Tunku Abdul Rahman (UTARRF/ 2016–C2/S03 and UTARRF/2017-C1/L02) and Ministry of Higher Education of Malaysia (FRGS/ 1/2015/TK02/UTAR/02/2 and FRGS/1/2016/TK02/UTAR/02/1).

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Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Characteristics of Paper and Pulp Mill Effluent . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Treatments of Paper and Pulp Mill Effluent . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biological Treatment of Paper and Pulp Mill Effluent . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Role of Bacteria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Role of Fungi . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Role of Microbial Enzymes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Technologies for Enhancing the Biological Treatment Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Bioreactors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cells/Enzyme Immobilization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Nano-biotechnology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Constrains and Future Thrust in Biological Treatment System . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cross-References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

Economic development of a nation is linked to industrialization, but should not take place at the expense of environmental degradation. The demand for paper and cardboard in packaging industries is continuously on increase. However, paper industry is extremely water intensive and also an obnoxious polluter of the environment, thereby being categorized under the red category of pollution control boards. Pulp and paper mill effluents consist of not only lignin and other naturally occurring polymers but also many xenobiotic compounds (chlorinated lignins, resin acids and phenols, dioxins, furans, chlorophenols, adsorbable organic halogens (AOX), extractable organic halogens (EOXs),

S. Chaudhry (*) · R. Paliwal Institute of Environmental Studies, Kurukshetra University, Kurukshetra, India © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_134

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polychlorinated biphenyls, polychlorinated dibenzodioxins, plasticizers, etc.), which can cause severe toxicity to aquatic life by bioaccumulation and may lead to biomagnification in food chains. Salt-rich black liquor can also deteriorate the soil structure, increase soil salinity, and cause nutrient imbalance in crops when used for irrigation without any prior treatment. Hence, appropriate treatment of this liquor prior to discharge into the environment is crucial. Although conventional treatments are quite effective in decolorization of paper mill effluents, all have severe setbacks such as high cost of treatment or unreliability in operation. Thus, this chapter deals with the recent developments in the technologies for treatment of wastewater generated from paper and pulp industries. Some of these include use of environmental biological agents with their enhanced enzymatic systems and new materials for cell immobilization that have received considerable attention in the recent years. The processes related to these technologies, their economic benefits, and constraints if any will be discussed. Keywords

Black liquor · Hardwood · Softwood · Agro-residues · Color · Lignin · Chlorinated xenobiotic · Resin acids · Pulping · Pulp · Kraft process · Sulfate process · Sulfite process · Chemical recovery · Bleaching · Decolorization · Washing of pulp · Recovery · Clarification · Conventional treatments · Biological treatment · Lagooning · Biocatalyst · Depolymerization · Enzymes · Extracellular enzymes · Lignin peroxidase · Laccase · Manganese peroxidase · Glutathione S-transferases · Dioxygenases · Monooxygenases · Phenol oxidases · Xylanase · Versatile peroxidase · Accessory enzymes · Bioreactors · Biofilm · Sequential bioreactor · Immobilization · Nano-biotechnology · Nanomaterials · Enzyme immobilization · Biomolecules · Magnetic support

Introduction In the past few decades, an enormous industrial development has taken place at a rapid pace. This industrial revolution has totally changed the environment from the earlier conditions. The paper and pulp industries are extremely energy and resource intensive in terms of fossil fuels, electricity, water, raw materials (wood and non-wood materials), and chemicals. In the production of 1 ton of paper, pulp and paper industry consumes 250–300 m3 water. As a result, paper and pulp mills discharge a variety of gaseous, liquid, and solid wastes into the environment. Thus, these have been categorized by the Central Pollution Control Board (CPCB) among the 17 most polluting industries. There are 759 paper and pulp mills in India, out of which 30 are wood-based large scale, 150 are agriculture residue-based medium-scale, and 579 are recycled fiber-based medium and small-scale mills (Rajwar et al. 2017). Pulping and bleaching are the two important processes involved in paper manufacturing. In pulping, wood chips are generally treated with the alkaline solution of sodium hydroxide and sodium sulfate/sulfite at high temperature and

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pressure to convert wood chips into pulp. During pulping, lignin and hemicelluloses are separated out of wood and released in the effluent as waste. The lignin-rich effluent generated during paper manufacturing is called as black liquor due to its black color. Another process called bleaching involves the brightening of pulp with chemical treatments such as chlorine, hydrogen peroxide, ozone, etc. Effluents generated during bleaching contain highly chlorinated xenobiotic compounds such as chlorinated lignin, resin acids, phenols, dioxins, furans, chlorophenols, adsorbable organic halogens (AOX), extractable organic halogens (EOXs), plasticizers, etc. These compounds are highly recalcitrant and tend to persist in nature. Moreover, the pollutants released from paper and pulp mill are reported to affect aquatic life, deteriorate the soil structure, increase soil salinity, and cause a nutrient imbalance in crops (Ali and Sreekrishnan 2001; Yadav et al. 2010; Pathak et al. 2013). Among the paper and pulp industries, the large-scale mills have a chemical and energy recovery system that leading to reducing the pollution loads. However, due to unavoidable installation and operational cost, small-scale mills often lack such recovery systems. As a result, these industries discard their unrecovered or partially treated effluent into the environment. Therefore, there is a need to employ a costeffective technique to remove the harmful pollutants from the effluent generated during paper manufacturing, prior to discharge. Conventional techniques, such as ion exchange, resin separation, reverse osmosis, and advanced oxidation, are available for the treatment of paper and pulp mill effluent, but these techniques are not cost-effective and may lead to generation of secondary pollutants. Biological methods involving the application of bacteria, fungi, and actinomycetes are comparatively preferable, in terms of cost and being environmentally friendly (Pant and Adholeya 2007). Although a plethora of information is available on biological treatment methods for black liquor, there is an acute shortage of efforts to make the process being implemented effectively on large scale.

Characteristics of Paper and Pulp Mill Effluent The final wastewater released from the paper and pulp industries is the combined effluent produced from different processes of paper manufacturing. The characteristics and the volume of effluent depend upon the nature of raw material such as hardwood, softwood, and agro-residues used and the type of manufacturing process adopted in papermaking. For example, softwood contains higher quantity of resin acids than hardwood, therefore, contributes tannins and resin acids to wastewater. Resin acids commonly found in tree bark and wood of conifers are the diterpenic carboxylic acids which are hydrophobic, nonvolatile, and unsaturated in nature. Their hydrophobicity makes them harmful to aquatic fauna (Lindholm-Lehto et al. 2015). Nature of the wastewater also depends on the different processes and steps of papermaking (Karrasch et al. 2006; Raj et al. 2014). Figure 1 summarizes the main pollutants, which are normally produced during several steps of pulp and papermaking process.

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Paper and Pulp Making Processes

1. Raw Material Preparation

Suspended solids including bark particles, fiber pigments, dirt, grit, BOD and COD

2. Pulping

Color, bark particles, soluble wood materials, resin acids, fatty acids, AOX, VOCs, BOD, COD and dissolved inorganic compounds

3. Bleaching

Dissolved lignin, color,COD, carbohydrate, inorganic chlorines, AOX, EOX, VOCs, chlorophenols and halogenated hydrocarbons

4. Paper Making

Particulate wastes, organic and inorganic compounds, COD and BOD

Fig. 1 Major pollutants released from different processes of paper manufacturing

• Pulping: Chemicals (such as sodium hydroxide and sodium sulfate/sulfite), used in the process of pulping makes the water highly alkaline. Kraft process (also known as sulfate process) uses sodium sulfate, sodium hydroxide, and sodium sulfide in pulp making, whereas the sulfite process uses magnesium or calcium bisulfate and sulfurous acid. The alkali process uses the sodium hydroxide or lime in pulping. The effluent released from these processes is known as black liquor and contains a large amount of lignin and unused chemicals. The black liquor may or may not be treated for chemical recovery, depending upon the economy and the size of mill. Generally, small-scale paper mills do not treat their black liquor for chemical recovery, therefore causing the severe pollution problem. • Bleaching is the decolorization or brightening of cellulose fibers (pulp) obtained from the process of pulping. It involves treatment of pulp with several bleaching agents, such as chlorine, chlorine dioxide, hydrogen peroxide, oxygen, ozone, etc., that contribute highly recalcitrant toxic compounds of chlorinated lignin and phenols to the effluent. • Washing of pulp: Washing pulp with alkali caustic soda removes all the color and bleaching agents from the pulp, further making the effluent more alkaline in nature (Ali and Sreekrishnan 2001). • The bleached pulp is then mixed with a variety of filler material such as alum, talc, dyes, etc. in beater tank. The refined pulp is further processed by diluting and screening to remove any lumps. The wastewater produced from this section is known as white water and contains fine fibers, alum, talc, etc. The white water is usually reused for wet chipping process (Rao and Dutta 2016). A well managed and kraft process employing paper and pulp mill, generally produces 225–320 m3 wastewater from the production of 1 ton of the paper. Characteristics of effluent released from large and small paper manufacturing industries are given in Table 1.

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Table 1 Characteristics of effluent released from large- and small-scale paper and pulp mills Parameters Volume of wastewater (flow/day) Color pH Total solids (mg/l) Suspended solids (mg/l) COD (mg/l) BOD (mg/l) COD/BOD ratio

Large-scale industry 222 m3/ton 7800 units 8.5–9.5 4410 3300 716 155 4.6

Small-scale industry 330 m3/ton – 8.2–8.5 – 900–2000 3400–5780 680–250 3.9–5

Source: Rao and Dutta (2016)

The wastewater produced from the small-scale mills is generally highly polluted compared to the large-scale mills, as the effluents produced from different sections of large mills are generally treated for chemical recovery prior to discharge. However, in small mills, the entire black liquor is wasted as effluent.

Treatments of Paper and Pulp Mill Effluent Conventional treatment of paper and pulp mill effluent involves various techniques for the reduction of pollution load. Different treatment techniques are discussed in Table 2.

Biological Treatment of Paper and Pulp Mill Effluent Biological treatment involves the application of bacteria, fungi, and their enzymes as single or in combination with the different conventional processes for the degradation of lignin-rich effluent. Biological methods are not only suitable for the reduction of pollution load from the effluents but they are also cost-effective and eco-friendly. Highly recalcitrant, complex, and random structure of lignin restricts the microbial degradation. However, microbes have developed a unique strategy to overcome this restriction for intact lignin molecules.

Role of Bacteria Bacteria can play a crucial role in the biological treatment of paper and pulp effluent due to their immense environmental adaptability and biochemical versatility. Several bacterial species have been reported to decolorize the paper and pulp effluent and remove lignin efficiently from the wastewater. The bacterial species such as Bacillus subtilis and Micrococcus luteus were found to reduce the major pollution load such as BOD (87.2%), COD (94.7%), and lignin (97%) from the paper and pulp mill

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Table 2 Different treatments for the reduction of pollution load from the paper and pulp mill Processes Recovery Color removal

Activated carbon for color removal Clarification (physical treatment) Biological treatment

Lagooning

Land treatment

Disposal by irrigation

Details The lignin from black liquor is recovered by acid treatment with carbon dioxide or sulfuric acid Lime treatment is used to remove color and BOD from the paper mill effluents. The color absorbed by the lime and the sludge thus produced is settled Acidic activated carbon can be used to remove color from generated wastewater The suspended solid from combine mill effluent are cleaned mechanically in clarifiers. Removal of more than 90% of suspended solid can be achieved in clarifier Major reduction in the BOD can be achieved by application of different biological treatment processes, such as aerated lagoon, anaerobic lagoon, activated sludge process, trickling filter, etc. The black liquor is segregated from other wastes and stored in a lagoon, which is discharged later into the stream under the favorable conditions Soil removes the color by the cation-anion exchange property. The wastes from paper industries are spread over soil to absorb the colorimparting agents The treated wastewater from paper industries can be utilized for irrigation

(Tyagi et al. 2014). Although a number of bacterial species can degrade only monomeric lignin forms, few strains are also reported to attack lignin derivatives released from different pulping processes (Hao et al. 2000; Chandra and Bharagava 2013). Bacillus megaterium and two strains of Pseudomonas aeruginosa have been reported to decolorize bleach kraft effluent by Tiku et al. (2010). Some most efficient bacterial strains reported to treat pulp and paper mill effluent are listed in Table 3. Keharia and Madamwar (2003) compared the degradation potential of Pseudomonas, Ancylobacter, and Methylobacterium for organochlorine from bleached kraft pulp and paper mill effluents. They observed that Ancylobacter showed the broadest substrate range but could significantly reduce the AOX from softwood effluents only, whereas Methylobacterium with limited substrate range was capable of degrading AOX from both hardwood and softwood effluents. The bacterial degradation of lignin is limited compared to fungi. However, bacteria offer great potential and possess a unique class of enzymes for the depolymerization of lignin. de Oliveira et al. (2009) reported the production of xylanases and manganese-dependent peroxidase during the treatment of paper and pulp mill effluent by Bacillus pumilus CBMAI 0008 and Paenibacillus sp. CBMAI 868. Some other bacterial-originated lignin-degrading enzymes involved laccases, glutathione S-transferases, ringcleaving dioxygenases, monooxygenases, and phenol oxidases (Paliwal et al. 2012). Bacteria cannot grow effectively on lignin, and the presence of all the lignin-degrading enzymes in a single culture is rather difficult; therefore, application of more than two strains could overcome the same problem. Bacteria can effectively

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Table 3 Bacterial species used for the treatment of paper and pulp mill effluent Bacterial species Pseudomonas aeruginosa

Pollution parameters Color (26–54%)

Aeromonas formicans

COD (71%) and lignin (78%)

Pseudomonas fluorescens

Color (75%), phenol (66%), COD (79%), and lignin (45%)

Pseudomonas, Ancylobacter, Methylobacterium

Organochlorines

Paenibacillus sp., Aneurinibacillus aneurinilyticus, and Bacillus sp. Novosphingobium sp. B-7

Color (39–61%), lignin (28–53%), BOD (65–82%), COD (52–78%), and total phenol (64–77%)

Serratia marcescens, Citrobacter sp., and Klebsiella pneumonia Pseudochrobactrum glaciale, Providencia rettgeri, and Pantoea sp. Cupriavidus basilensis

COD (34.7%) COD (83%), BOD (74%), color (85%)

References Blair and Davis (1980) Gupta et al. (2001) Chauhan and Thakur (2002) Keharia and Madamwar (2003) Raj et al. (2007)

Chen et al. (2012) Chandra et al. (2011)

Color (96.01%), COD (91%), BOD (92.59%)

Chandra and Singh (2012)

Kraft lignin (44.4%)

Shi et al. (2013)

degrade the low-molecular-weight components produced from lignin degradation by fungi. Therefore, bacteria can work in synergy with fungi for complete removal of lignin from paper and pulp mill effluent.

Role of Fungi Among the biological agents, fungi are unique as their extracellular enzyme system such as, lignin peroxidase (LiP), laccase (Lac), and manganese peroxidase (MnP) can depolymerize lignin non-specifically. A number of fungal species have been identified as major lignin degrader. Compared to bacteria, fungi can survive the higher effluent load (Kamali and Khodaparast 2015). Lignin degradation by fungi is essentially a secondary metabolic process, as fungi do not utilize lignin as a carbon source for their growth. This unique feature makes fungi suitable for their application in pulp pretreatment, which can reduce the energy requirement during the mechanical pulping and also will increase the efficiency of bioconversion (Skyba et al. 2013). Some important species of fungi found to degrade lignin include Schizophyllum commune, Tinctoria borbonica, Phanerochaete chrysosporium, Trametes versicolor, Aspergillus niger and Trichoderma sp., Lentinus edodes,

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Datronia sp., Thelephora sp., Coriolus versicolor, Fomes lividus, Pleurotus sp., and Ceriporiopsis sp. due to their advanced extracellular enzyme system (Chedchant et al. 2009; Dashtban et al. 2010; Paliwal et al. 2012; Kamali and Khodaparast 2015). Apiwattanapiwat et al. (2006) reported the decolorization efficiency of Datronia sp. KAPI0039 (54.9%) and Trichaptum sp. KAPI0025 (54.4%) for paper and pulp mill effluent. Yadav et al. (2010) treated the kraft pulp of mixed hardwood with lignin-degrading fungi Ceriporiopsis subvermispora during the bleaching pretreatment. They observed that the fungal treatment made the bleaching process energy efficient and reduced the chlorine consumption up to 4.8%, lignin content 4.7%, and pollution load in terms of COD and BOD by 32.6% and 41.5%, respectively.

Role of Microbial Enzymes Microbial (fungal and bacterial) enzyme systems can serve as an important tool for the degradation of lignin as well as to convert different components of black liquor into useful intermediate metabolites during the treatment of paper and pulp mill effluents. Some fungi such as white-rot fungi utilize a mixture of extracellular ligninolytic enzymes, organic acids, mediators, and accessory enzymes for the depolymerization of lignin (Fig. 2). These enzymes include laccase (Lac), lignin peroxidase (LiP), manganese peroxidase (MnP), and versatile peroxidase (VP). Apart from the main lignin-degrading enzyme, there are some accessory enzymes that facilitate the activity of principal enzymes. These include glyoxal oxidase (GLOX), aryl-alcohol oxidase (AAO), pyranose-2 oxidase (two flavoenzymes), aryl-alcohol dehydrogenase (AAD), glucose oxidase, and methanol oxidase. Cellulolytic enzyme, i.e., cellobiose dehydrogenase (CDH), produced by many fungi is also involved in degradation of lignin. However, the effect of CDH on lignin degradation is through the reduction of quinones, which can be used by ligninolytic enzymes or the support of a MnP reaction (Fig. 2) (Dashtban et al. 2010). Different ligninolytic and accessory enzymes are categorized on the basis of their cofactor, mediator, and mode of action (Table 4). The information on bacterial enzymes for lignin degradation is limited compared to fungal enzymology. However, bacteria utilize some unique enzymes along with principal extracellular enzymes for the lignin degradation. These include bacterial laccase, glutathione S-transferases, ring-cleaving dioxygenases, monooxygenases, and phenol oxidases (Masai et al. 2003; Masai et al. 2007; Allocati et al. 2009). Bacterial enzymes have an advantage over their fungal counterparts with regard to specificity, thermostability, halotolerance, and activity on a wide range of environmental conditions. Application of lignin-degrading enzymes in pre-bleaching of pulp could prove to be effective and also reduces the pollution that results from the chemical bleaching (Yadav et al. 2010). Jain et al. (2007) treated the pulp with xylanase enzyme that effectively reduced the chlorine demand by 15–18% followed by conventional bleaching sequence, i.e., chlorination, extraction, and hypochlorite (CEH), and also the brightness was increased by 2–3% (ISO, International

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Lignin degradation process (mainly by basidiomycetes white-rot fungi) Enzymatic lignocellulose degradation Phenol oxidases

Lac

LiP

MnP h

a

e

Free radicals

VP

i

Intermediate radicals

j

Quinone redox cycling

Glycopeptides-catalyzed Fenton reaction

Quinones

Fenton reaction

Hydroquinones

AAO & GLOX

m Fenton reaction Reduction of quinones

H2O2

n

k

CDH catalyzed reaction

H2O2-GO AAD, QR

l

Chelated MN(III)

f

b

Accessory enzymes

Mn(II) Mn(III)

H2O2

O2 2H2O

Heme peroxidases

Non-enzymatic lignocellulose degradation

o

o

o

•OH

Second mediators

LMS c

Reactive radicals

Mediators

d

More lignocellulolytic enzymes (cellulases & hemicellulases)

g

Lignin Cellulose Hemicellulose

Lignin degradation

Lignocellulose degradation

Fig. 2 Schematic diagram of lignin degradation by white-rot fungi; steps of enzymatic reactions involved in depolymerization of lignin are as follows: (a) Substrate oxidation leading to the formation of free radicals. (b) Free radicals act as intermediate substrates for the enzyme laccasemediator system (LMS). (c) Mediator formation creates the nonenzymatic routes of oxidative polymerization/depolymerization of lignin. (d) LMS involve in ligninolysis. (e) Lignin peroxidases (LiPs) oxidized the substrate result in formation of intermediate radicals (phenoxy radicals and veratryl alcohol radical cations). (f) Intermediate radicals undergo nonenzymatic reactions such as polymerization, intermolecular rearrangements, etc. (g) Direct oxidation of non-phenolic aromatic substrates by LiPs. (h) Manganese peroxidase (MnP) catalyzes the peroxide-dependent oxidation of Mn (II) (as the reducing substrate) to Mn (III). (i) Which is then chelated by oxalate or other chelators? (j) Chelated Mn(III) complex acts as a reactive low-molecular-weight, diffusible redox mediator for oxidation of phenolic substrate. (k) Formation of reactive radicals in the presence of a second mediator (such as acetic acid radicals, peroxyl radicals, superoxide, and formate radicals) for oxidation of non-phenolic units of lignin. (l) Aryl-alcohol oxidase (AAO) and glyoxal oxidase (GLOX) generate hydrogen peroxide (H2O2) required by peroxidases. (m) Aryl-alcohol dehydrogenases (AAD) and quinone reductases (QR) reduce the lignin-derived compounds. (n) Cellobiose dehydrogenase (CDH) involved in lignin degradation in the presence of H2O2 and chelated Fe ions. (o) The nonenzymatic reactions are involved in lignin degradation through generation-free hydroxyl radicals (˙OH) by CDH-catalyzed reactions, low-molecular-weight peptides/quinine redox cycling and glycopeptide-catalyzed Fenton reactions (Source: Dashtban et al. 2010)

Organization for Standardization). Reduction in AOX, BOD, and COD further proves the applicability of enzymatic treatment of bleached effluents (Jain et al. 2007). However, the introduction of enzymatic pre-bleaching technology on the commercial scale requires the optimization of environmental conditions such as pH, temperature, enzyme activity, etc. as these conditions affect the activity of enzymes.

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Table 4 Enzymes involved in the degradation of lignin and their main reactions Enzyme activity (Abbreviation) Lignin peroxidase (LiP) Manganese peroxidase (MnP) Versatile peroxidase (VP) (hybrid peroxidases) Laccase (LAC)

Cofactor H2O2

Substrate, “mediator” Veratryl alcohol

H2O2

Mn, organic acid as chelator, thiols, unsaturated lipids Same or similar compounds as LiP and MnP

H2O2

O2

Glyoxal oxidase (GLOX) Aryl-alcohol oxidase (AAO) Other H2O2producing enzymes

Mediators (hydroxybenzotriazole, ABTS) Glyoxal, methylglyoxal Aromatic alcohols (anisyl, veratryl alcohol) Many organic compounds

Main effect or reaction Aromatic ring oxidized to cation radical Mn2+ oxidized to Mn3+, further oxidation of phenolic compounds to phenoxyl radicals Same effect on aromatic and phenolic compounds as LiP and MnP Phenol is oxidized to phenoxyl radicals, mediator radicals Glyoxal oxidized to glyoxylic acid, H2O2 production Aromatic alcohol oxidized to aldehydes, H2O2 production O2 reduced to H2O2

Source: Hatakka (2001)

Technologies for Enhancing the Biological Treatment Processes Industrial application of microbial cells and/or their enzymes for effluent treatment is limited by several factors such as loss of active biomass and their enzymes, incompatibility with the native microbes, etc. Therefore, it is crucial to develop advanced techniques for maintaining the activity of applied microbes in large-scale industrial effluent treatment processes through the bioreactor development, immobilization, and nano-biotechnology.

Bioreactors An important strategy for the treatment of large-scale effluent generated from industrial processes is to develop bioreactors to deal with specific environmental pollutants. Bioreactors are the large reactor vessels or tanks used in effluent treatments and also to transform waste materials into biochemical products. A bioreactor may contain inoculation of competent microorganisms suspended or immobilized as biofilms in the reaction medium. It is designed to provide the environment to the microbial cells or enzymes for product formation. Bioreactor may be either batch, fed batch, or continuous type. The design may be aerobic, anaerobic, and microaerobic conditions. In aerobic process, the method of providing oxygen has resulted

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in mechanically agitation bioreactors, airlift columns, bubbler column, and membrane reactors. The sterility requirements of pure culture process, which develop microbial strains, differ from those of environmental mixed culture processes, which are based on natural selection (Erickson 1992). Microbial growth in the bioreactor with process optimization ascertains the utilization of components in the effluents as carbon and energy source. Several reports are available on the principles and design of bioreactor (Godjevargova et al. 2003). A sequential bioreactor has been used for the treatment of pulp and paper mill effluent and proved effective in absorbing the shock load to microorganisms during its continuous presence in toxic compounds (Valenzuela et al. 1997; Thakur 2004). The advantage of using two-step sequential treatments would be to use fungi for removal of coloring material in the first step in the reactor and subsequent treatment of effluent by bacteria for effective removal of the recalcitrant organic compounds. The idea of sequential treatment with previously studied microorganisms appears to be a better approach, as it may allow a greater removal of color and lignin, COD using different removal mechanisms after providing them optimum growth conditions (Thakur 2004). Singh and Thakur (2006) reported that sequential anaerobic and aerobic treatment is more efficient in removal of color and chlorinated compounds, because anaerobic microorganisms were able to remove highly chlorinated substances more efficiently than aerobic microorganisms. The combined treatments typically removed 82% of the AOX, COD, and chlorinated phenolics and completely eliminated chlorate. Lafond and Ferguson (1991) reported that anaerobic treatment in an upflow hybrid reactor removed 17–40% of AOX. Similarly, Mishra et al. (2016) compared the effluent treatment efficiency of hybrid unit of upflow fixed-bed anaerobic bioreactor (UFBAB) along with slow sand filter (SSF) with the single-unit UFBAB for paper and pulp mill effluent. The hybrid system showed better treatment efficiency as the SSF provides a polishing effect to the effluent generated by after the UFBAB treatment. Shim and Kawamoto (2002) investigated the production of lignin peroxidase by a white-rot fungus, Phanerochaete chrysosporium, in batch and a reactor system with various carriers. Immobilization of mycelia cell culture was more effective in promoting cell growth and lignin peroxidase production compared to conventional stationary liquid culture. Biostage carrier, commonly used for biochemical treatment in a fluidized bed disposal system, greatly improved production of lignin peroxidase up to 8.1 U mL1 in the batch system. The packed bed reactor system was operated using a repeated batch technique, consisting of alternating growth and production phases, to sustain lignin peroxidase growth and production during the entire experiment period. Steady-state continuous PCP degradation over an extended period was accomplished with a mineralization ratio exceeding 80%.

Cells/Enzyme Immobilization Selection and application of appropriate biological systems for paper mill effluent treatment is very important. However, application of suspended cells and enzymes is not technically viable for continuous operation of a treatment system due to washout

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Table 5 Advantages and disadvantages associated with the application of immobilized cells/enzymes for effluent treatment Advantages Enhanced retention period of biomass and biocatalyst Increasing process stability Reusability of biological system Convenience in continuous operation of system and tolerance to shock loadings Allowing higher hydraulic loading rates Generating relatively less biological sludge Maintaining the activities of slow-growing microbes Allow development of multienzyme activity

Disadvantages Loss or reduction of the catalytic activity of some cells or enzymes Diffusion limitation for enzymes Additional cost for isolation, purification, and recovery of active biocatalysts (enzymes)

of cells and/or their enzymes. Application of immobilized whole cells or biocatalyst (enzymes) therefore can overcome the problems associated with the long-term utilization of suspended cells or enzyme for the treatment of industrial effluent. Therefore, immobilization is the confinement of the cells or any biological agent to a defined region (support/matrix) or resisting their movement in a space while retaining their enzymes’ activity in the medium. Immobilized systems can offer several advantages and improve the efficiency of an effluent treatment system (Chen et al. 2005; Engade and Gupta 2010). Some advantages and disadvantages of application of immobilization techniques wastewater treatment are summarized in Table 5. A variety of support/matrices can be used for immobilization of cell or enzymes, which are grouped into three groups such as natural polymers, synthetic polymers, and inorganic materials (Fig. 3). Synthetic polymers are ion exchange resins/polymeric and insoluble supports with porous surface, which trap and hold the cells and enzymes. In last few decades, various techniques have been developed for the immobilization of cells and enzymes as well. These biological agents can be immobilized to different matrices by various physical and chemical methods (Fig. 4). In the recent years, application of immobilized cells and enzymes for wastewater treatment has been reported by many workers. Malaviya and Rathore (2007) reported the bioremediation of pulp and paper mill effluent by an immobilized fungal consortium. The treatment resulted in the reduction of color, lignin, and COD of the effluent in the order of 78.6%, 79.0%, and 89.4%, respectively, in 4 days. Ortega-Clemente et al. (2009) also reported the posttreatment of pulp mill effluents using an aerobic, upflow column reactor packed with immobilized Trametes versicolor. Arica et al. (2009) used immobilization of laccase onto nonporous poly (GMA/EGDMA) beads for degradation of industrial effluent. In addition, Sharma et al. (2008) immobilized the enzyme tannase (E.C.3.1.1.20) to possess desirable properties such as stability at extreme pH and temperature, board substrate specificity for industrial applications. Couto et al. (2004) used stainless steel sponge as carrier materials to immobilize white-rot fungus for textile dye decolorization. Wu et al. (2005) conducted an investigation to explore the lignin-degrading capacity of attached-growth white-rot fungi. Different white-rot fungi, Phanerochaete chrysosporium, Pleurotus

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Matrix/Supports used for Immobilization

Natural Polymers

Agar, Carrageenan, Calcium alginate, Chitosan and Chitin, Collagen, Gelatin, Cellulose, Starch, Pectin

Synthetic Polymers

Polyacrylamide, Polyurethane, Polyvinyl Chloride (PVC), DEAE Cellulose, UV active polyethylene glycol

Inorganic Material

Zeolite, Ceramics, Diatomaceous Earth, Silica, Glass, Active Carbon. Charcoal

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Fig. 3 Different matrix/supports used for cell/enzyme immobilization

ostreatus, Lentinus edodes, and Trametes versicolor, immobilized on porous plastic media individually and were further assessed to treat black liquor. Vikineswary et al. (2006) used sago hampas for laccase production by Pycnoporus sanguineus in solidstate fermentation. Corncob has also been used by several researchers as a SSF substrate for enhanced enzyme production (Cabaleiro et al. 2002; Oliveira et al. 2006). These studies presented a new approach of utilizing the agro-industrial waste as fermentation feedstock for the production of value-added enzymes.

Nano-biotechnology Researches in the field of nano-biotechnology have received increased attention in the last few decades. Nanomaterials due to their nanoscale size can serve as excellent tools with improved structural and functional properties for their application in treatment of industrial effluents. The interest in nano-biotechnology has been increasing in analytical practices and for developing efficient biosensors through controlled immobilization of biomolecules (enzymes) on different nanosized surfaces. Lignin-degrading enzyme such as laccases as biosensor has been studied in the detection of industrial pollutants (Couto and Herrera 2006). In order to develop multifunctional biosensor, Roy et al. (2005) reported the application of cross-linked enzyme crystals (CLEC) of laccase isolated from Trametes versicolor as biosensor that has great advantage over the soluble enzyme. Several researches have been made for enzymes immobilization on nonmagnetic micron-scaled natural support material such as chitosan to achieve efficient performance and reusability of the enzyme. However, removal of immobilized biocatalyst

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Cell/Enzyme Immobilization Methods

Chemical Methods

Physical Methods

Entrapment (Physical trapping of cells/ enzymes in the polymeric network of natural or synthetic polymer)

Adsorption (Involves the physical binding of biocatalyst on the surface of carrier matrix)

Microencapsulation (Enclosing the cells/enzymes in a semi-permeable membrance capsule)

Covalent Attachment (Involves the formation of covalent bond between the biocatalyst and the support)

Cross-linking (Covalent bonds formation between the a three the enzymes via poly-functional reagents forming three dimensional crossed linked aggregates)

Conjugation by Affinity Ligands (Involves the covalently binding of closely affinity ligands (e.g., antibody or lectin)to the matrix)

Fig. 4 Methods for immobilization

from the reaction media is mostly achieved by filtration or centrifugation methods that may cause loss of activity of the enzymes due to high mechanical stress (Kalkan et al. 2012). In the recent years, researches are being focused on development and application of cell or enzyme immobilization on magnetic support materials functionalized with natural and synthetic polymers. Application of magnetic support nano-materials for enzyme immobilization is easily separable from the reaction mixture and also advantageous for maintaining the catalytic activity of enzymes and improving the storage stability. Jiang et al. (2005) observed the improved thermal, operational, and storage stabilities of the enzyme laccase after immobilization on magnetic chitosan microspheres by adsorption and cross-linking. The optimum conditions (pH 3.0 and temperature 10  C and 55  C) for maximum catalytic activity of immobilized laccase was also observed. Pich et al. (2006) have compared the activity of immobilized laccase on three different carrier particles of poly(styrene-co-acetoacetoxyethyl methacrylate) (PS-AAEM). In polymeric particles

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Fig. 5 Laccase immobilization onto Fe3O4-CS nanoparticles (a) by EDAC and (b) by CC (Source: Kalkan et al. 2012)

PS-AAEM, the PS was a polystyrene core bearing reactive β-diketone groups (AAEM) on the particle surface that act as binding sites for the enzymes. The other two particle carriers PS-AAEM-DM and PS-AAEM-EM additionally contain maghemite nanoparticles, fixed on the surface and in the core, respectively. They observed the similar activity and immobilization efficiency in all the three carriers. However, the immobilization of laccase over the two hybrids PS-AAEM-DM and PS-AAEM-EM has shown to exhibit better storage stability and increased pH and temperature resistance of enzyme, compared to PS-AAEM. Similarly, Kalkan et al. (2012) immobilized laccase (L) enzyme on magnetite (Fe3O4) nanoparticles coated and functionalized with chitosan (CS). Laccase was immobilized on (Fe3O4-CS) by adsorption or covalent binding after activating the hydroxyl groups of chitosan with carbodiimide (EDAC) or cyanuric chloride (CC) (Fig. 5). The results showed that all immobilized systems retained more than 71% of their initial activity at the end of 30 batch uses.

Constrains and Future Thrust in Biological Treatment System The effluent generated from paper and pulp industries has tremendous impact on the health of ecosystem functioning. Microorganisms (fungi and bacteria) and their enzyme system have great potential for their application in the treatment of paper and pulp mill effluent generated from different processes. The broad spectrum for biodegradation potential of bacteria is due to adaptability and versatility of their enzymes for a wide range of environmental conditions and for the fungi is because of extracellular and non-specific nature of their enzyme system. The application of

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such versatile biological agents for the treatment of industrial wastewater on large scale is still limited due to several factors, such as limited amount and sources of biocatalysts, lack of optimum substrate specificities, environmental and cultural conditions required for the growth of microorganisms, competition from native microbes, lack of efficient microbial expression, effluent load, appropriate treatment reactors vessels, etc., that make the biological treatment processes slow compared to the conventional processes. However, the activities of the bacterial and fungal enzymes system can be enhanced by the utilization of innovative advanced techniques such as cell/enzymes immobilization, nanotechnology, etc. The biodegradation efficiency can further be enhanced by exploring the microbial expression for specific enzyme. Furthermore, isolation, characterization of new microbial strains, immobilization, and genetics of lignin-degrading microorganisms are the area of future research required to make the direct use of biological agents in industrial treatment processes.

Conclusion Paper and pulp mill is extremely water intensive and polluting sector. Conventional treatment techniques are available but small-scale mills hesitate to install them due to their being extremely cost demanding. Biological treatments are effective in removing the major pollution load such as BOD and COD from the effluent and easy to install and cost-effective. However, biological treatment techniques are comparatively slow, and available natural enzymes sources (microorganisms) cannot meet the market demand due to low yields and their incompatibility toward standard industrial processes. The large amount of enzymes required for the effluent treatment and presence of all the enzymes in a single microorganism is difficult, thus becoming a bottleneck for industrial application of microbes and their enzyme system in paper and pulp mill effluent treatment. However, application of two or more microbes in combinations can solve the problem of limited resource availability of enzymes. Therefore, to achieve the desired standard norms for effluent, successful implementation of microbes in biological treatment processes of paper and pulp mill effluent requires the identification of optimum application conditions such as pH, temperature, substrate specificities, reaction media, etc. The treatment efficiency of biological agents (fungi, bacteria, and their enzymes) for industrial effluent can be further enhanced by utilizing the modern cell and enzyme immobilization and nano-biotechnological techniques.

Cross-References ▶ Biofilm-Based Systems for Industrial Wastewater Treatment ▶ Environmental Nanotechnology ▶ Environmental Treatment Technologies: Adsorption ▶ Technologies for Treatment of Colored Wastewater from Different Industries ▶ Wastewater Management to Environmental Materials Management

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Meryem Asri, Soumya Elabed, Saad Ibnsouda Koraichi, and Naïma El Ghachtouli

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Wastewater Characteristics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biofilm Fundamentals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Potentiality of Biofilm in Wastewater Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biofilm Application in Wastewater Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Supports in Biofilm-Based Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Bioreactors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cross-References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Over the last few decades, the worldwide growth in industrial activity has resulted in the discharge of significant pollutant quantities in the aquatic environment. These contaminants are generally characterized by their toxicity for living organisms and the environment. In response to these dangers, environmental regulations are imposing limitations on a wide variety of pollutants within industrial wastewaters. The great diversity of raw materials and production operations

M. Asri · S. Elabed · N. El Ghachtouli (*) Laboratoire de Biotechnologie Microbienne, Faculté des Sciences et Techniques, Université Sidi Mohamed Ben Abdellah, Fes, Morocco e-mail: [email protected] S. Ibnsouda Koraichi Laboratoire de Biotechnologie Microbienne, Faculté des Sciences et Techniques, Université Sidi Mohamed Ben Abdellah, Fes, Morocco Centre Universitaire Régional d’Interface, Université Sidi Mohamed Ben Abdellah, Fes, Morocco © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_137

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employed by industries poses a matter of concern for scientific community attempting to define proficient control technologies. Biological treatment technologies were viewed as attractive alternatives to conventional methods. Indeed, biofilm-mediated processes for industrial wastewater treatment are among the most proficient technologies due to their distinct advantages over conventional methods. Biofilm mode proved its capacity to enhance the overall pollutant degradation efficiency. Expanding the knowledge regarding biofilm wastewater treatment would contribute to its full-scale use. This chapter presents an overview of the beneficial use of biofilm mode in depollution technologies and a critical discussion of a relevant number of recent investigations that have dealt with biofilm-based processes. It also provides a review of the various types of bioreactors and biofilm supports currently used. Keywords

Industrial effluents · Wastewater treatment · Biofilm · Supports · Bioreactors

Introduction The rescue of industrial wastewaters has become a significant problem throughout the world (Segura et al. 2015). These wastewaters generally contain high quantities of hazardous pollutants such as persistent organic pollutants, heavy metals, phenols, dyes, pesticides, humic substances, and detergents (Ferronato et al. 2016). These contaminants are mostly characterized by their toxicity for living organisms, their persistence against chemical or biological decomposition, their high environmental mobility, and their extreme bioaccumulation tendency in the food chain (Bahafid et al. 2013; Liu et al. 2013b). In response to these dangers, environmental regulations are imposing limitations in a wide variety of pollutants within industrial wastewaters (Day 1993). In order to remove or transform undesirable elements from industrial effluents, various physicochemical treatment processes such as electrochemical, coagulation–flocculation, adsorption, and membrane treatment are currently employed depending on the wastewater characteristics (Shammas 2005; Leyva-Ramos et al. 2008). Biological treatment technologies were viewed as an appealing approach compared to these conventional methods. Among these, biosorption defined as the capacity of biological materials to accumulate and/or transform pollutants from contaminated sites had gained especial attention (Fomina and Gadd 2014). This technology can use different materials such as plant biomass or animal polymers. Nevertheless, biodepollution plants mainly employ microorganisms. Microorganisms can be employed in different forms within wastewater treatment systems, as living or dead (Bahafid et al. 2013), suspended or biofilm-immobilized biomass (Pan et al. 2014), etc. Biofilm is a growth mode distinct from planktonic form of microorganisms. It can be defined as a complex structure of cells and extracellular products that can be formed spontaneously as dense granules

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(Lettinga et al. 1980). It can also be fixed on a solid surface or suspended carrier supported on particles (Davey and O’toole 2000). This microbial mode of development is of great interest in biotechnology applications. Bioreactors based on fixed microorganisms in biofilm growth mode are increasingly employed for bioremediation applications (Lewandowski and Boltz 2011). They represent a recent proficient alternative for assessing environmental effects. This is by means of their significant capacity to resist to stressful conditions and their various catabolic pathways allowing the degradation of various hazardous contaminants. The extracellular products of biofilms are of great importance. They form a barrier for microorganisms against the toxicity of wastewaters and thus enhance the pollutant immobilization (Quintelas et al. 2011). This review summarizes existing knowledge on various aspects of the implication of biofilm in wastewater treatment. It (i) provides essential insight into the benefits of biofilms in wastewater treatment technologies, (ii) presents an overview on the supports currently used in these biofilm-based technologies, and (iii) gives a rundown of the commonly employed biofilm-mediated processes. This chapter provides a source of useful information for future areas of research and efforts, aiming to fully implement these methods for wastewater treatment.

Wastewater Characteristics Wastewater treatment process should be decided based on a balance between technical and economical aspects and customized according to the wastewater characteristics. Industrial wastewaters are commonly characterized by the basic parameters which are represented by chemical oxygen demand (COD), biochemical oxygen demand (BOD), suspended solids (SS), ammonium nitrogen, heavy metal concentration, pH, turbidity, color, and biological parameters. These characteristics vary in a wide range depending on the industrial activity. Industrial processes constituting the major sources of polluted effluents include petroleum, textile, tannery, food processing, pharmaceutical, and manufacturing industries. Typical characteristics of industrial wastewaters are summarized in many previous reviews (Lin et al. 2012). It offers a useful guideline for the decision of the treatment process despite the variation of wastewater characteristics. The wastewater from industries varies so greatly in both flow and pollution strength. Thus, it is impossible to give fixed values to their composition. Conventionally, industrial wastewaters may contain suspended, colloidal, and dissolved (mineral and organic) solids. In addition, it may contain colored matter at varied concentrations. It may also contain inert, organic, or toxic materials and possibly pathogenic microorganisms. Compared with municipal wastewaters, industrial effluents generally present a higher organic strength (>1000 mg COD/L) and an extreme physicochemical nature with extreme acid or alkaline pH values, high salinity (e.g., from petroleum refining, textile processing, leather processing), and a variable temperature. It may also contain high concentrations of toxic substances

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(natural or synthetic) which may present an obstacle to biological treatment plants of the effluents (Lin et al. 2012). The pH variation and salinity were always the serious challenges of biological treatment. These conditions lead to the inhibition of both microbial cell activity and the flocculation of sludge flocs (Lefebvre and Moletta 2006). In the case of industrial wastewater with high concentration of toxic compounds (heavy metals, phenols, surfactants, pesticides, etc.) mainly characterized by marked persistence against chemical or biological degradation, a pre- or posttreatment should be carried using performing microorganisms able to remove the pollutant of concern or in some cases the biological treatment should be combined to other physic-chemical treatment strategy for an optimal treatment efficiency (Lin et al. 2012). With an appropriate analysis and control of the effluent, generally wastewaters with a BOD/COD ration of 0.5 or greater may be treated with a biological treatment system (Metcalf and Eddy 2003). Compared to other wastewater treatment strategies, biological treatment strategies present the advantages of being eco-friendly and economic. It can be assessed by either aerobic or anaerobic processes. Aerobic processes refer to the use of dissolved oxygen by microbial cells in the conversion of the organic matter to biomass and CO2, while anaerobic process involves the degradation of organic wastes into methane, CO2, and H2O in the absence of oxygen. Each of the treatments presents various merits. In fact, both systems are able to remove effectively organic pollution. However, generally, aerobic systems are preferably dedicated to the treatment of wastewaters containing biodegradable COD concentrations less than 1000 mg/L. A crossover point ranging from 300 to 700 mg/L influent wastewater ultimate BOD (BODu) was determined by Cakir and Stenstrom (2005) as a crucial interval for an effective treatment of wastewaters using aerobic systems, while anaerobic systems become favorable for the treatment of higher strength wastewater (COD concentration over 4000 mg/L) (Chan et al. 2009).

Biofilm Fundamentals Since the first paper, where the term “biofilm” was coined and described by Costerton et al. (1978), great efforts have been made to deeply and properly understand this complex structure. Thus, many definitions were attributed to the term biofilm in the literature. It is commonly defined as being a complex coherent structure of cells (aggregates) and cellular products (Nicolella 2000). Most definitions include the attachment of microorganisms to a solid surface or carriers (Quintelas et al. 2008; Yang et al. 2015); though few definitions do not consider the surface as an essential element in the biofilm mode, it can be formed spontaneously as granules (Lettinga et al. 1980). Biofilms are heterogeneous and enormously complex communities of microbial cells suspended in a matrix of extracellular polymeric substances (EPS) (Quintelas et al. 2011). They are highly affected by the surrounding environmental and mechanical conditions. According to these latter, the biofilm may use its ability to regulate many activities such as attachment,

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mobility, and detachment through the cell-to-cell molecular signaling, called “quorum sensing” (Singh et al. 2006). Biofilm is a predominant microbial growth mode in the environment distinct from planktonic growth mechanism. It is commonly found in soil and aquatic systems. It has been proved to be an advantageous form offering protection against different dangers including predation, and chemical and biological toxicants. It also offers resistance to stressing conditions such as dehydration and lack of nutrients (Singh et al. 2006). It is hence increasingly employed in wastewater treatment systems (Butler and Boltz 2014). Nevertheless, these microbial aggregates are with relevance to serious problems in diverse fields, essentially in terms of material corrosion and degradation (Hamadi et al. 2005).

Potentiality of Biofilm in Wastewater Treatment The use of biofilm-based technology in wastewater treatment dates from 1893 with the first utilization of trickling filters in England (Lohmeyer 1957). Although microbial biofilms cause detrimental effects in various environments, they are still considered as useful in biodegradation of complex pollutants. Biofilm-mediated depollution presents an efficient and cost-effective option rather than the use of planktonic microbial cells. This is due to the better survival and adaptation ability to stressing conditions. Indeed, biofilm-forming microorganisms are good competitors with nutrients and oxygen. They are also known to survive and manage the most stressful conditions and harsh hydrodynamics forces. It makes them excellent candidates for bioremediation applications. In fact, depending on the environmental conditions, the biofilm via the EPS matrix can develop different structures to overcome environmental challenges (Kreft and Wimpenny 2001; Miqueleto et al. 2010; Jung et al. 2013). Thus, it forms mushroomlike shapes in the case of static water and appears in filamentous structure in fast-moving water flux (Edwards et al. 2000; Reysenbach and Cady 2001). In fact, in recent years, biofilm-based processes are being increasingly employed as an appealing strategy representing an environment-friendly and cost-effective option for pollutant removal (Das et al. 2012). The degradation of xenobiotics is more effective within biofilm systems, owing to the close, mutually beneficial physical and physiological interactions among biofilm-forming cells. The use of the biofilm-mediated bioremediation plants became hence common for the removal of pollutants (Das et al. 2012). The beneficial use of biofilm processes for the adsorption, immobilization, and degradation of various pollutants was pointed out (Quintelas et al. 2010). In fact, these technologies have been widely used for the removal of both organic and inorganic compounds from aqueous media (Chen et al. 2008; Quintelas et al. 2013; Hai et al. 2015). Many contaminants including heavy metals, petroleum, dyes, and pesticides have been successfully remediated using microbial biofilms (Mitra and Mukhopadhyay 2016). In order to resist predatory protozoa, in biofilm mode, microbial cells are immobilized in a self-synthesized matrix which offers protection from stress,

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contaminants, and predatory protozoa (Quintelas et al. 2011). They also tend to form large inedible microcolonies that offers a better resistance to protozoa under harsh environment (Matz and Kjelleberg 2005; Mitra and Mukhopadhyay 2016). Furthermore, the gene expression within biofilm mode is distinct from planktonic form of microbial cells. The heterogeneous assemblage of microorganisms in the biofilm in terms of microbial species offers a diversity of metabolic pathways. It enables the biofilm to degrade several types of pollutants either individually or correctively (Gieg et al. 2013; Horemans et al. 2013). Various microbial species were successfully employed as biofilm form for the wastewater treatment, including bacteria (Abzazou et al. 2016), yeast (Cong et al. 2014), fungi (Badia-Fabregat et al. 2017), and algae (Hoh et al. 2016). The introduction of algae into the biofilm reactors is considered by some researchers as a nuisance due to the clogging issues, while it is viewed by others as a source of great amounts of oxygen beneficial to the growth of the other species constituting the biofilm (Kesaano and Sims 2014). Biofilm processes favor selective development of slow-growing microorganisms such as autotrophs (i.e., nitrogen-oxidizing bacteria) and phosphorus-accumulating microorganisms by the maintenance of high biomass age, which reduces their washout from the system (Lee et al. 2006). Since the researchers became aware of the importance of biofilm in treatment systems, the growth conditions were thoroughly studied for their great influence on the system efficiency. This included temperature, nutrients, substrata, extracellular polymeric substances (EPS), species interactions, and light for algal based systems (Kesaano and Sims 2014). These studies are of extreme importance for the design and scale-up of effective treatment systems. To successfully achieve the treatment of wastewater using biofilm-based systems, optimal conditions must be provided to the microbial consortia presenting the ability to remove or transform the pollutant of concern.

Biofilm Application in Wastewater Treatment Supports in Biofilm-Based Processes In biofilm reactors, microorganisms may be supported on various materials. The depollution efficiency is strongly related to the properties and the nature of these supports and the ability of the biofilm to be attached to the chosen support (Asri et al. 2017). Indeed, the commonly used supports are of various natures including ceramic, clays including zeolithes, seashell and charcoal, plastic materials, sintered glass, fire bricks, sand, natural stones like limestone and gravel, pumice, and rocky aggregates (Silva et al. 2008; Tarjányi-Szikora et al. 2013). These supports were used directly without any modification or after treatment to modify certain properties (porosity, surface charges, etc.) in order to enhance their effectiveness, while others can be commercially synthesized. For microbial fuel cells (MFCs), the used material

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should present not only an adhesive property but also an electrical conductivity and a chemical stability. The commonly used supports in this technology are carbon-based materials such as carbon cloth, carbon paper, or carbon felt (Liu et al. 2013a; Zhang et al. 2013; Alatraktchi et al. 2014). The effective depollution of the biofilm systems was related to support properties mainly the surface area, surface morphology, and capacity of microbial adhesion. Numerous works mainly focused on microbial adhesion studies aiming the determination of the influence of surface characteristics on biofilm formation. Indeed, this step is viewed as a key element promoting the cell attachment and the long-term stability of the biofilm (Zainul Akmar et al. 2007). Many surface properties were proved to influence microbial initial attachment to various substrata. For instance, roughness is among the most reported parameters (El Abed et al. 2012). In fact, rough or porous surfaces were more favorable for cell adhesion than regular surfaces, which was ascribed to the increased surface area and the protection against hydraulic shear forces (Hoh et al. 2016). Surface energy components were also showed to strongly influence microbial initial attachment and treatment performance of the biofilm-based system (Asri et al. 2017). Among these components, hydrophobicity character was reported to be of extreme importance in wastewater applications. Indeed, a greater adhesion of different microbial species to hydrophobic surfaces such as titanium, Perspex, and stainless steel was showed (Kesaano and Sims 2014). Conversely, Irving and Allen (2011) showed no correlation between surface hydrophobicity of the substratum and cell adhesion of microbial biofilms grown in wastewater. On this basis, there is no recommended standard material as support for biofilm formation in wastewater treatment scale-up operation. However, in order to select adequate materials for biofilm development for successful application of this technology, the support must be ideally chosen. Thus, many factors should be taken into consideration such as its availability, cost, durability, and compatibility to the selected microorganisms from a thermodynamical point of view (Asri et al. 2017). In the majority of the wastewater treatment plants using microbial biofilm, granular activated carbon (GAC) was used as a support material (Quintelas et al. 2010; Muhamad et al. 2013). However, its high cost and the challenges associated to its regeneration have strongly limited its utilization. At present, the works are mostly oriented toward the utilization of cheap alternatives. Hence, works in this regard have been significantly increased. This recent approach for the choice of supports is attempting to use natural wastes as supports; this adsorbent-support category is called low-cost adsorbents especially lignocellulosic wastes and by-products as an economical and eco-friendly alternative to conventional supports (Abdolali et al. 2014). This class includes a large variety of adsorbents such as fruit peel (Babel and Kurniawan 2004; Memon et al. 2009) and wood husk (Zainul Akmar et al. 2007; Asri et al. 2017). Various used support materials within different biofilm-based wastewater treatment systems are illustrated in Table 1.

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Table 1 Bioremediation of different pollutants using biofilms in bioreactors Type of reactor Column lab scale

Microorganism or culture Bacillus coagulans

Pollutant Cr (VI)

Pseudomonas sp., Rhodococcus sp.

Chlorophenol Laboratory and pilot scale (FBR)

Sphingomonas xenophaga QYY

Bromoamine acid

Arthrobacter viscosus

Cr (III) and Cr(VI)

Proteobacteria Bacteroidetes, Nitrospirae, Cyanobacteria Actinobacteria

COD

Ammonia

Geobacter COD metallireducens + bacteria present in the wastewater Desulfovibrio Sulfate desulfuricans Secondary sludge from Toluene wastewater treatment plant

Laboratory scale (MBR) Suspended biofilm

Natural ventilation trickling filters (NVTFs) MFC

MFC

Membrane biofilter reactor Mixed culture of herbicide- MCPP; 2,4-D Biofilm degrading bacteria (herbicides) column reactor

Material support Granular activated carbon (GAC)

Silica-based spherical Celite R-633 (Celite Co.) microcarriers

NaY zeolite

Sponge Zeolite Ceramsite Sponge Zeolite Ceramsite Graphite

Activated carbon cloth Hollow-fiber membrane GAC

Treatment efficiency 5.34 mg.g1 biosorbent for initial concentrations of 100 mg.L1 99.9%

Reference Quintelas et al. (2008)

Puhakka et al. (1995)

90% (color) – 50% (COD)

Qu et al. (2009)

14 mg/gZeolite (Cr III) – 3 mg/ gzeolite for Cr (VI) 37.33% 53.83% 47.87% 84.90% 65.28% 63.77% 80%

Silva et al. (2008)

99%

Zhao et al. (2008) Parvatiyar et al. (1996) Oh and Tuovinen (1994)

84%;

100% (2,4-D); partially (MCPP)

Zhang et al. (2016)

Liu et al. (2004)

Bioreactors Aerobic/Anaerobic Treatment In order to achieve a high degree of treatment efficiency, organic wastewaters are preferably treated in aerobic biological processes. These processes allow the achievement of a higher removal of soluble biodegradable organic material compared to the anaerobic treatment. Furthermore, the produced biomass is easily separated from the aqueous media presenting a better flocculation property which provides an effluent of lower suspended solid concentration and thus higher quality. However, anaerobic treatments overcome the aerobic approach when treating influents with higher COD concentration. Anaerobic treatment is more suitable for the treatment of highly contaminated industrial effluents because of the high COD concentration. It also requires less energy and nutrient recovery with a low sludge

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production. Whereas anaerobic treatment is a promising approach from the resource recovery and utilization point of view, more efforts are still required for a better pollution control (Seghezzo et al. 1998; Chan et al. 2009). Indeed, this treatment suffers from a low growth rate of microbial cells, a and low quality of effluent due to the low flocculation ability and settling rate of produced biomass. In practical applications, anaerobic treatment suffers also from the difficulty of complete stabilization of organic matter because of the high COD level which provides a final effluent containing solubilized organic matter. To overcome the disadvantages of both aerobic and anaerobic biofilm-based reactors, the use of anaerobic-aerobic systems presents several benefits and has been remarkably employed in industrial and municipal wastewater treatment in order to meet the effluent discharge standard (Frostell 1983; Cervantes et al. 2006). The biofilm-based approaches of wastewater treatment are generally characterized by the ease and the safety of handling. In this section of the chapter, the most commonly used biofilm-based bioreactors are presented highlighting the advantages and disadvantages of each treatment system.

Types of Biofilm-Based Bioreactors Biofilm-based wastewater treatment technology has been heralded as a promising cost-effective clean-up technology. Biofilm bioreactors are playing an extremely important role in environmental biotechnology. Despite the fact that many aspects of their design and technical operations remain poorly understood, a variety of these latter are installed worldwide while the researchers are still conducting intensive investigations for a better control of these promising depollution strategies. Biofilm reactors are essentially composed of five compartments, while some additional components may be typical of a type of reactor: (1) influent, referring to the wastewater containing a given concentration of the pollutant of concern; (2) containment structure; (3) biofilm carrier or substratum referring to the used material for the growth and attachment of microbial cells; (4) effluent water collection system; and (5) an aeration or a mixing system for agitation and carrier distribution. In this section, some of the commercially available biofilm-based processes are discussed. This includes membrane biofilm reactors (MBR), moving-bed biofilm reactors (MBBR), fluidized-bed reactors (FBR), trickling filter (TF), and microbial fuel cells (MFCs). Membrane Biofilm Reactors Membrane biofilm reactors (MBR) have been claimed for a long time as a promising biotechnology for pollutant removal and/or recovery from aqueous solutions. It refers essentially to a combination of a biological degradation of waste compounds using a biofilm-based system and physical separation realized by a membrane unit replacing the secondary settler. Presenting distinct advantages compared to other treatment technologies, it serves for the treatment of various industrial and urban wastewaters, proving an effective removal capability of both organic and inorganic matter (Di Fabio et al. 2013). Indeed, it provides an excellent quality of the effluent, reduced sludge production. It also provides a great

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flexibility toward influent variability, high volumetric loading, and good disinfection efficiency. It can be implemented within two different configurations (Lin et al. 2012). The first configuration is called submerged or immersed configuration and the second is the external or sidestream configuration. The submerged configuration came for the first time in 1989 with the idea of Yamamoto et al. (1989) that consists of the direct immersion of the membrane module inside the reactor. The driving force in this configuration is created by a negative pressure on the permeate side or pressurizing the bioreactor. This configuration showed distinct advantages such as the lower energy consumption and thus the lower operating cost and less cleaning procedures. These advantages have encouraged the development of this system. Regarding the external configuration, the membrane is placed outside the bioreactor allowing the recirculation of the mixed liquor. This configuration allows an easy control and membrane replacement. The driving force in this configuration is related to the high cross-flow velocity (CFV) through the bioreactor (Le-Clech et al. 2006; Liao et al. 2006). To date, both MBR configurations have been successfully utilized for various industrial effluents for their several distinct advantages, mainly for the lower energy consumption. Recently, an innovative configuration has appeared consisting of the development of air-lift sidestream MBRs (Lin et al. 2012). The main idea in this configuration is to exploit all advantages of MBRs above mentioned, with the application of the sidestream airlift principle (Chen and Liu 2006; Shariati et al. 2010). This concept showed its great efficiency in the treatment of industrial and municipal wastewater (Futselaar et al. 2007) and many efforts are still made for its development for a better application. Biofilm processes in MBR could be done by the addition of media in moving- or fixed-bed configurations, or aerated membranes in the bioreactor as a support for biofilm growth. Numerous materials may be employed for biofilm support. To date, cord media, RBC media, sponge, plastic media, and GAC are commercially applied in full-scale systems (Ivanovic and Leiknes 2012). This technology has showed its efficiency in the treatment of various pollutants. For instance, it allowed an overall efficiency of degradation over 90% (Chang et al. 2003, 2004). It also showed an excellent removal of both COD and inorganic nitrogen in a further work (Semmens et al. 2003). The membrane biofilm bioreactors have become an option of choice and efficient alternative for the treatment of domestic and industrial effluents. However, its widespread application suffers from major limitation due to the membrane fouling and clogging layers and their consequences in terms of plant maintenance and operating costs (Le-Clech et al. 2006). Its wider application is also limited by the high energy demand related to the air scouring demand and the high price of membranes. Membrane fouling is a common phenomenon in membrane applications, including MBR systems (Ngo et al. 2006). However, the fraction mostly contributing to this problem remains unclear; it may be caused by colloidal and soluble organic content such as biopolymers or EPS, suspended solids, and physical properties (Ivanovic and Leiknes 2012).

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Many works have dealt with these disadvantages by the control membrane fouling mechanisms and finding cheaper membrane materials. It also aimed the optimization of energy consumption and hence made this system more realistic and a reliable alternative to activated sludge processes and other conventional technologies (Ivanovic and Leiknes 2012). Moving-Bed Biofilm Reactors The moving-bed biofilm reactor (MBBR) is a wastewater biofilm-based technology presently implemented in more than 50 countries. It was developed in Norway in the late 1980s and early 1990s (Ødegaard et al. 1994).The MBBR plants were successfully used for municipal and various types of industrial wastewater treatment (Bassin and Dezotti 2018). The principle behind this process is the combination of the best features of activated sludge process and those of biofilter. This type of reactor may be used for aerobic, anoxic, or anaerobic processes. The carriers that serve as housing for biofilm growth in this system move freely in the tank volume. In aerobic case, their movement is caused by the agitation set up by the air, while in anoxic and anaerobic processes the carriers are kept in movement by a mechanical mixing (Ødegaard 2006). MBBRs are continuous-flow reactor units in which the most commonly used biofilm carrier is named K1. They present a cylindrical form constituted of highdensity polyethylene (density 0.95 g/cm3) with a cross on the inside of the cylinder and “fins” on the outside. There is no filling fraction of the bioreactor with these carriers; it can vary from 25 to 70% of the total tank volume. However, it is recommended to be below 70% (corresponding to 350 m2.m3 effective specific area in the case of K1) (Ødegaard 2006). Biofilms primarily grow in the inside of the plastic carriers, protected from external abrasion. The implementation of moving beds rather than fixed ones presents the advantage of minimizing the clogging limitation and the ability to utilize the whole volume of the bioreactor. As for other biofilm-based processes, the transport of substrates is of extreme importance. In this process, the turbulence in the reactor due to the shearing forces assures not only the appropriate diffusion of compounds with the biofilm, but also the low thickness of the formed biofilm. Additionally, the use of this process minimizes or eliminates the need for biomass recirculation, which is a major problem of fixed-bed biofilm process and activated sludge systems. Only the excessive biomass has to be separated from the solution (Ivanovic and Leiknes 2012). In comparison with fixed- or fluidized-bed biofilm reactors, moving-bed biofilm reactor provides a higher available surface for microbial growth and attachment due to the carrier materials. Moreover, it allows an efficient mixing condition inside the reactor, favoring hence the liberation of the biogas and the dispersing of volatile acids throughout the aqueous solution (Karadag et al. 2015). As for other treatment processes, the combination of moving-bed biofilm bioreactors to other treatment techniques was previously proposed. Indeed, the combination of this technology

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with coagulation and flotation process for high-rate secondary treatment was previously reported (Ødegaard 2006). The MBBR has been demonstrated as a well-proven, compact, and robust reactor for wastewater treatment that served for carbon oxidation as well as for nitrification and denitrification goals as single stage or in combined systems (Gilbert et al. 2014; Malovanyy et al. 2015) and operational results were satisfying in both lab scale and larger scales. A noted disadvantage of these systems is that the carrier must be removed in order to benefit the reactor components (Butler and Boltz 2014). Fluidized-Bed Biofilm Reactors Fluidized-bed biofilm reactors (FBBR) are based on the use of small carriers, forming a bed inside a column kept in fluidized movement due to the flowing wastewaters and the bed hence expands. Within this system, a recycle line is used in order to maintain a fixed, vertical hydraulic flow of introduced wastewater. The aeration is typically realized during recycle, where the influent wastewater mixes with the effluent recycled from the top of the bed. The air addition to the recycle stream is possible; however, it was found to cause a turbulence inside the reactor which may cleave the attached biofilm from the carriers. Generally, media particles are distributed in FBBR within an increasing gradient of size from the top to the bottom of the bioreactor. Depending on the degree of particle expansion, the bed is classified as deemed expanded or fluidized. Many biofilm support materials have been typically used within this system, such as silicabased materials (Puhakka et al. 1995) zeolite, and GAC (Kida et al. 1990). However, in order to provide greater specific surface area, which is a key point of this technology, small materials (below 1 mm) have been used at pilot-scale experiments. Once the driving force of this flowing inside the bioreactor exceeds the gravity (i.e., 30–50 m.h1), the small particles become suspended and separated (Lewandowski and Boltz 2011). Fluidized-bed biofilm reactor (FBBR) is widely recognized to present better mass transfer characteristics in comparison with fixed biofilm reactors. It showed its efficiency for tertiary denitrification in the case of the municipal wastewater treatment. For wastewater treatment, FBBR are used for the removal of oxidized contaminants (McCarty et al. 2005). The fluidization of the media particles presents the advantage of maximizing the contact surface between microbial cells and effluents. It increases also the mass transfer and consequently the treatment efficiency. However, a low degree of bed expansion is recommended, as it decreases the flow velocity, consumes less energy, and increases the concentration of effective biomass. The amount of attached microbial cells in FBBR is very important, exhibiting a high microbial diversity. This parameter permits a rapid recovery of the system after the variation of instability conditions (Malaspina et al. 1996; Borja et al. 2004). Nevertheless, it increased volumetric oxygen biomass because of the important biomass concentration.

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Many successful applications of this system have been reported. For instance, Puhakka et al. (1995) used Pseudomonas sp. and Rhodococcus sp. in a laboratoryscale fluidized-bed biofilm reactor for the remediation of chlorophenolcontaminated groundwater. The treatment allowed an appreciable mineralization of chlorophenol efficiency (over 99.9% of 2,3,4,6-tetrachlorophenol, 2,4,6-trichlorophenol, and pentachlorophenol removal efficiency) at chlorophenol loading rates of 1000 mg.L1.d1 and hydraulic retention times of less than 1 h.

Trickling Filter The trickling filter (TF) has been in use for more than 50 years. It is a three-phase biofilm reactor, including an influent recirculation pump station, the TF, and a clarification unit. It is generally composed of (1) an influent water distribution system, through which wastewaters are introduced into the reactor. The distribution may occur either by fixed-nozzle or rotary distributors; (2) a containment structure; (3) a support media; (4) an underdrain system; and (5) a ventilation system. Wastewater treatment using TFs requires a further liquid-solid separation for the elimination of suspended solids as the TF treatment results in the total suspended solid production. This step is typically carried out using circular or rectangular secondary clarifiers. The rotatory distributor is advantageous for the influent distribution; it allows an intermittent wastewater application and an effective substratum wetting. These parameters avoid the odor emission and dry pockets and permit the biofilm to have resting periods serving primarily as a process aeration mechanism (Lewandowski and Boltz 2011). The gradient of temperature between ambient air and air inside the trickling filter may provide a natural ventilation. When there is a difference between both temperatures, the provided dose of the oxygen is not suitable. In this case, the air supply may be achieved by the underdrain system that provides a space below the trickling filter for the collection of treated influent (Grady et al. 2011). TFs have been proved for their capacity of meeting treatment objectives in terms of carbon oxidation and nitrification. The supported biofilm on the filter media uses oxygen in the form of air for the carbon oxidation. TFs are suitable for carbon oxidation and combined carbon oxidation and nitrification when a solid separation is included in the treatment train. Good results of nitrification have been observed with a combined oxidation and nitrification, where the concentration of ammonia was below 3 mg NH4+ N.L1 and the reached BOD concentration is below 10 mg.L1. The ammonia concentration is even lower when the nitrification is the main treatment goal (Metcalf and Eddy 2003). A key element that should be taken into consideration in the design of TFs is the selection of filter media. The most commonly used material in this system was for a long time stones and gravel. However, these materials were restricting the air circulation in the filter and consequently the available oxygen quantity for the growth of microbial biofilm. This problem has limited the quantity of the treated

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wastewater and also reduced the specific surface area for microbial attachment that can accommodate the BOD loading for the reactor. Furthermore, stone bed trickling filters were limited by the clogging of the void spaces when treating high organic loads because of the excessive microbial cell growth. However, rock-media TFs were able to provide great treatment performance under low organic loading (i.e., 1.5 mg/l) in drinking water is harmful to the human health. Although fluoride is an essential trace element for humans, excessive fluoride intake may cause adverse health effects on humans, viz. dental fluorosis, skeletal fluorosis, and muscular fluorosis. Fluoride not only affects humans but also the animals (skeletal and dental lesions) and plants (hinders plant growth, leaf narcosis,

Fig. 1 Countries affected with excess fluoride in drinking water (http://www.nofluoride.com/ Unicef_fluor.cfm)

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leaf chlorosis, leaf tip burn). India is among 23 nations where a large population suffers from these diseases. Various treatment technologies for removing fluoride from groundwater have been investigated in the past. After studying the various technologies that are used in past years and considering the demerits and merits, the adsorption techniques have been opted and are being used through which the low-cost adsorbents material are used in the successful removal of fluoride.

Fluoride in Drinking Water Vardhan and Karthikeyan (2011) estimated that around 260 million people (in 30 countries) are drinking water with fluoride content more than 1.0 mg/l. According to the study done by Kumar and Saxena (2011), the Indo-Gangetic Alluvium Plain has been found to be contaminated with highly fluoride content in the groundwater. The sources of water like dug wells and hand pumps show increased concentration of fluoride in water, and their study area included few villages of Unnao district, Uttar Pradesh, where about the 85% of rural population use this water as the main source for their drinking purpose, out of which almost 80% of the population were found to be adversely affected by fluorosis (skeletal, dental, and muscular).

Sources of Fluoride Contamination in Groundwater Geogenic Sources Study conducted by Koteswar and Metre (2014) revealed that the release of fluoride in the groundwater can be through the fluoride-bearing rocks like fluorspar, cryolite, and fluorapatite that are found naturally and dependent on the parameters like the solubility of minerals, availability, concentration, pH, and velocity of flowing water. The weathering processes of these rocks are also one of the major factors of fluoride in water. Chemically fluoride ions and OH ions are negatively charged, hence during chemical reaction, fluoride ions easily replace OH ions present in the rock and enhance its concentration in rocks and minerals. Whenever carbonate- and bicarbonate-rich water is passed through such type of rocks, fluoride ion is released due to chemical reactions which are listed below and percolates to groundwater and causes change in the concentrations (Saxena and Ahmed 2001). CaF2 þ Na2 CO3 ! CaCO3 þ 2F þ 2Naþ CaF2 þ 2NaHCO3 ! CaCO3 þ 2Naþ þ 2F þ H2 O þ CO2 Kumar and Saxena (2011) concluded in their studies, which was conducted in the central part of Indo-Gangetic Alluvium, that the probable source mineral for fluoride in the groundwater of the study area was mainly due to weathering of mica mineral, biotite, which was altering to chlorite and was responsible for leaching of fluoride in the groundwater. The incorporation of fluoride from the geologic formations occurs

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through the rain water and irrigation in agricultural fields; as the rainfall occurs, the water seeps into soil and further comes in contact with rocks and minerals in the aquifer materials due to which the fluoride content gets combined into the groundwater which contributes to fluoride contamination (Saxena and Ahmed 2001). Anthropogenic Sources Fluoride can be incorporated into the water through many point and nonpoint sources, i.e., industrial effluents discharged from the aluminum industries, phosphate industries, and coal plants, or it may result from the agricultural runoff from the fields which use phosphate-rich fertilizers and chemical fertilizers. Fluoride discharged from the fertilizers manufacturing processes is basically in the form of silica tetrafluoride, as result of handling phosphate-rich rocks such as fluorspar and limestone which are among the basic fluxing materials used in steel making. The aluminum processing industries use the fluoride compound cryolite as catalyst in bauxite ore reduction process and the gaseous fluoride which is produced is directly discharged into atmosphere or in the waste streams (Samal and Nayak 1988). An average fluoride value for aluminum reduction plants is 107–145 mg/l in wastewater streams. Fluoride concentrations ranging 1000–3000 Mg/L have been reported from glass manufacturing process (Sun et al. 1998). The other nonpoint sources of groundwater pollution also contribute fluoride to groundwater. Modern agricultural practices used by agricultural fields, i.e., chemical fertilizers (phosphate rich) and pesticides that contain 1–3% of fluoride and leach down into the soil due to rainfall, increase the level of fluoride in groundwater. Fluoride has also been found in some food stuffs (e.g., wheat, rice, pulses), cosmetics, and drugs. The use of drugs containing sodium fluoride for osteoporosis and dental carries is very common; different brands of toothpaste use certain excessive amounts of fluoride through which the fluoride gets introduced to the oral cavity and reaches human body (Meenakshi and Maheshwari 2006).

Health Impact Due to Fluoride Contamination For drinking water World Health Organization has prescribed the range of fluoride as 1.5 mg/l, and Bureau of Indian Standards (BIS 2012) has set the above value as desirable range and the permissible range as 1.2 mg/l. If the fluoride intake is more than the permissible limit (BIS Standards), then disease called fluorosis occurs (skeletal, dental, and muscular). Higher concentrations of fluoride have many detrimental impacts on humans: (a) Skeletal fluorosis: It is an illness in which the excessive accumulation of fluoride in bones occurs and causes pain, damage brittleness in the bones, and movement in joints become difficult. Figure 2 shows the problem of skeletal fluorosis in which deformation of leg bones is clearly visible (Patel and Saxena 2016). (b) Dental fluorosis: It is also referred as mottling of enamel, a defect caused in teeth due to excess level of fluoride (Fig. 3) in which the tooth enamel porosity increases and teeth become to erode and crumble.

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Fig. 2 Skeletal fluorosis (http://www.inrem.in/ fluorosis/about.html)

CAUSES DENTAL FLUOROSIS “Fluo*ro*sis is poisoning by fluorides” –Websters Encyclopedic Unabridged Dictionary

VERY MILD WATER @ 1PPM San Francisco

1MG PER DAY PRESCRIPTION FLUORIDE White and Brown Enamel

SEVERE WATER 4MG/DAY Brown Pitted Enamel

Fig. 3 Stages of dental fluorosis (http://cleanwatercalifornia.nationbuilder.com/infant_formula)

(c) Muscular fluorosis: It is an illness in which the muscle weakness is a common symptom and the muscle attachment protrudes with occurrence of pain. Figure 4 shows muscular fluorosis in leg muscle.

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Fig. 4 Muscular fluorosis (After Saxena 2005)

Fig. 5 Problem in thyroid gland (http://www.medhealth.net/Symptoms-OfThyroid-Problems.html)

Thyroid Gland Right Lobe Enlarged Thyroid Gland

Voice Box (Larynx) Thyroid Gland left Lobe Lathmus Trachea (Wins Pipe)

(d) Other diseases: Include thyroid disorder, osteoporosis, and lower IQ level in children, which inhibit or stop their growth (Patel and Saxena 2016). The problem of thyroid gland and osteoporosis has been shown in Figs. 5 and 6. The severity of fluorosis depends on the concentration of fluoride in drinking water, daily intake, continuity and duration of exposure, and climatic conditions. WHO recommends that water containing a minimum of 0.6 mg/l and a maximum of 1.5 mg/l fluoride is considered safe for drinking purposes. Table 1 shows effect of

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Fig. 6 Bones affected with osteoporosis (http://www. thehealthsite.com/diseasesconditions/osteoporosis/001/)

Osteoporosis Bone with osteoporosis Normal bone matrix

Table 1 Effects of fluoride in water on human health Fluoride concentrations (mg/l) SO42 > NO3. Iron-Based Sorbents for Defluoridation of Water Most of the adsorbents for fluoride removal have been tested for drinking water and would not be stable at extreme pH values unless the pH is adjusted. Thus, polishing industrial wastewaters

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Drinking water

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Warm Water Cold Water Cartridge textile filter

Flow controller & water meter

Activated alumina filter

Max flow 21/min

Fig. 14 Activated alumina process (http://www.appropedia.org/Water_defluoridation)

containing excess fluoride stands as a major problem. Due to the stability at low pH and its magnetic properties, granular ferric hydroxide (GFH) for the removal of various ions including fluoride was used for the treatment of fluoride contaminated wastewater (Mohapatra et al. 2004). • Other metal oxides/hydroxide/oxyhydroxide, mixed metal oxides, metal impregnated oxides as sorbents for defluoridation of water: The metal oxyhydroxides have surface oxygen which differs in the number of coordinating metal ions to facilitate the adsorption of different cations and anions. This property of oxide minerals was considered as an advantage for fluoride removal. Refractory grade bauxite (RGB), feed bauxite (FB), manganese ore, magnesiaamended silicon dioxide granules (MAS), titanium oxysulfate TiO (SO4), and hydrated oxides of Zirconium ores (WAD) (Blackwell and Carr 1991) were taken as test adsorbents for fluoride removal from aqueous solutions. These adsorbents were able to remove fluoride in real waste water at very low concentration. Fluoride ions were easily desorbed using a high pH solution, completely

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regenerating for further removal process at acidic pH, and the capacity for fluoride adsorption remained almost unchanged three times after repeated adsorption and desorption (Biswas et al. 2009). • Natural materials as sorbents for defluoridation of water: Various naturally occurring materials available abundantly free of charge have been explored as adsorbents for the removal of fluoride from water. The efficiency of four different coal-based sorbents, lignite (LN), fine coke (FC), raw laterite, and bituminous coal (BC) was evaluated for fluoride sorption from water and found to be good if they are given treatment (Ma et al. 2003). Bioadsorbents for Defluoridation of Water Biosorption is an emerging technique for water treatment utilizing abundantly available biomaterials. Various bioadsorbents have been developed for fluoride removal, and among those chitin and chitosan-derivatives have gained wide attention as effective bioadsorbents due to their low cost and high contents of amino and hydroxyl functional groups which show significant adsorption potential for the removal of various aquatic pollutants. Algal biomass pretreated with Ca2+ was also evaluated for the biosorption of fluoride from polluted waters. The fluoride sorption phenomenon on fungal bioadsorbent was attributed to the chemical type of interaction. Collagen fiber, a profuse natural biomass, has abundant functional groups, is capable of chemically reacting with different types of metal ions, and can be used as a carrier of metal ions (Jagtap et al. 2011). Agricultural Wastes as Sorbents for Defluoridation of Water Agricultural waste materials being economic and eco-friendly due to their unique chemical composition, availability in abundance, renewable nature, and low cost are viable option for water and wastewater remediation. Low-cost adsorbents from different agricultural waste materials such as coconut shell, coconut shell fibers, and rice husk were developed and employed for the removal of various pollutants in industrial wastewater including fluoride (Alagumuthu and Rajan 2007). Industrial Waste as Many Sorbents for Defluoridation of Water Widespread industrial activities generate huge amount of solid waste materials as by-products and these products can be used as defluoridating agents. Various industrial wastes have the potential with or without treatment for fluoride removal from water. Fly ash (a thermal power plant waste), carbon slurry (a fertilizer industry waste), red mud, original waste mud, acid-activated and -precipitated waste mud, spent bleaching earth, a solid waste from edible oil processing industry, and alum sludge are some wastes that can be used for defluoridation purposes after some treatment (Kemer et al. 2009). Nanosorbents for Defluoridation of Water As per the study of Sarkar et al. 2007, nanotechnology has emerged as a promising technology in past decade in various fields and use of nanoparticles as sorbents for water treatment is also gaining wide attention in recent years. Aligned carbon nanotubes, prepared by catalytic

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A1 Anode

+ + + + + + + + + +

A13+

Cl–

F– A13+

A13+ F–

A13+ A13+

Cl–

Na+

F–

Cl– Cl–

Cl–

F–

F– Cl–

Fig. 15 Representation of electrocondensation effect (Hu et al. 2003)

decomposition of xylene using ferrocene as catalyst, proved their good performance for fluoride removal from water. Both the surface and inner cavities of aligned carbon nanotubes were found to be readily accessible for fluoride sorption. A variety of nano-sized inorganic oxides prepared through thermolysis of a polymeric-based aqueous precursor were capable to give the solution of the desired inorganic ions (Sarkar.et al. 2007). Like this various efforts in this field have been made to take the advantage of nanoparticles (Ayoob and Gupta 2009).

Electrochemical Technique Electrocoagulation Electrocoagulation (EC) utilizes an electrolytic process to generate a coagulant in situ by oxidation of an appropriate anodic material. The coagulant ions then react with the target pollutant ions, initiating normal coagulation. When an electric current passes through the aluminum electrodes, an anodic reaction releases Al(III) ions and these ions react with hydroxide ions produced at the cathode and with fluoride ions in solution. Defluoridation efficiency of the EC system may exceed that of the traditional coagulation process because of electrocondensation. Since fluoride ions are attracted to the anode, fluoride concentration near the anode exceeds that in the bulk solution, which may lead to higher efficiency (condensation effect) for the EC process. The modification of the process was done by introducing bipolar electrodes. A conductive plate without any electric connection, placed between two electrodes having opposite charges, will develop bipolarity and hence it is known as the bipolar electrode (Mameri et al. 1998). Figure 15 shows the electrocondensation effect, and Fig. 16 shows bipolar electrode system.

Coagulation As per the study of Gregory and Duan (2001), chemical coagulants are effective in removing a broad range of impurities from water, including colloidal particles and dissolved organic substances. Two mechanisms essentially operate: charge neutralization of negatively charged colloids by cationic hydrolysis products and addition

1808 Fig. 16 Schematic diagram of bipolar electrode system (After Mameri et al. 1998)

A. Saxena and A. Patel Al

Al

Al

Al

F– Al3+ F– Al3+ OH–

of impurities in an amorphous precipitate of metal hydroxide, generally known as “sweep flocculation.” Charge neutralization is a simple mechanism for destabilization of negatively charged particle by adsorption of cations from solution, and this adsorption of positively charged species on negative surfaces may occur for simple electrostatic reasons or by formation of some types of surface complexes. “Sweep flocculation” refers to the “sweeping out” of particles from water by hydroxide precipitate and generally gives improved particle removal. Thus, coagulation by hydrolyzing coagulants may involve both charge neutralization and sweep flocculation as essential steps in particle removal (Gregory and Duan 2001). Electrosorption Electrosorptive techniques have been used to treat various contaminated waters and also to enhance sorption capacity of the conventional systems. Electrochemical cell, equipped with two stainless-steel electrodes, was introduced into a polyvinyl chloride column to produce an electrical field in the activated alumina bed. Normally, chemical desorption and regeneration result in reduced sorption capacities. But the electrode sorption techniques are found to be better. Ninety percent less cleaning agent (NaOH) was used as compared to conventional regeneration techniques. A 95% recovery of the adsorption capacity was realized with the electrosorption system, and the volume of water required to regenerate the saturated bed was only 6% of the treated water volume, a much lower value than with current regeneration techniques (Lounici et al. 2004). Nanofiltration Study done by Velizaro et al. 2004 gives details about nanofiltration, which is also a pressure-driven membrane process, which is generally used for separating multivalent ions from monovalent species. It is also possible to achieve separation of ions of same valence by selecting a proper membrane and operating conditions. NF membranes are essentially low-pressure RO membranes. Because NF membranes provide higher water fluxes at lower transmembrane pressures than traditional RO membranes, they are sometimes referred to as “lowpressure RO membranes” or “loose RO membranes.” The NF membranes are usually asymmetric with negative charge at neutral and alkaline environment and also have a “loose” polymer structure. The origin of surface charge on NF membranes are the anions adsorbed on its surface, whereas in ion exchange membranes, it is due to various fixed charged groups bonded to the polymer structure. As a result of

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this additional mechanism of ion exclusion (in addition to size-based exclusion), high degrees of ion separation (ion rejections) similar to those in RO, but with higher water fluxes through the NF membranes could be achieved. But the NF process is much more sensitive than RO to pH and ionic strength of source water (Velizaro et al. 2004).

Ion Exchange Method As per the study done by the Ingle et al. 2017, fluoride removal can be obtained from this method also by using anion exchange resin containing quaternary ammonium groups. The chemical representation of the process is as follows: MatrixNR3 þ Cl þ F ! MatrixNR3 þ F þ Cl Fluoride ion gets replaced by chloride ions of the resin; the process works until resin completely gets used up and then further backwashed with the solution which is supersaturated with sodium salt. The new chloride ions replace the fluoride ion. The advantages of this process are that it has highest fluoride removal efficiency and maintains the quality of the water. The main disadvantage of this process is that it is very costly and pH of the water is very low and chloride concentrations are higher. Figure 17 shows an illustration of ion exchange process: Electrodialysis Electrodialysis (ED) is an electromembrane process developed for desalting and demineralization of water. In this process, the transport of ions present in contaminated water is accelerated due to an electric potential difference applied externally. The separation is accomplished by alternately placing cationand anion-selective membranes in a parallel fashion across the current path to form an ED cell. A schematic representation of a typical ED cell is given in Fig. 18 (Marder et al. 2003). Donnan Dialysis Donnan dialysis (DD) is a membrane separation process that uses an ion exchange membrane similar to electromembrane processes but with a concentration gradient as driving force rather than applying an external electric potential Fig. 17 Typical representation of ion exchange process (https:// iaspub.epa.gov/tdb/pages/ treatment/treatmentOverview. do?treatmentProcessId= 263654386)

Influent Upper Distributor

Zeolite/Resin

Underdrain Effluent

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ANODE

CATHODE

A

Anions Cation

C

A

C

DS - Distilled Solution CS - Concentrated Solution ERS - Electrode Resin Solution

A

C

Feeding Solution

A - Anion selective membrane C - Cation selective membrane

Fig. 18 Schematic diagram of a typical ED cell (After Marder et al. 2003)

AEM Water (A)

NaCl and xM A13+ (R)

(b)

(a)

Receiver

Feed Circuit

(a) OPEN RECEIVER (b) CLOSED RECEIVER

Fig. 19 Schematic flow diagram of DD system (After Hichour et al. 2000)

difference across the membrane. Its operation requires addition of a driving counterion to stripping solution, which is transported in a direction opposite to that of the target ion in order to maintain electroneutrality. The ions which are permeable to the membrane will equilibrate until the conditions of Donnan equilibriums are satisfied (Sarkar et al. 2010). Figure 19 shows the flow diagram of Donnan dialysis.

Reverse Osmosis In this process the hydraulic pressure is applied on one side of the semipermeable membrane which forces the water across and the leftover salt is obtained; the membranes are sensitive to pH and temperature. The demerits of this method are it

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Fig. 20 Process of reverse osmosis (http://puretecwater. com/reverse-osmosis/what-isreverse-osmosis)

requires skilled supervision, the units might get prone to chemical attack, it is not an economical process or is an expensive option to adopt, and large quantity of waste is generated (Renuka and Puspanjali 2013). The reverse osmosis process is shown in Fig. 20.

Comparison of Different Defluoridation Technologies Used On comparison of different methods of defluoridation mentioned above for the successful removal of fluoride, it was found that each methodologies have some merits and demerits. Table 4 shows the comparison between different techniques and Table 5 shows the comparison of adsorption capacity of different adsorbents.

Adsorption Treatment Techniques Fluoride can be removed by adsorption onto many adsorbent materials; the criteria for selecting suitable sorbents are cost of medium and running costs, ease of operation, adsorption capacity, and possibility of regeneration and reuse. Adsorption technique uses a bed of defluoridating material along which the raw water is passed. The defluoridated material restricts the fluoride either by physical, chemical, and ion exchange mechanisms. Hence the adsorbent material gets saturated after a time period of operation and requires regeneration, pH dependent. The methods can be used to eliminate fluoride contamination due to physical and chemical behavior with the adsorbents. Adsorbent is a porous substance that has a high surface area and has the capability to absorb or adsorb other substances using intermolecular forces onto its surface and a suitable example of an adsorbent material is activated charcoal. Adsorption is the attachment or adhesion of atoms, ions, and molecules (adsorbates) from a gaseous, liquid, or solution medium onto the surface of an adsorbent – activated carbon. The porosity of activated carbons offers a vast surface on which this adsorption can take place. Adsorption occurs in pores slightly larger than the

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Table 4 Comparison of different defluoridation techniques Defluoridation Defluoridation techniques capacity/dose Electrochemical methods Electrocoagulation Efficiency ion 100%

Electrosorption

Highly efficient

Membrane processes Reverse Highly efficient Osmosis

Nanofiltration

Highly efficient

Electrodialysis

Highly efficient

Donnan dialysis

Highly efficient

Adsorption

Efficient

Strength

Limitations

Emerging technique. Efficiency of EC system is very high compared to the traditional coagulation process

Interference from other anions like sulfate. Need for regular replacement of sacrificial electrodes, due to high consumption of electric power Costly due to high consumption of electric power

Emerging technique. Capacity of adsorbent enhanced by more than 0%. Excellent regeneration potential Well-studied and established Technology. Immense applications, commercial applications. Dominant in many developed countries. Small foot print. Organics and salts are also removed Well-accepted membrane separation process. Handles higher water fluxes at lower transmembrane pressures than RO. Excellent technique for simultaneous defluoridation and desalination. Commercially established. More economical than RO. More resistant to fouling Recently, received attention in treating fluoride. Electromembrane processes but with concentration gradient as driving force. A permanent separation between solutions which is not reversed even if the system is closed to the surroundings Widely used

Sensitive to polarization Phenomenon. Chances of biological and mineral fouling. Treated water may lack the right balance of minerals. Poor water recoveries. High cost More sensitive than RO to pH and ionic strength. Leaves large concentration of retentate fraction. Expensive technique. Require high degree of pretreatment. Ineffective in removing lowmolecular-mass noncharged compounds

Operation requires addition of a so-called driving counterion to stripping solution. Reduced efficiency in high-saline waters. Expensive technique

Easy in operation

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Table 5 Comparison of adsorption capacity different adsorbent Adsorbent Bone

Defluoridation capacity/dose 0.9 mg F/g

Bone char

2–4 mg F/g

Clays

0.03–0.35 mg F/g

Economical. Very limited local application

Activated alumina

1.0 mg F/g

Very well established technique. Regarded as one of the best available technology. Best performance at pH 5. Minimum interference from counterions with consistent potential

Calcium-based adsorbent

Up to 84.8 mg F/g

Can defluoride highly contaminated water. Very quick action

Iron-based sorbents

Up to 58.6 mg F/g

Metaloxides/ hydroxides/ oxyhydroxide, mixed metal oxides, metalimpregnated oxides.

Up to 98 mg F/g

Stable at low pH, have magnetic properties. Insensitive to temperature change. Do not affect the adsorption of arsenate Cost-effective alternative, removal can occur at low pH

Carbon-based adsorbents

Up to 3.13 mg F/g

Strength Long-established technique for local applications Well-known and established technique. Good potential. Local availability and processing facilities aids local applications. Ability to remove fluoride to very low levels

Cheap, easily available

Limitations Impart taste to water. Limited social acceptance Capacity reduces drastically after successive regenerations. More expensive than coagulation techniques. The use is constrained by the religious beliefs in many societies and communities. Limited social acceptance Defluoridation potential is generally low. Regeneration is very difficult Costly compared to coagulation processes and bone char. High pH reduces potential. Regeneration results in a reduction of about 5–10% in material, and 30–40% in capacity with increased presence of aluminum (>0.2 mg/L) Cannot be regenerated easily. Do not work in acidic conditions. (pH < 3) Very sensitive in high pH range (>8)

Impart other ions to the defluoridated water. Phosphate and sulfate ions put adverse effect. Efficiency decreases at high pH (>10) Low defluoridation capacity (continued)

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Table 5 (continued) Adsorbent Natural materials

Defluoridation capacity/dose 3–5 mg F/g

Biosorbents

Up to 4 mg F/g

Agricultural wastes

Up to 18.9 mg F/g

Industrial wastes

Up to 11.8 mg F/g

Nanosorbents

Up to 28.7 mg F/g

Strength Almost free of cost, easily available, do not impart any contaminant to the treated water Emerging technique, low cost and high concentration of amino and hydroxyl group that can remove other contaminants too Economic, eco-friendly, available in abundance, renewable in nature, low cost Waste materials and by-products can be reused, cheap, and easily available Small size, large surface area, high mechanical strength, and conductivity

Limitations Can treat water having low concentration of fluoride Low removal capacity, affected by pH variation

Removal is affected by the presence of carbonates, sulfates, and nitrate and bicarbonate ions Large amount of sludge generated, highly sensitive to pH change High pressure drop if used in column, phosphate, chloride, nitrate, etc.

molecules that are being adsorbed, which is why it is very important to match the molecule you are trying to adsorb with the pore size of the activated carbon. These molecules are then trapped within the carbon’s internal pore structure by Van der Waals forces or other bonds of attraction and accumulate onto a solid surface. The two types of adsorption are: • Physical adsorption – During this process, the adsorbates are held on the surface of the pore walls by weak forces of attraction known as Van der Waals forces or London dispersion forces (http://www.haycarb.com/activated-carbon). • Chemisorption – This involves relatively strong forces of attraction, actual chemical bonds between adsorbents and chemical complexes on the pore wall of the activated carbon (http://www.haycarb.com/activated-carbon).

Low-Cost Adsorbents for Fluoride Removal Different low-cost adsorbent materials are available for effective removal of fluoride from water. The naturally available adsorbents, chalk powder, pineapple powder, orange peel powder, multani mitti (special type of clayey soil), activated neem and

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low cost adsorbents

Percentage of Fluoride Removal by Low Cost Adsorbents Multani Mitti Neem Leaves Kikar leaves Orange Peel Powder Rice Husk Chalk Powder Pineapple Peel…

Percentage of Fluoride Removal

0%

50%

100%

Percentage of fluoride removal

Fig. 21 Fluoride removal by low-cost adsorbents (After Koteswar and Metre 2014)

kikar leaves, rice husk, etc., are some of the different materials investigated for adsorptive removal of fluoride from water. Figure 21 gives a comparative fluoride adsorption of few such adsorbents. (a) Chalk Powder: The main component of chalk is calcium carbonate (CaCO3), a form of limestone. Limestone deposits develop as coccoliths (minute calcareous plates created by the decomposition of plankton skeletons) accumulate, forming sedimentary layers. Plankton, a tiny marine organism, concentrates the calcium found naturally in seawater from 0.04% to 40%, which is then precipitated when the plankton dies. The base of pastel chalks is calcium sulfate (CaSO4), which is derived from gypsum (CaSO4 .2H2O), an evaporate mineral formed by the deposition of ocean brine; it also occurs disseminated in limestone. Chalk and dehydrated gypsum thus have similar origins and properties although great care is taken to eliminate contaminants when chalk is manufactured; some impurities inherent to the mineral remain (http://www.madehow.com/Volume-1/Chalk. html). As calcium carbonate decomposes only at 900  C, the adsorption taking place, chalk powder due to certain porosity adsorbs fluoride from aqueous solution. The adsorption of fluoride is 86%. (b) Pineapple peel powder: Pineapple (Ananascosmosus) is a tropical fruit which grows in countries which are situated in the tropical and subtropical regions. Pineapple belongs to the Bromeliaceae family and grows on the ground. It can grow up to 1 m in height and 1.5 m wide. Other bromeliads live on trees (epiphytes). There are many cultivars of Ananas. Total pineapple production worldwide is around 16 to 18 million tons. There are several countries (e.g., Thailand, Brazil, India, Philippines, and China) which contribute to the total production (http://foodscience.wikispaces.com/Pineapple). The pineapple peel is the waste product generated from processing pineapples. Hence it is an agricultural waste which is cheap for modern communities and easily available (Hajar et al. 2012). The adsorption of fluoride is 86%.

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(c) Orange peel powder: Botanically, orange (Citrus sinensis) is the citrus fruit belonging in the Rutaceae family, of the genus, Citrus. The family also includes other related species of oranges such as pomelo, tangerine (mandarin orange), yuzu, lemon, and grapefruit. Orange is a tropical to semitropical, evergreen, small flowering tree, growing to about 5–8 m tall, and bears seasonal fruits that measure about 3 inches in diameter and weigh about 100–150 g. Oranges are classified into two general categories: sweet and bitter, former being the most commonly consumed type (http://www.nutrition-and-you.com/orange-fruit. html). Orange peel are rich in flavonoids that have chemical composition as well as some trace elements, ascorbic acid, carotenoids, dietary fibers, and total polyphenols, and the antiradical efficiency was assessed in the dried peels; due to certain porosity of orange peel powder adsorbs fluoride from aqueous solution. The fluoride removal is 79%. (d) Multani mitti: Multani mitti or fuller’s earth is clay material that is popularly used for skin care ingredient. It is rich in magnesium chloride and has great surface area with excellent bonding and sealing properties (http://stylesatlife. com/articles/multani-mitti-benefits/). It contains grains of fine sand particle; they contain complex multicenter crystalline structures of oxides and hydroxide of magnesium, aluminum, zinc, and silicon. It has the tendency to filter, decolorize, and clarify properties and other liquids without any chemical treatment. The efficient removal of fluoride can be done up to 56%. (e) Kikar leaves: Kikar is the small thorny tree. It grows to the height of 7–12 m. It has yellow round head flowers which are nectarless. The bark is red-brown to blackish and rough. The leaves are light green and fern-like up to 120 mm long and 50 mm wide. Flowers are 10–15 mm in diameter, which grow in cluster between 4 and 6 and are sweetly scented (http://www.ecoindia.com/flora/trees/ kikar-tree.html). The leaf, bark, seeds, and the gum are used for medicinal purposes here in India. Babul is well known for its amazing use in treating gum problems. Chemical constituents: Leaves and flowers contain tannin, bark contains tannin and other polyphenolic compounds, and fruits contain gallic acid. Babul tree has many medicinal properties as babul has wonderful antiviral and antifungal properties. It also has antidiarrheal, antioxidant, antibacterial, antimalarial, anthelmintic, and anti-inflammatory properties (http://www. wildturmeric.net/2016/12/babul-babool-acacia-arabica-medicinal-uses-side-effe cts.html). Kikar leaves are cheap adsorbents for fluoride removal with the removal efficiency of 59.2%. (f) Rice husk: Rice milling generates a by-product know as husk. This surrounds the paddy grain. During milling of paddy, about 78% of weight is received as rice, broken rice, and bran. The remaining 22% of the weight of paddy is received as husk. This husk is used as fuel in the rice mills to generate steam for the parboiling process. This husk contains about 75% organic volatile matter and the balance 25% of the weight of this husk is converted into ash during the firing process known as rice husk ash (RHA). This RHA in turn contains around 85–90% amorphous silica. So for every 1000 kgs of paddy milled, about

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220 kgs (22%) of husk is produced. India is a major rice producing country, and the husk generated during milling is mostly used as a fuel in the boilers for processing paddy, producing energy through direct combustion and/or by gasification. About 20 million tons of RHA is produced annually (http://www. ricehuskash.com/application.htm). Composition of rice husk ash: Rice husk contains abundant floristic fiber, protein, and some functional groups such as carboxyl, hydroxyl, amidogen, which makes adsorption processes possible. There is marginal variation in fluoride removal by rice husk over pH range of 2 to 10 from 83% to 84% removal efficiency. (g) Neem leaves: The neem tree (Azadirachta Indica) is a tropical evergreen tree native to India and is also found in other southeast countries. In India, neem is known as “the village pharmacy” because of its healing versatility, and it has been used in ayurvedic medicine for more than 4000 years due to its medicinal properties. Neem is also called “arista” in Sanskrit, a word that means “perfect, complete and imperishable.” The seeds, bark, and leaves contain compounds with proven antiseptic, antiviral, antipyretic, anti-inflammatory, antiulcer, and antifungal uses (http://www.organeem.com/neem_tree.html). Neem is extremely beneficial to save the environment from pollution; since its in-florescence is purifying “with its feathery crests tossing fifty feet into the sky” neem is a veritable “Kalpataru” for giving healthy environs. Like other trees, it exhales out oxygen and keeps the oxygen level in the atmosphere balanced. Neem is a natural resource to keep environment clean. In villages and cities as well as on farms, it is useful as a windbreak. As a source of shade, it is excellent for parks, roadsides, etc. Because of its so many qualities, it is a common practice in rural India to have a neem tree within the compounds of most of the houses. Neem is also regarded as a valuable forestry species in India (http://www.neemfoundation.org/about-neem/neem-environment/). Neem leaves can thus be used by rural communities as it is a most common and easily available tree. With the help of neem leaves, fluoride can be removed up to 58%.

Cross-References ▶ Advanced Treatment Technologies ▶ Bioremediation of Mined Waste Land ▶ Fluoride Contamination in Groundwater and the Source Mineral Releasing Fluoride in Groundwater of Indo-Gangetic Alluvium, India ▶ Global Status of Nitrate Contamination in Groundwater: Its Occurrence, Health Impacts, and Mitigation Measures ▶ Identification of the Source Mineral Releasing Arsenic in the Groundwater of the Indo-Gangetic Plain, India ▶ Micro-remediation of Metals: A New Frontier in Bioremediation

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Meenakshi, Maheswari RC (2006) Fluoride in drinking water and its removal. J Hazard Matter 137:526–527 Mjengera H, Mkongo G (2003) Appropiate Defluoridation technology for use in fluoride area in Tanzania. Phys Chem Earth 28:1097–1104 Mohapatra D, Mishra D, Mishra SP, Chaudhary GR, Das RP (2004) Use of oxide minerals to abate fluoride from water. J Colloid Interface Sci 275:355–359 Patel A, Saxena A (2016) Role of bioremediation as a low cost adsorbant for excessive fluoride removal in groundwater. In: International conference on emerging trends in civil engineering (ICETCE) Quershi N, Malmberg RH (1985) Reducing aluminium residuals in finished waters. J Am Water Works Assoc 77:101–108 Rajiv Gandhi National Drinking Water Mission (RGNDWM) (2001) Making water safe, news letter on rural water and sanitation in India, Published by RGNDWM and Water and Sanitation Program South Asia Renuka P, Puspanjali K (2013) Review on defluoridation techniques of water. Int J Eng Sci 2(3):86–94 Samal UN, Naik BN (1988) Dental fluorosis in school children in the vicinity of aluminium factory in India. Fluoride 21(3):142–148 Sarkar S, Blaney LM, Gupta A, Ghosh D, Gupta AKS (2007) Use of ArsenXnp, a hybrid anion exchanger for arsenic exchanger for arsenic removal in remote villages in the Indian subcontinent. React Funct Polym 67(12):1599–1611 Sarkar S, Greenleaf JE, Gupta A, Ghosh D, Blaney LM, Bandhopadhay P, Biswas R, Dutta AK, SenGupta AK (2010) Evolution of community- based arsenic removal systems in remote villages: assessment of decade-long operation. Water Res 44:5813–5822 Saxena A (2005) Sedimentological and mineralogical studies of the quaternary sediments of Unnao District (Nawabganj Area) with special reference to fluoride contamination in the groundwater. – India [unpublished], pp 63–76 Saxena VK, Ahmed S (2001) Dissolution of fluoride in groundwater: a water rock interaction study. Environ Geol 40:1084–1087 Saxena R, Nathawat GS (1990) Germination behaviour of Hordium vulgare with respect to increasing concentration of sodium fluoride. Acta Ecologica 12(12):103–105 Sorg TJ (1978) Treatment technology to meet the interim primary drinking water regulation for inorganics. J Am Water Works Assoc 70:105–112 Sun Z, Cheng Y, Zhou J, Wei R (1998) Research on effect of fluoride pollution in atmosphere near an aluminium electrolysis plant on regional fall wheat growth. In: Proceedings of air and waste management association 91st annual meeting, San Diego, TPE 09/P1-TPE09/P7 Suttie JW, Kolstad DL (1977) Effects of dietary fluoride ingestion on ration intake and milk production. J Dairy Sci 60:1568–1573 Turner BD, Binning P, Stipp SLS (2005) Fluoride removal by calcite: evidence for fluorite precipitation and surface adsorption. Environ Sci Technol 39(24):9561–9568 Vivek Vardhan CM, Karthikeyan J (2011) Removal of fluoride from water using low cost materials. International Water Technology Journal, I(2):1–12 Velizaro S, Crespo GJ, Reis AM (2004) Removal of inorganic anions from drinking water supplies by membrane bio/processes. Rev Environ Sci Bio/Technol 3(4):361–380 Wasay SA, Haron JM, Tokunaga S (1996 part 1) Adsorption of fluoride, phosphate and arsenate ion on lathanum-impregnated silica gel. Water Emission Res 68:295–300 Weinstein LH, Davison AW (2004) Fluoride in the environment effect on plants and animals. CAB International, Wallingford Yadav RK, Sharma S, Bansal M, Singh A, Pandey V, Maheshwari R (2012) Effects of fluoride accumulation on growth of vegetable and crops in Dausa District Rajasthan, India. Adv Biores 3(4):14–16. ISSN: 0976–4875

Biotechnological Approach for Mitigation Studies of Effluents of Automobile Industries

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N. N. Bandela, P. N. Puniya, and Geetanjali Kaushik

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Literature Review . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Results and Discussion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cross-References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

Automobile industry effluents are highly contaminated with various heavy metals like Zn, Ca, Pb, Ni, Cr, and Fe, paint particles, coolants, phosphate coating, and oil and grease. The discharge of such toxic effluents without any treatment contaminates natural water bodies. To study the efficiency of biological treatment of the feeding effluent of automobile industries, two pilot plants were set up at a lab scale: one was the conventional bioreactor plant and another was the novel bioreactor with modified design concept. In the novel bioreactor, inside baffles are constructed, and two impellers are used: one at the surface and the other at the bottom. After the comparative study, it was finally concluded that the novel bioreactor efficiency was two times more than the conventional bioreactor. Hence, it is recommended that novel bioreactors can play a vital role in treating the effluent of automobile industries. The microbe of the activated sludge helps to

N. N. Bandela · P. N. Puniya Department of Environmental Sciences, Dr. Babasaheb Ambedkar Marathwada University, Aurangabad, Maharashtra, India G. Kaushik (*) MGM’s Jawaharlal Nehru Engineering College, Mahatma Gandhi Mission, Aurangabad, Maharashtra, India © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_146

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adsorb various heavy metals from the effluent. Pseudomonas aeruginosa was found abundant in the effluent of automobile industries. Keywords

Treatment · Automobile industry · Heavy metals · Bioreactor · Reduction

Introduction Since the 1990s, the surge in the automobile industry resulted in a robust growth of the auto component sector within the country. In tandem with the industry trends, the Indian component sector has shown great advances in recent years in terms of growth, spread, absorption of new technologies, and flexibility. The Indian auto component industry has seen major growth with the arrival of world vehicle manufacturers from Japan, Korea, the United States, and Europe. Today, India is emerging as one of the key auto component centers in Asia and is expected to play a significant role in the global automotive supply chain in the near future (Joseph and Nagendran 2004). The auto component industry is today considered as the rising industry with huge growth potential. This industry is also expected to drive the growth of the engineering sector in view of its strong downstream and upstream linkages with many other segments of the engineering sector like raw materials, capital goods, intermediate products, etc. Today, the auto component industry has emerged as a highly competitive segment of the manufacturing sector. The Indian auto component industry is wide (over 400 firms in the organized sector producing practically all parts and more than 10,000 firms in small unorganized sector, in tierized format) and has been one of the fastest-growing segments of the auto industry. During the year 2006–2007, the auto component industry continued its high growth path and emerged as one of the fastest-growing sectors in the Indian engineering industry by clocking 21% growth in output during the year. The industry crossed a total turnover of over US $ 15 billion (Rs. 64,500 crore), with exports of US $ 2.9 billion (Rs. 12,643 crore) during the year. It supports industries like automobiles, machine tools, steel, aluminum, rubber, plastics, electrical, electronics, paints, forgings, and machining. India has also emerged as an outsourcing hub for auto parts for international companies such as Ford, General Motors, Daimler Chrysler, Fiat, Volkswagen, and Toyota. Aurangabad city in Maharashtra state of India has emerged as an industrial hub (Table 1) with the presence of various automobile giants in its industrial clusters. The Shendra, Chikalthana, and Waluj MIDC industrial areas are prominent industrial zones on the outskirts of the city, with various major multinational groups having set up manufacturing or processing plants in and around the city Brief Industrial Profile of Aurangabad District (2013–2014). In Waluj MIDC area, almost 70% industries manufacture automobile parts. As per MPCB (Maharashtra Pollution Control Board), effluent quantity of approximately 5 MLD is released per day from

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Biotechnological Approach for Mitigation Studies of Effluents of. . .

Table 1 Aurangabad city Present population at a glance Present households Area sq. kms. Height (MSL) Latitude Longitude Annual rainfall (mm-average) Main tourist spots

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12,000,00 (approx.) 1,55,000 (approx.) 138.53 611 m 19o530 5000 N 75o220 4600 E 725 Ellora caves, Bibi ka Maqbara, Panchakki, etc.

Source: AMC (2016)

this area. To study the characteristics of automobile industrial effluent sample was collected from the Waluj MIDC area (MSME 2014).

Literature Review The automobile industry provides employment to a large section of the population. Thus, the role of the automobile industry cannot be overlooked in Indian economy. All kinds of vehicles are produced by the automobile industry. It includes the manufacturing of trucks, buses, passenger cars, defense vehicles, two-wheelers, etc. The industry can be broadly divided into the car manufacturing, two-wheeler manufacturing, and heavy vehicle manufacturing units. The automobile industry is a metallurgic and engineering industry, wherein water is used in huge quantities for coolant preparation, surface preparation for electroplating, surface preparation for phosphate coating, washing and cleaning processes used for removing soils and other impurities from iron and aluminum sheets, and surface preparation for painting purposes. On an average, each leading industry generates a minimum of 3200 M3 wastewater per day. From this average, one can get an idea of water pollution load generated per day by the automobile industry. The generated effluents from the automobile industry are highly contaminated with various heavy metals like Zn, Ca, Pb, Ni, Cr, and Fe, paint particles, coolants, phosphate coating, oil and grease, etc. (Davies 2003; Degarmo et al. 2003). The main units of water pollution in automobile industries are shown in Fig. 1. If the generated effluent is discharged as it is in natural water resources, it creates toxicity for aquatic flora and fauna. During the survey, it was observed that generated wastewater in automobile industries was treated by chemical methods only and then directly discharged in the natural resource of water. However, due to the scarcity of water, huge quantities of wastewater can be reused for washing, cleaning, and plantation purposes after proper treatment. After chemical treatment, biotechnological method should be employed. In general, physicochemical methods such as chemical precipitation, chemical oxidation or reduction, electrochemical treatment, evaporative recovery, filtration, ion exchange, and membrane technologies are widely used to remove heavy metal

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Coolant Treatment Plant (CTP)

Synthetic Coolant

Electro-plating

Phosphate Coating

Paint Shop

Water based Paint

Zn Plating Vegetable Ni & Cr Plating Blakodizing Anodizing

Applied for steel processing

Plastic or top Coating Paint

Applied for Aluminium processing

Fig. 1 Major units generating effluents from automobile industries

ions from the effluent in automobile industries. All these processes are used especially when heavy metal ions are in solution form present in the order of 1–100 mg dissolved heavy metal ions/L (Volesky 1990). The operating costs of these processes are expensive. Even after employing all these methods, trace amounts of heavy metals still remain within the effluent. Therefore, biological methods such as bioaccumulation for the removal of heavy metal ions will prove an attractive alternative technique for physicochemical methods. The removal of heavy metals by the bioaccumulation process was also studied by Kapoor and Viraraghavan (1995). Bioreactors are at the heart of biotechnological processes and device used in industries for wastewater treatment. The biological activities take place in a tank, i.e., called bioreactor. The bioreactor’s environmental conditions such as gas (i.e., air, oxygen, nitrogen, carbon dioxide) flow rates, temperature, pH, and dissolved oxygen levels, and agitation speed/circulation rate need to be closely monitored (Metcaff and Eddy 2001). In automobile industries, aerobic biological treatment is basically involved in the stabilization of biodegradable organic contents of wastewater by the mixed population of microorganisms. During the stabilization of organic content, biodegradable organic matter is oxidized or synthesized by microorganisms in aerobic conditions to produce new cells, inert solids, and other simple end products. Rubin (2001) has also designed bioreactors as a suitable environment of nutrients, moisture, dissolved oxygen, and enriched culture microorganisms, which enabled the breakdown of the diesel oil. This case study indicated that the target of 99% of diesel oil cleanup could be achieved by using the technology of bioremediation. Methodology The Waluj MIDC area in Aurangabad of Maharashtra state was selected for studying the effluent characteristics and treatment methodology of automobile industries. To know the characteristics of the generated effluents of automobile industries, parameters such as color, odor, temperature, TSS, TDS, oil and grease, pH, DO, COD, BOD, PO4, OUR (oxygen uptake rate), MLSS (mixed liquor suspended solids), MLVSS (mixed liquor volatile suspended solids), SVI

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1825

(sludge volume index), and heavy metals were studied. The effluent generated from the automobile industry is mainly contaminated with aromatic chemicals (as thinners and paint particles), oil and grease (used for lubrication), phosphate and various heavy metals like chromium, copper, zinc, and lead (used for plating), etc. Hence, to maintain the effluent quality as per the MPCB norms is one of the greatest challenges for automobile industries. During the analysis, various characteristics of the inlet effluent of automobile industries were recorded with the use of APHA standard methods.

Results and Discussion pH was near to neutral, whereas other parameters were recorded average mean values of TSS (287.6 mg/l), TDS (2673.2 mg/l), COD (6899.16 mg/l), BOD (1403.75 mg/l), O & G (5.95 mg/l), PO4 (29.61 mg/l), Cu (12.58 mg/l), Cr (8.71 mg/l), Zn (19.15 mg/l), Fe (71.95 mg/l), Pb (3.92 mg/l), and Ca (0.53 mg/l) were recorded, respectively (Table 2). Dissolve oxygen (DO) was totally absent in the inlet effluent of automobile industries; the 12-month average data is shown in Table 2. After the analysis of the inlet effluent of automobile industries, the results show that each parameter concentration is very high as compared to MPCB norms. Therefore, before biological treatment, chemical treatment is very essential. In general practice, automobile industries use sodium hydroxide (NaOH) and BTS-809 polyelectrolyte (cation) for treating the effluent in order to remove the impurities from the effluent. This practice shows that it is not a satisfactory combination for the effluent treatment, because sodium compound increases the salinity and yellowish color of the effluent. Treated effluent with sodium hydroxide cannot be used for plantation and agricultural purposes due to the high salinity. The market cost of sodium hydroxide is 25 Rs. per kg. Therefore, during investigation again, it is tried with another alternative treatment with more effective results, i.e., Ca(OH)2 and BTS-809 polyelectrolyte (cation). During work, 50 mg/l of Ca(OH)2 (conc. 5%) and 10 mg/l of BTS-809 polyelectrolyte (conc. 0.1%) are used for the treatment of the effluent. Immediate separation was observed. The cost of calcium hydroxide in the market is 5 Rs. per kg. Calcium compound contributed to lowered total dissolved solids (TDS) level as compared to sodium hydroxide. Therefore, treated effluent with calcium hydroxide can be used for various purposes such as plantation, cleaning, washing, etc. The inlet effluent of automobile industries is highly dark blackish, reddish, and whitish in color due to the presence of paint particles and other impurities like oil and grease, phosphate, and heavy metals like Cr, Cu, Zn, Pb, and Fe. During the course of chemical treatment, the pH of the effluent is increased to 9.5 up to 10.5 due to the addition of Ca(OH)2 some chemical alteration took place. The hydroxide groups from calcium are attached to other heavy metals like Zn, Cu, Cr, Pb, and Fe, and get converted into hydroxide compounds like zinc hydroxide, chromium hydroxide, copper hydroxide, lead hydroxide, and iron hydroxide, etc. within 1 minute the polyelectrolyte is added to the effluent.

Months Nov./07 Dec./07 Jan./08 Feb./08 Mar./08 April/08 May/08 June/08 July/08 Aug./08 Sep./08 Oct./08 Average

pH 6.79 7.06 7.39 7.91 6.69 7.73 7.15 6.57 7.82 7.48 7.23 6.86 7.22

TSS mg/l 236 279 310 347 300 280 297 273 317 278 254 281 287.6

TDS mg/l 2750 2833 2910 2521 2571 2784 2879 2667 2457 2770 2345 2592 2673.2

Table 2 Inlet effluent characteristics

COD mg/l 7480 5244 9040 7082 6680 8496 5632 6248 9320 7084 5042 5322 6899.16

BOD mg 1438 1275 1510 1224 1372 1480 1256 1530 1438 1532 1422 1368 1403.75

O & G mg/l 5.73 8.43 6.51 7.01 5.68 3.79 7.21 4.71 5.72 6.40 5.50 4.73 5.95

PO4 mg/l 28.49 30.12 34.62 37.10 27.42 34.63 19.79 32.20 28.34 23.17 33.08 26.43 29.61 Cu mg/l Cr mg/l 10.38 9.30 15.33 10.41 9.03 9.43 11.42 8.28 10.72 9.75 20.08 7.55 13.19 8.39 12.89 7.23 8.19 9.71 16.11 8.19 13.20 8.87 10.45 7.50 12.58 8.71

Zn mg/l 20.07 22.07 19.41 18.77 20.19 17.28 19.07 17.45 18.43 17.26 19.39 20.43 19.15

Fe mg/l 80.58 74.01 77.14 70.47 80.31 83.70 55.73 63.67 76.23 79.14 64.39 58.03 71.95

Pb mg/l 4.38 5.62 3.47 5.22 2.87 3.66 4.50 5.61 2.07 3.89 3.37 2.40 3.92

Ca mg/l 0.50 0.57 0.60 0.53 0.48 0.55 0.51 0.54 0.49 0.63 0.44 0.59 0.53

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Polyelectrolyte helps to adhere impurities such as paint particles, oil and grease, phosphate, etc., and each hydroxide group such as zinc hydroxide, chromium hydroxide, copper hydroxide, lead hydroxide, and iron hydroxide is attached and get converted into the flocs. The flocs are easily separated out from the effluent by gravitational force; with the help of this force, the flocs are settled at the bottom of the reaction tank, and finally get cleaned and become alkaline effluent. The alkaline effluent is neutralized by the addition of dilute sulfuric acid, i.e., in the ratio of 1:1. After neutralization, the effluent is used as a feeding effluent in the bioreactor for biological treatment. The 12-month data is exhibited in Table 3. The average mean values of the 12-month data of each parameter like pH (7.46), TSS (119.1 mg/l), TDS (1457.50 mg/l), COD (1655.6 mg/l), BOD (629.5 mg/l), DO (0.37), O & G (0.86 mg/l), PO4 (4.87), Cu (2.91), Cr (2.60 mg/l), Zn (2.69 mg/l), Fe (9.08 mg/l), Pb (0.67 mg/l), and Ca (0.34 mg/l) are recorded, respectively. Due to the chemical treatment, approximately 65% impurities load is reduced from the inlet effluent. After this treatment, still some concentration of metals is found in the effluent. Therefore, these effluents are again treated by biological way. Chemical sludge is generated after chemical treatment. The quantity of the sludge and heavy metal concentration are analyzed from the chemical sludge; the 12-month mean values average data are given in the table no. 3.2. The average mean values of the chemical sludge per month is 45.38 gm, pH of the sludge is 10.06, oil and grease (5.07 mg), PO4 (27.23 mg), Cu (9.67 mg), Cr (6.11 mg), Zn (16.45 mg), Fe (62.84 mg), Pb (3.24 mg), and Ca (0.23 mg) are recorded, respectively. To study the efficiency of biological treatment of the feeding effluent of automobile industries, two pilot plants were set up at a lab scale: one was the conventional bioreactor plant and another was the novel bioreactor with modified design concept. High-density polyethylene (HDP) sheets were used for both bioreactors (conventional and novel), and the thickness of the sheet was 5 mm. Height (30 cm), width (20 cm), and length (20 cm) were the same for both bioreactors, but the surface area of conventional bioreactor was 400 sq. cm and the surface area of novel bioreactor was 313.76 sq. cm, because 86.24 sq. cm. was covered by baffles. Inside of novel bioreactor, two types of baffles were designed as half circular and small square. Half circular baffles covered 50.24 sq. cm and small square baffles covered 36 sq. cm area from the total surface area, i.e., 400 sq. cm, but the working volume of both bioreactors was maintained the same, i.e., 7 L; in the same way, the feeding volume in each bioreactor was also maintained the same, i.e., 2 L. After maintaining working and feeding values the same of both bioreactors than automatically other operating parameters as the flow rate of the effluent in bioreactor, retention time of both bioreactors was also maintained the same (Table 4). The position of stirrer and impeller is shown in Fig. 2a, b. 2800 rpm motors were used for aeration in both bioreactors, but the impeller positions of the stirrer were different. In the novel bioreactor, two impellers of the stirrer were used (the first impeller position was 6 cm and the second impeller position was 13.5 cm from the bottom of the reactor), whereas in the conventional bioreactor, only one impeller was used (its position was 6 cm from the bottom of the reactor).

Months Nov./07 Dec./07 Jan./08 Feb./08 Mar./08 April/ 08 May/08 June/08 July/08 Aug./08 Sep./08 Oct./08 Average

TSS mg/l 115 120 146 109 99.0 103

110 130 124 94.0 137 143 119.16

pH 7.43 7.56 7.31 7.73 7.50 7.04

7.49 7.57 7.12 7.84 7.30 7.63 7.46

1468 1410 1396 1471 1384 1359 1457.50

TDS mg/l 1325 1389 1480 1505 1573 1530

1842 1508 1824 1744 1564 1642 1655.66

COD mg/l 1480 1672 1542 1712 1704 1634

Table 3 Feeding effluent characteristics

741 560 530 588 624 574 629.58

BOD mg/l 560 678 800 720 487 693 0.42 0.30 0.45 0.34 0.38 0.26 0.37

DO mg/l 0.48 0.51 0.33 0.37 0.40 0.29 0.46 1.32 0.12 0.56 1.03 0.63 0.86

O & G mg/l 1.73 0.47 0.76 1.35 0.92 1.20 4.50 3.68 2.94 5.07 6.45 4.73 4.87

PO4 mg/l 6.23 3.71 5.06 6.15 5.23 4.79 2.49 2.53 3.30 3.09 2.87 2.26 2.91

Cu mg/l 3.19 3.23 2.78 3.36 2.81 3.04 2.60 2.51 2.03 2.82 2.20 2.73 2.60

Cr mg/l 3.56 2.84 3.05 2.37 2.08 2.49 3.56 2.83 2.94 3.05 2.13 2.06 2.69

7.76 9.20 10.37 8.19 7.80 8.14 9.08

Zn mg/l Fe mg/l 2.31 10.34 3.27 9.54 2.34 9.40 2.68 10.32 2.72 8.56 2.44 9.59

0.71 0.80 0.73 0.60 0.50 0.68 0.67

Pb mg/l 0.83 0.49 0.83 0.35 0.94 0.65

0.32 0.34 0.36 0.39 0.40 0.36 0.34

Ca mg/l 0.40 0.37 0.33 0.38 0.35 0.41

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Table 4 Design aspect and operating parameters of both bioreactors Sr. no. 1. 2. 3. 4. 5. 6. 7. 8.

Design aspect and operating parameters of bioreactor Feeding capacity of effluent per day Flow rate of effluent Retention time of bioreactor Working effluent capacity of each bioreactor Operating hours Inner surface of bioreactor Number of stirrer Impeller position of stirrer

Conventional bioreactor 2 lit. 83.33 ml/hrs. 7 days 7 lit. 24 Non baffles 1 (1 impeller) Only surface

9.

Motor rpm

2800

Novel bioreactor 2 lit. 83.33 ml/hrs. 7 days 7 lit. 24 Baffles 1 (2 impeller) Surface and bottom 2800

Fig. 2 (a and b) shows the surface aerator of both bioreactors

The related opinion was also given by many workers for the treatment of wastewater (Brindle et al. 1996). In the novel bioreactor, inside baffles are constructed, and two impellers are used: one at the surface and the other at the bottom. Due to these differences in the bioreactor system, changes in the effluents are recorded in both bioreactors.

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The 12-month average mean values data of the novel bioreactor such as pH (7.43), SVI % (57.41), TDS (1268.83 mg/l), COD (320.58 mg/l), MLSS (4768.2 mg/ l), MLVSS (2412.75), DO (3.14 mg/l), OUR (0.81 mg/l), PO4 (1.18 mg/l), Cu (0.80 mg/l), Cr (0.24 mg/l), Zn (0.24 mg/l), Fe (2.28 mg/l), Pb (0.19 mg/l), and Ca (0.08 mg/l) were recorded, respectively, whereas the parameters of conventional bioreactor like pH (7.43), SVI % (47.91), TDS (1312.75 mg/l), COD (833.66 mg/l), MLSS (4062.75 mg/l), MLVSS (2204.83 mg/l), DO (1.56 mg/l), OUR (0.43 mg/l), PO4 (1.83 mg/l), Cu (0.91 mg/l), Cr (0.73 mg/l), Zn (0.48 mg/l), Fe (2.94 mg/l), Pb (0.29 mg/l), and Ca (0.13 mg/l) were recorded, respectively. Similar findings were also reported by many workers from the metal-contaminated water (Anderson and Smith 1987; Dvorak et al. 1992; Allard and Neilson 1997; Ansola et al. 2003). Some parameters like SVI (sludge volume index), MLSS (mixed liquor suspended solids), MLVSS (mixed liquor volatile suspended solids), DO, and OUR (oxygen uptake rate) were studied in order to recognize the microbial growth rate in both bioreactors. In the novel bioreactor, SVI was recorded as 57.41%, MLSS 4768.25 mg/l, MLVSS 2412.75 mg/l, DO 3.14 mg/l, and OUR 0.81 mg/l, respectively, whereas in the conventional bioreactor, SVI was recorded as 47.91%, MLSS 4062.75 mg/l, MLVSS 2204.83 mg/l, DO 1.56 mg/l, and OUR 0.43 mg/l, respectively. This difference indicated that the bacterial growth rate in the novel bioreactor was more than the conventional bioreactor. The study concluded that two impellers and baffles assisted in increasing the DO rate in the effluent; when the dissolved oxygen rate increased in the effluent simultaneously, the bacterial growth rate also increased in the reactor. Proper mixing of the effluent and bacterial biomass is very essential during the biological treatment. The bottom impeller of the stirrer helps to mix the bacterial biomass in the effluent uniformly in the novel bioreactor. The baffles and the bottom impeller are absent in the conventional bioreactor. Therefore, the bottom effluent and the biomass remain unchanged, and hence the anaerobic condition is revealed. Therefore, the rate of bacterial growth decreases. When the parameters like SVI, MLSS, MLVSS, DO, and OUR are at increasing side, then the concentration of COD, BOD, PO4, and heavy metals shows towards decreasing. Some experts have also reported the same finding in the mineral hydrometallurgy and from oil refinery streams (Crundwell 2003; Babich and Moulijn 2003). The treated effluent sample from both bioreactors were collected and analyzed. The outlet effluent quality of the novel bioreactor is analyzed, and their parameter results such as pH (7.30), TSS (31.83 mg/l), TDS (1236.0 mg/l), COD (68.25 mg/l), BOD (27.16 mg/l), DO (2.92 mg/l), O & G is BDL (Below detection limit), PO4 (0.25 mg/l), Cu (0.14 mg/l), Cr (0.17 mg/l), Zn (0.19 mg/l), Fe (0.31 mg/l), Pb (0.02 mg/l), and Ca (0.01 mg/l) are recorded respectively, at the same time, the outlet effluent of the conventional bioreactor is also analyzed, and the results are pH (7.36), TSS (46.75 mg/l), TDS (1286.41 mg/l), COD (618.58 mg/l), BOD (98.66 mg/l), DO (1.36 mg/l), O & G is BDL (below detection limit), PO4 (1.74 mg/l), Cu (1.0 mg/l), Cr (1.06 mg/l), Zn (1.01 mg/l), Fe (3.33 mg/l), Pb (0.23 mg/l), and Ca (0.12 mg/l) are recorded respectively. The 12-month average values of both bioreactor data are illustrated in Tables 5 and 6.

April/ 08

Mar./ 08

Feb./08

Jan./08

Dec./ 07

Months Nov./ 07

Bioreactors name Conventional bioreactor Novel bioreactor Conventional bioreactor Novel bioreactor Conventional bioreactor Novel bioreactor Conventional bioreactor Novel bioreactor Conventional bioreactor Novel bioreactor Conventional bioreactor Novel bioreactor

40

26

7.39

34

7.20

7.04

47

39

7.50

7.41

66

30

7.24

7.45

58

32

7.33

7.18

40

7.36

1368

1442

1368

1416

1243

1308

1200

1228

1120

1164

68

615

58

660

80

674

90

642

120

698

87

20

132

18

130

22

162

22

164

20

148

18

1208

7.15

37

BOD mg/l 156

Outlet effluent characteristics TSS TDS COD pH mg/l mg/l mg/l 7.20 43 1253 684

2.82

1.35

3.0

1.15

3.03

1.40

3.23

1.37

2.98

1.39

2.56

DO mg/l 1.29

Table 5 Monthly average mean values of the outlet effluent of both bioreactors

BDL

BDL

BDL

BDL

BDL

BDL

BDL

BDL

BDL

BDL

BDL

O &G mg/l BDL

0.21

1.70

0.19

2.14

0.36

2.12

0.42

1.75

0.38

1.19

0.48

PO4 mg/l 3.26

0.15

1.05

0.09

0.91

0.06

1.10

0.23

1.02

0.42

1.16

0.23

Cu mg/l 1.07

0.09

0.75

0.17

0.94

0.23

1.28

0.28

1.36

0.32

1.54

0.38

Cr mg/l 2.0

0.15

0.80

0.24

1.08

0.28

0.96

0.30

1.02

0.18

1.16

0.48

Zn mg/l 1.20

0.31

3.50

0.12

2.79

0.06

3.53

0.64

4.36

0.07

4.20

0.54

Fe mg/l 3.21

BDL

0.20

BDL

0.23

0.04

0.14

0.09

0.39

BDL

0.17

BDL

Pb mg/l 0.38

BDL

0.09

BDL

0.13

0.03

0.10

BDL

0.14

0.02

0.11

0.03

Ca mg/l 0.21

73 Biotechnological Approach for Mitigation Studies of Effluents of. . . 1831

Average

Oct./08

Sep./08

Aug./08

July/08

June/08

Months May/08

Bioreactors name Conventional bioreactor Novel bioreactor Conventional bioreactor Novel bioreactor Conventional bioreactor Novel bioreactor Conventional bioreactor Novel bioreactor Conventional bioreactor Novel bioreactor Conventional bioreactor Novel bioreactor C.B N.B

46.75 31.83

27

7.29

7.36 7.30

39

28

7.20

7.63

40

35

7.34

7.30

56

33

7.42

7.84

49

7.12

31

7.37

7.27

38

30

pH 7.30

7.57

TSS mg/l 45

1286.41 1236.08

1181

1251

1068

1139

1315

1360

1203

1235

1210

1250

1354

TDS mg/l 1391

618.58 68.25

51

526

60

545

50

564

57

592

46

570

52

COD mg/l 653

Outlet effluent characteristics

98.66 27.16

50

62

26

48

36

52

38

50

30

44

24

BOD mg/l 36

Table 6 Monthly average mean values of the outlet effluent of both bioreactors

1.36 2.92

3.07

1.28

2.92

1.10

2.75

1.64

2.88

1.53

2.97

1.60

2.94

DO mg/ l 1.27

BDL BDL

BDL

BDL

BDL

BDL

BDL

BDL

BDL

BDL

BDL

BDL

BDL

O &G mg/l BDL

1.74 0.25

0.13

1.27

0.34

2.14

0.20

1.48

0.06

0.70

0.14

1.73

0.16

PO4 mg/l 1.42

1.00 0.14

BDL

1.13

0.22

1.04

0.10

1.12

0.13

1.02

0.03

0.72

0.11

Cu mg/l 0.90

1.06 0.17

0.11

0.80

0.18

0.87

0.14

0.93

0.03

0.65

0.13

0.84

0.08

Cr mg/l 0.78

1.01 0.19

0.12

0.90

0.09

0.85

0.10

1.09

0.08

0.92

0.10

0.97

0.22

Zn mg/l 1.23

3.33 0.31

0.30

2.89

0.36

2.93

0.32

3.07

0.35

3.81

0.28

2.89

0.45

Fe mg/l 2.80

0.23 0.02

BDL

0.18

BDL

0.20

0.03

0.24

0.04

0.29

0.06

0.27

BDL

Pb mg/l 0.16

0.12 0.01

BDL

0.10

0.01

0.12

0.03

0.14

BDL

0.09

0.04

0.11

0.02

Ca mg/l 0.12

1832 N. N. Bandela et al.

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Biotechnological Approach for Mitigation Studies of Effluents of. . .

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The bio-sludge was generated after biological treatment. The quantity of the bio-sludge and heavy metal concentration were analyzed from the sludge. The average mean value of the novel bioreactor biomass sludge per month was 0.133 gm, PO4 (3.43 gm), Cu (1.96 gm), Cr (2.18 gm), Zn (2.23 gm), Fe (6.50 gm), Pb (0.46 gm), and Ca (0.25 gm) were recorded respectively, whereas conventional bioreactor biomass per month was 0.059 gm, PO4 (1.30 gm), Cu (0.97 gm), Cr (0.81 gm), Zn (1.19 gm), Fe (2.82 gm), Pb (0.14 gm), and Ca (0.11 gm) difference in results were recorded respectively. Some of the authors opined that there is continuous reduction in excess sludge production from the activated sludge process (Liu and Tay 2001; Curds and Cockburn 1970). The percent reductions of the conventional and novel bioreactors were TDS (10.93%), COD (63.05%), PO4 (65.10%), Cu (64.89%), Cr (61.06%), Zn (63.41%), Fe (64.43%), Pb (64.86%), and Ca (66.26%) recorded respectively; similarly, the percentage mean values of the novel bioreactor parameters were TDS (14.51%), COD (96.12%), PO4 (96.03%), Cu (94.45%), Cr (93.83%), Zn (93.72%), Fe (95.72%), Pb (94.95%), and Ca (94.15%) recorded respectively. After the investigation of both bioreactor effluents, it was finally concluded that the treatment efficiency of the novel bioreactor was between 29.56 and 33.07%, which was more than the conventional bioreactor. Seviour et al. (2003) and Kong et al. (2007) also reported the reduction of phosphorus from the activated sludge system. Today, various industries are well acquainted with diffuse bioreactor, because its efficiency is six times more than the surface (conventional) bioreactor. Even though it is feasible, industries have neglected the diffuse bioreactor, because of its high initial cost and maintenance cost. During the survey, it was observed that all automobile and electroplating industries were engaged with surface bioreactor (conventional bioreactor). To improve the efficiency of outlet effluent, a novel bioreactor has been designed during the work. After the comparative study, it is finally concluded that the novel bioreactor efficiency is two times more than the conventional bioreactor. Hence, it is recommended that novel bioreactors can play a vital role in treating the effluent of automobile industries. Then novel bioreactor will maintain the middle stage between the surfaces and the diffuse reactor. It is simple for installation and maintenance, and it is also economically reliable for operation. It too helps in oxygen transfer and for mixing thoroughly. The microbe of the activated sludge helps to adsorb various heavy metals from the effluent. Therefore, during the work, some selective species of the bacteria such as Pseudomonas aeruginosa, Pseudomonas fluorescens, Alcaligenes faecalis, Enterobacter aerogenes, Bacillus subtilis, Aeromonas hydrophila, Agrobacterium tumefaciens, Escherichia coli, and Zoogloea ramigera; fungi such as Aspergillus niger; plankton such as Scenedesmus, Ankistrodesmus, Paramecium, and Arthrodesmus; and diatoms such as Navicula were identified from the activated sludge of both bioreactors of automobile industries. Utgikar et al. (2000), Ganguli and Tripathi (2002), Curtin (1983), Bozkurt et al. (2008), Andrade et al. (2003), Volesky and Holan (1995), and Abbas (2006) – all these workers also identified some species of microbes like P. fluorescens,

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Alcaligenes faecalis, Enterobacter aerogenes, Bacillus subtilis, Aeromonas hydrophila, Agrobacterium tumefaciens, Escherichia coli, and Zoogloea ramigera and fungi such as Aspergillus niger isolated from the acid mine drainage as well as from heavy metal effluents. It is found that after the biological treatment, most of the heavy metals were reduced in the outlet effluent from automobile industries. The initial feeding concentration of heavy metals like Cu (2.91 mg/l), Cr (2.60 mg/l), Zn (2.69 mg/l), Fe (9.08 mg/l), Pb (0.67 mg/l), and Ca (0.34 mg/l) and at outlet effluents results of novel bioreactor were Cu (0.14 mg/l), Cr (0.17 mg/l), Zn (0.19 mg/l), Fe (0.31 mg/l), Pb (0.02 mg/l), and Ca (0.01 mg/l) are recorded respectively. It seems that the remaining concentration of heavy metals is accumulated in microbes and in the algal biomass. Many workers have also reported that these same species are highly useful in the reduction of heavy metal concentration by the bioremediation method from the wastewater – Das et al. (2008), DeLeo and Ehrlich (1994), Gonzalez et al. (2005), McLean and Beveridge (2001), and Ray and Ray (2009). Most prominently from algal species like Scenedesmus, Ankistrodesmus and Navicula species are found from bioreactor sludge. Interesting results were recorded during the analysis of bio-sludge. It seems that these identified algal species along with some microbes are helping in the removal of phosphorus and also useful in the progressive reduction in COD analysis. The average mean values of PO4 (96.03%) and COD (96.12%) are recorded respectively. Hammouda et al. (2002), Bojarajan et al. (2007), and Chang et al. (2004) also reported the same from microalgae aquaculture wastewater treatment. The algal species which are identified during work shown interesting relation in the bio-accumulation of oil, special Ankistrodesmus species in the inlet feeding effluents (0.86 mg/l) after the process it is below detection limit with the supporting references of Ben-Amotz and Tornabene (1985) and Castanier and Brigham (2003) from the wastewater. All species of the bacteria, fungi, plankton, and diatoms play a vital role in the treatment of the effluent of automobile industries. Out of the identified bacterial species, it is noted that P. aeruginosa was found abundant in the effluent of automobile industries.

Conclusion The automobile industry is a metallurgic and engineering industry, wherein water is used in huge quantities for coolant preparation, surface preparation for electroplating, surface preparation for phosphate coating, washing and cleaning processes, and surface preparation for painting purposes. Such processes result in the generation of enormous water pollution load each day by the automobile industry. The generated effluents from the industry are highly contaminated with various heavy metals like Zn, Ca, Pb, Ni, Cr, and Fe, paint particles, coolants, phosphate coating, oil and grease, etc. Further, the discharge of such toxic effluents without any treatment would contaminate natural water bodies. The preliminary survey prior to the study

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Biotechnological Approach for Mitigation Studies of Effluents of. . .

1835

revealed that the effluent from industries was treated only by chemical methods and then was directly discharged in the water bodies. However, the region faces water scarcity so huge quantities of this wastewater post treatment could be reused for washing, cleaning, and plantation purposes. And clearly subsequent to chemical treatment, the application of biotechnological treatment should be undertaken to further reduce the pollution load. The physicochemical methods such as chemical precipitation, chemical oxidation or reduction, electrochemical treatment, evaporative recovery, filtration, ion exchange, and membrane technologies are expensive, and even after the application of all these methods, traces of heavy metals still remain within the effluent. Therefore, the biological methods such as bioaccumulation for the removal of heavy metal ions are an attractive alternative. Effluent treated with sodium hydroxide cannot be used for plantation and agricultural purposes due to the high salinity and cost. Therefore, during investigation, another alternative treatment with more effective results, i.e., Ca(OH)2 and BTS-809 polyelectrolyte (cation), was used. Calcium compound was cheaper and contributed to lowered total dissolved solids (TDS) as compared to sodium hydroxide. Therefore, treated effluent with calcium hydroxide can be used for various purposes such as plantation, cleaning, washing, etc. To study the efficiency of biological treatment of the feeding effluent of automobile industries, two pilot plants were set up at a lab scale: one was the conventional bioreactor plant and another was the novel bioreactor with modified design concept. In the novel bioreactor, inside baffles are constructed, and two impellers are used: one at the surface and the other at the bottom. The study concluded that two impellers and baffles assisted in increasing the DO rate in the effluent; when the dissolved oxygen rate increased in the effluent simultaneously, the bacterial growth rate also increased in the reactor. Proper mixing of the effluent and bacterial biomass is very essential during the biological treatment. The bottom impeller of the stirrer helps to mix the bacterial biomass in the effluent uniformly in the novel bioreactor. The baffles and the bottom impeller were absent in the conventional bioreactor. After the comparative study, it was finally concluded that the novel bioreactor efficiency was two times more than the conventional bioreactor. Hence, it is recommended that the novel bioreactors can play a vital role in treating the effluent of automobile industries. The novel bioreactor will maintain the middle stage between the surfaces and the diffuse reactor. It is simple for installation and maintenance, and it is also economically reliable for operation. It too helps in oxygen transfer and for mixing thoroughly. The microbe of the activated sludge helps to adsorb various heavy metals from the effluent. Therefore, during the work, some selective species of the bacteria such as P. aeruginosa, P. fluorescens, Alcaligenes faecalis, Enterobacter aerogenes, Bacillus subtilis, Aeromonas hydrophila, Agrobacterium tumefaciens, Escherichia coli, and Zoogloea ramigera were found within the activated sludge of bioreactors. Out of the identified bacterial species, it was noted that P. aeruginosa was found abundant in the effluent of automobile industries.

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Cross-References ▶ Application of Novel Microbial Consortia for Environmental Site Remediation and Hazardous Waste Management Toward Low- and High-Density Polyethylene and Prioritizing the Cost-Effective, Eco-friendly, and Sustainable Biotechnological Intervention ▶ Biostimulation and Bioaugmentation: An Alternative Strategy for Bioremediation of Ground Water Contaminated Mixed Landfill Leachate and Sea Water in Low Income ASEAN Countries ▶ Decentralized Integrated Approach of Water and Wastewater Management in Rural West Bengal ▶ Micro-remediation of Metals: A New Frontier in Bioremediation ▶ Remediation of Industrial and Automobile Exhausts for Environmental Management

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Crundwell FK (2003) How do bacteria interact with minerals? Hydrometallurgy 71:75–81 Curds CR, Cockburn A (1970) Protozoa in biological sewage-treatment processes—I. A survey of the protozoan fauna of British percolating filters and activated-sludge plants. Water Res 4:225–236 Curtin ME (1983) Microbial mining and metal recovery. Biotechnology 1:228–235 Das N, Vimala R, Karthika P (2008) Biosorption of heavy metals – an overview. Indian J Biotechnol 7:159–169 Davies G (2003) Materials for automobile bodies. Butterworth-Heinemann, Amsterdam, p 157 Degarmo EP, Black JT, Kohser RA (2003) Materials and processes in manufacturing, 9th edn. Wiley, Upper Saddle River, p 793 DeLeo PC, Ehrlich HL (1994) Reduction of hexavalent chromium by Pseudomonas fluorescens LB300 in batch and continuous cultures. Appl Microbiol Biotechnol 40:756–759 Dvorak DH, Hedin RS, Edenborn HM, McIntire PE (1992) Treatment of metal-contaminated water using bacterial sulfate reduction: results from pilot-scale reactors. Biotechnol Bioeng 40:609–616 Ganguli A, Tripathi AK (2002) Bioremediation of toxic chromium from electroplating effluent by chromate-reducing Pseudomonas aeruginosa A2Chr in two bioreactors. Appl Microbiol Biotechnol 58:416–420 Gonzalez AR, Ndung'u K, Flegal AR (2005) Natural occurrence of hexavalent chromium in the aromas red sands aquifer, California. Environ Sci Technol 39(15):5505–5511 Hammouda O, Gaber A, Abdelraouf N (2002) Microalgae & wastewater treatment. Ecotoxicol Environ Saf 31:205–210 Kapoor A, Viraraghavan T (1995) Fungal biosorption- an alternative treatment option for heavy metal bearing wastewater: a review. Bioresour Technol 53(3):195–206 Kong Y, Xia Y, Nielsen JL, Nielsen PH (2007) Structure and function of the microbial community in a full-scale enhanced biological phosphorus removal plant. Microbiology 153:4061–4073 Joseph K, Nagendran R (2004) Essentials of environmental studies. Pearson Education Pvt. Ltd., Delhi, pp 123–137 Liu Y, Tay JH (2001) Strategy for minimization of excess sludge production from the activated sludge process. Biotechnol Adv 19:97–107 McLean J, Beveridge TJ (2001) Chromate reduction by a pseudomonad isolated from a site contaminated with chromated copper arsenate. Appl Environ Microbiol 67:1076–1084 Metcaff and Eddy (2001) Wastewater engineering treatment disposal reuse, 3rd edn. Tata McGrawHill Publishing Company Limited, New Delhi Ray SA, Ray MK (2009) Bioremediation of heavy metal toxicity-with special reference to chromium. Al Ameen J Med Sci 2(2 Special):57–63. ISSN:0974–1143 Rubin H (2001) Feasibility of on-site bioremediation of loam soil contaminated by diesel oil. J Environ Sci Health A Tox Hazard Subst Environ Eng 36(8):1549–1558 Seviour RJ, Mino T, Onuki M (2003) The microbiology of biological phosphorus removal in activated sludge systems. FEMS Microbiol Rev 27:99–127 Utgikar V, Chen BY, Tabak HH, Bishop DF, Govind R (2000) Treatment of acid mine drainage: I. Equilibrium biosorption of zinc and copper on non-viable activated sludge. Int Biodeter Biodegr 46:19–28 Volesky B (1990) Removal and recovery of heavy metals by biosorption. In: Biosorption of heavy metals. CRC Press, Boston, pp 7–43 Volesky B, Holan ZR (1995) Biosorption of heavy metals. Biotechnol Prog 11:235–250

New Techniques for Treatment and Recovery of Valuable Products from Olive Mill Wastewater

74

Reda Elkacmi and Mounir Bennajah

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . COD, Coloration, and Polyphenols Removal from Olive Mill Wastewater by Electrocoagulation: Equilibrium and Kinetic Studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Effect of Experimental Parameters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Kinetic Study . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Variable Order Kinetic Approach (VOK) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A New Protocol for Isolation and Purification of Fatty Acids and Other Components from Olive Mill Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Method Details . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cross-References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

1840 1844 1844 1845 1845 1846 1846 1855 1855 1855

Abstract

The olive oil industries produce large quantities of wastewater having an enormous amount of pollutants that provide a deleterious effect on environment drastically if discharged without proper treatment. Despite its treatment throughout the extraction process, this liquid waste still contains a very important oily residue, always considered as a pollutant waste. In this context, a new upgrading technique has been developed for the treatment and valorization of olive mill R. Elkacmi (*) Department of Chemistry and Valorisation, Faculty of Sciences Ain-Chock, HASSAN II University of Casablanca, BP, Casablanca, Morocco Department of Process Engineering, National School of Mineral Industries of Rabat, BP, Rabat, Morocco M. Bennajah Department of Process Engineering, National School of Mineral Industries of Rabat, BP, Rabat, Morocco © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_157

1839

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wastewater (OMW), to ensure a better environmental protection and to contribute to the improvement of the profitability of the oleicol field through the recovery of a multiple valuable products. This chapter reviews both the treatment and valuation of OMW: The first objective of this study is to describe the treatment of OMW by electrocoagulation (EC) in a stirred tank reactor (STR); a variable order kinetic (VOK) derived from the Langmuir-Freundlich equation was developed to simulate the kinetics of the detoxification of OMW with EC using bipolar aluminum electrodes. The results showed good agreement between the predictive equation and the experimental data. The second part of this work was intended to extract and separate two polyunsaturated free fatty acids, with high purity by a fractional crystallization with urea from OMW collected from three olive oil extraction processes (traditional, modern, and semimodern). A comparative study was investigated with different OMW samples. By centrifugation, the oily phase would be extracted from OMW, which is destined to be a platform to obtain two fatty acids (oleic and linoleic acid), and to ensure total valorization of our effluent, a complementary process was carried out for production of a high quality soap, glycerol, and polyphenols. Finally, a pilot plant was designed and developed for a full olive harvesting period (100 days), in one of the biggest olive production areas of Morocco, in order to carry out a socioeconomic study, which consists in studying the feasibility of OMW valuation project for the local olive oil extraction processes.

Keywords

Olive mill wastwater · Valuable products · Electrocoagulation · Variable Order Kinetic approach · Isolation and purification · Fatty acids · Polyphenols

Introduction The olive oil industries are large consumers of water and also produce large quantity of effluents called “olive oil mill wastewater” (OMW), having an enormous amount of pollutants which provide a deleterious effect on environment drastically if discharged without proper treatment. OMW is a heavy polluted liquid stream characterized by an acid pH value, black color, and very high chemical oxygen demand (COD) and biological oxygen demand (BOD) concentrations comprising polyphenols, polysaccharides, sugars, proteins, nitrogenous organics, tannins, and fats (Dhaouadi and Marrot 2010; Adhoum and Monser 2004). Processes used at olive oil production have two types as discontinuous pressing in the traditional mills or continuous centrifuging in modern units according to oil separation techniques (Dermeche et al. 2013; Elkacmi et al. 2017a, b). Continuous separation systems are two types as two-phase or three-phase in accordance with features of decanter and phase separation level. Both two-phase centrifugal and

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1841

three-phase centrifugal processes seem to produce larger quantities of OMW per unit of treated olives than the traditional ones (Sassi et al. 2006). For every tons of olives processed, an amount between 1 and 1.5 m3 of effluent is generated according to the production method (Paraskeva and Diamadopoulos 2006), and the annual quantity of effluent discharged by Moroccan industries exceeds 250,000 m3/year (Razouk et al. 2012). OMW also varies considerably in volume and concentration due to the type of olive, harvest season, production process, and conditions of operation; the main physicochemical characteristics of the OMW from olive oil extraction processes are summarized in Table 1. These liquid effluents not only present an environmental nuisance which should be mitigated or even neutralized, but also constitute a source of energy, of citizen activities contributing to the jobs creation. Waste management is one of the major concerns of sustainable development. Their valorization and the prevention of the reduction of their production are at the heart of many debates within the scientific community, local authorities, companies, and civil society. The valorization of by-products of the olive tree has become a necessity to increase overall efficiency, and to reduce the health risks and environmental impacts. There are various techniques available for prevention of pollution caused by OMW physical processes such as: lagooning (natural evaporation) (Benyahia and Zein 2003), forced evaporation (Fiestas Ros de Ursinos and Borja 1992), membrane filtration (El-Abbassi et al. 2013), biological treatments: aerobic digestion (Sayadi et al. 2000) and anaerobic treatment (Antonacci et al. 1981; Balice et al. 1988), and chemical processes: adsorption (Galiatsatou et al. 2002) and coagulation-floculation (Ginos et al. 2006). Several publications demonstrated that the electrocoagulation was employed successfully not only for the detoxification of OMW but also for the treatment of many types of wastes, such as chemical and mechanical polishing (Belongia et al. 1999), textile wastewaters (Bennajah et al. 2009), restaurant wastes (Chen et al. 2000), suspended particles (Matteson et al. 1995), oil refinery (Un et al. 2009), leachate (Meunier et al. 2006), mine wastes (Jenke and Diebold 1984), and heavy metals (Kabdaşli et al. 2009). Table 1 Physicochemical characteristics of raw OMW

pH Electrical conductivity (ms/cm) Total dry matter (g/L) Total suspended solids (g/L) Chemical oxygen demand (g/L) Biological oxygen demand (g/L) Total phenol (g/L)

Achak et al. (2009) 4.12 9.46 – 6.1  0.09 64.92  4.27 – 7.28  0.6

Mouncif et al. (1993) 4.73 18.6 16.9 4.99 224.1 98.18 64.78

Borja et al. (2006) 4–6.7 8–16 – 1–9 45–170 35–110 –

Morillo et al. (2009) 4.84 8.36 6.72 – 124.67 65 4.98

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The valorization of olive by-products and especially OMW is likely to contribute to improving the profitability of the olive sector. It also makes it possible, on the one hand, to solve to a large extent the problems posed by these effluents, which have a very high polluting power, and on the other hand, to help fill the forage deficit which is found mainly in North Africa and Middle East. Many biotechnological applications have been made to utilize these liquids that we mentioned; the most commonly used application are: composting (Paredes et al. 2002; Aviani et al. 2010), use as fertilizer (Roig et al. 2006; Vlyssides et al. 2004), use as animal feed (Hamdi 1993; Martilotti 1983), biogas production (Gelegenis et al. 2007; Demirer et al. 2000), extraction of products of interest in nutrition and health such as polyphenols (Frascari et al. 2016; Bertin et al. 2011), dietary fiber suspensions (Galanakis et al. 2010), lipase producing Bacillus pumilus (Ertuğrul et al. 2007), and volatile fatty acids (VFAs) (Scoma et al. 2016). These studies focus mainly on the aqueous part of OMW; however, the oily part remaining in these effluents attracts less scientific attention, especially since this fatty part is considered a rich source of natural products such as fatty acids. Today, as a result of population growth and the scientific and industrial revolution, several studies in chemistry and biology require large amounts of fatty acids with high purity (Hinton and Ingram 2000; Francoeur et al. 1990; Larrucea et al. 2001). The extraction of these acids and especially the most expensive ones such as oleic and linoleic acid has been the subject of several studies and by various techniques: Liquid–liquid extraction, classified as a rapid, cheap, and simple technique, has been used as an alternative method for oleic and linoleic acids recovery (Belfrage and Vaughan 1969; Chang and Gladstone 1965). The limitations of this operation are the low recovering efficiency when it comes to high and pure fatty acids. The poor selectivity toward some fatty acids has been reported to be a downside for this method. Fractional distillation has been excessively studied for separation of fatty acids. This technique is based on the chain length of the fatty acids and their boiling point. This technique is not satisfying when used for the separation of the unsaturated fatty acids with the same chain length; the unsaturated fatty acids are thus inseparable (Murray 1955; Loury 1968). Also, this method cannot separate oleic and linoleic acids, given their identical volatilities, and must always be followed by a complementary treatment. Inclusion with urea is one of the simple and efficient methods for the purification of fatty acids; it has demonstrated its ability to isolate oleic, linoleic, and other fatty acids (Keppler et al. 1959; Swern and Parker 1952). This process is characterized by its low cost, high efficiency, and the fewer materials and equipments it requires, as well as the low operation temperature for the extraction. Indeed the low operation temperature helps protect oleic and linoleic acid from oxidation. Most of these researches were carried out from olive and safflower oil, which makes them costly and limits their economic profitability because of the high cost of the raw materials used.

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In addition to fatty acids, the phenolic compounds are very abundant in the OMW and are the major responsible factors for its black color and polluting load. Not only are they crucial for the oil stability but also the source of its antioxidant properties (Brenes et al. 1999). Many factors are interfering on determination of their concentration in oils factors: geographical origin, processing technique, storage time, and irrigation system (De Ursinos and Padilla 1992). A particular focus was devoted to separate phenolic compounds from OMW by using adsorption (Soto et al. 2011), integrated membrane processing (Garcia-Castello et al. 2010), cloud point extraction (Katsoyannos et al. 2006), supercritical fluid extraction (Takaç and Karakaya 2009), cooling crystallization (Kontos et al. 2014), and liquid–liquid extraction (De Marco et al. 2007). The use of this last technique was preferred for its simplicity and convenience; it remains more advantageous not only because of its lower cost and its better yield, but also because it could be applied even in small, family-owned olive oil mills that exist in most Mediterranean countries. All the processes previously indicated were carried out from virgin oils, such as olive oil, safflower oil, and seed oil, neglecting the wealth of their waste, especially OMW, what makes them less effective, because of the high economic value of the raw materials used, this makes it possible to move toward a “green” separation technology and the development of nonpolluting and energy-saving processes. This chapter is therefore organized into two major parts: The first proposed process aims to evaluate the electrocoagulation treatment of OMW using aluminum electrode. The effect of three operational parameters, namely, electrolysis time, current density, and initial pH has been examined on COD, polyphenols, and dark color removal efficiency. In order to determine for the first time, the kinetic of the removal of pollutants which will be performed to estimate the time required for treatment. In addition, a modeling approach was conducted to simulate EC data for a better understanding of the mechanisms for OMW treatment. The second part is devoted to develop an innovative process to recover valuable products from these effluents that are discharged directly into the environment in huge amount without effective treatment. After separation by centrifugation of the oily phase contained in OMW, it was the source of extraction of two fatty acids (oleic and linoleic acids) with high purities, biodegradable soap, and glycerol as a coproduct. On the other hand, the aqueous part was used for the recovery of phenolic compounds. Based on the experimental results, a pilot plant was designed and developed for a full olive harvesting period (100 days), in the one of the biggest olive production areas of Morocco, in order to carry out a socioeconomic study, which consists in studying the feasibility of OMW valuation project for the three local olive oil extraction processes.

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COD, Coloration, and Polyphenols Removal from Olive Mill Wastewater by Electrocoagulation: Equilibrium and Kinetic Studies Electrocoagulation methods present a strong and innovative approach to in situ coagulation of dilute media by electro-oxidation of sacrificed electrode and destabilizing contaminants. The three basic steps of an electrocoagulation process are: (i) formation of coagulants by electrolytic oxidation of the sacrificial electrode; (ii) destabilization of the contaminants, suspension of particulates, and breakdown of emulsions; and (iii) aggregation of the destabilized phases to form flocs. Many studies have been carried out for the olive mill wastewater treatments using electrocoagulation, (Hanafi et al. 2010) which are capable of removing more than 70% of COD, polyphenols, and dark color. The optimum condition for treating OMW diluted 5 times was found at current density 250A/m2, NaCl concentration 2 g/L, initial pH 4.2, and 15 min electrolysis time. Good removals were already attained for COD about 76%, polyphenols about 95%, and dark color about 95% by Adhoum and Monser (2004). In 2007, Khoufi et al. (2007) proposed pretreatment process consisted in a combination of electrocoagulation and sedimentation; about 76.2%, 75%, and 71% of phenolic compounds, turbidity, and suspended solid, respectively, were eliminated after 3 days of settling, and the addition of electrocoagulation has eliminated 43% of COD and 90% of color. A new study on the detoxification of OMW was developed by Elkacmi et al. (2017c), which determined that treatment of olive mill wastewater using electrocoagulation (EC) technique was highly affected by electrolysis time, current density, and pH. In addition, a kinetic study was conducted for the first time to describe the removal rates of the polluting load. Furthermore, a variable order kinetic (VOK) model derived from the Langmuir-Freundlish equation was proposed to determine the kinetics of pollutant removal reactions with EC. Results showed that the model equations strongly fit the experimental concentrations of the three pollutants.

Effect of Experimental Parameters Treatment of olive mill wastewater (OMW) by electrocoagulation (EC) was investigated in a stirred tank reactor (STR), filled with 3.5 cm3 of fresh OMW. The active area of each electrode (aluminum) was 12  5 cm with a total area of 60 cm2. The distance between electrodes was 1 cm. Current density values (j) between 25 and 66.66 mA cm2were investigated, which corresponded to current (I = j S) in the range of 1.5–4 A. The effect of different influential parameters, namely, contact time, current density, and pH, was determined. The main results of this study are presented in Table 2. The results showed that electrocoagulation can remove 72.63% of COD, 92.53% of polyphenols, and 95.20% of dark color, with current density of 58.33 mA.cm2, just after 45 min of treatment without any pH adjustment (initial pH = 5.2).

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Table 2 Pollutants removal rate of COD, polyphenol, and dark color with solution volume = 350 cm3, initial pH = 5.2, and j = 58.33 mA/cm2 COD PP Coloration

Initial parameters 27.6 3.75 8.5

Removal efficiency (%) 72 93 95

Kinetic Study As mentioned above, none of previous studies modeled the kinetic mechanism for the detoxification of OMW. In this context, it should be remembered that the kinetic mechanism for pollutants removal using EC process is quite complicated. In addition, the nature of pollutants and their characteristics are quite different. The authors noted that the detoxification rate of the EC follows first order kinetics for COD and dark color removal, and the polyphenols reduction fits the pseudo second-order model with current dependent parameters. The results of the model could produce an R2 value of 0.9844 and 0.9728 for DCO and polyphenols, respectively, and an R2 value of 0.9669 was produced for decolorization.

Variable Order Kinetic Approach (VOK) The experimental adsorption equilibrium isotherms are useful for describing the adsorption capacity of a specific adsorbent. Moreover, the isotherm plays a vital role for the analysis and the design of adsorption systems as well as for model prediction. In order to examine the controlling mechanism of adsorption, two general purpose models and a modified combined model were used in an attempt to fit the experimental data: the Langmuir model, the Freundlich model, and the LangmuirFreundlich model. After representing the coefficient of the three models with the regression coefficients R2, Elkacmi et al. (2017c) concluded that the Langmuir-Freundlich model can be used to ensure a better representation of the experimental data of adsorption isotherms. VOK model (variable order kinetic) was developed for the first time by Hu et al. (2007), in order to best represent experimental results. The aim of the VOK is to describe as accurately as possible the adsorption kinetics for a comprehensive estimation of the time required for treatment due to the change in mass of adsorbent. By presenting a comprehensive mathematical model, all factors are taken into consideration. The VOK model was conducted for the first time to describe and study the EC mechanisms effect of detoxification of OMW in STR using aluminum electrodes (Elkacmi et al. 2017c). The retention time required (tN) to eliminate an acceptable concentration of pollutants (Ce) can be expressed as:

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 1n  ZFV 1 1n C tN ¼ ðC0  CÞ þ C φAl φC Iqmax kLF ð1  nÞ 0 Where Z, F, and V are the valence of the Al (Z = 3), Faraday’s constant, and the volume of the reactor (m3), respectively. ØC is the faradic yield or the current efficiency, ØAl is the efficiency of the formation of flocs, I is the applied current (A), qmax is the maximum pollutant adsorption, kLF is Langmuir-Freundlich constant (g/L), n is the constant which shows greatness of relationship between adsorbate and adsorbent (index of heterogeneity), and C0 is the initial concentration. The VOK model combined to the Langmuir Freundlish equation will be applied to the experimental data obtained, and the effect of current density was studied at optimal conditions (pHi = 5.2 and t = 45 min). The variations of pollutants concentration during EC and VOK model versus time at I = 1.5, 2, 2.5, 3, 3.5, and 4 A corresponding to a current densities of CD = 25, 33.33, 41.66, 50, 58.33, and 66.66 mA/cm2, respectively. Results showed that VOK model strongly fits the experimental data, demonstrating that detoxification kinetics could be simulated using the variable-order-kinetic approach (VOK) coupled with Langmuir-Freundlich adsorption isotherm.

A New Protocol for Isolation and Purification of Fatty Acids and Other Components from Olive Mill Wastewater The fractional crystallization was proposed as a simple, rapid, and green method for the extraction of a higher purity and polyunsaturated fatty acids; in this study, the olive mill wastewater (OMW) was used for the first time as the raw material to recover two expensive products (oleic and linoleic acids). After a separation of the oily phase contained in these effluents, the triglycerides forming this part were transformed into methyl esters, and then crystallized four times with urea at 4  C and 20  C for methyl oleate and 5  C for methyl linoleate. The originality of this method lies in: • Extraction of two fatty acids, oleic and linoleic acids with high purity 97.86% and 95.71%, respectively, from oily wastes. • A biodegradable soap and extra pure glycerin (>99%) were produced from the oily phase. • The aqueous phase sedimented by separation has been treated with ethyl acetate to recover the polyphenolic compounds with a yield of 68.5%.

Method Details Sampling Olive mill wastewater was collected from an olive extraction plant which uses a traditional process. No chemical additives were used during the olive oil production.

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OMW was stored in a closed plastic container of 20 L and kept in a dark place at room temperature.

Separation After the storage of the sample, the OMW can be separated either naturally or by centrifugation. Even if they have the same performance, but the slow and the area requirement may limit the first one. The centrifugation allowed us to accelerate the separation of oil and water phase. This separation technique is performed by filling 4 test tubes with 250 mL of OMW solution; they are then centrifuged at a speed of 4000 rpm during 30 min using a centrifuger. The oil, having a lower density than water, will be found above the water in the centrifuger. In order to finally isolate the oil, it will be collected using an automatic adjustable pipette. Oil percentage was determined using weight difference. Extraction of Oleic and Linoleic Acid from the Oily Phase Transesterification This technique is based on the transformation of the triglycerides to methyl esters (methyl oleate and methyl linoleate). To do so: • 200 mL of the separated oil sample was mixed with 500 mL of methanol. • After preparing a solution of sodium methoxide, by mixing of 1 g of sodium hydroxide with 100 mL of methanol, 100 mL was used as catalyzer. • To ensure a perfect homogenization of the mixture, the preparation is put under refluxing for 3 h at 60  C. After 3 h of natural separation in a separatory funnel, two distinct phases are obtained, an upper phase rich in methyl ester and a lower phase concentrated in glycerol. The lower portion is recovered from the bottom of the separatory funnel, and the separated upper phase is washed with hydrochloric acid to neutralize excess sodium hydroxide; it is rewashed a second time with 100 mL of distilled water to form two phases – a high phase rich in pure methyl esters and a second one rich in water and methanol. To maximize this separation efficiency and recover a maximum of methyl esters, the hexane was used in the aqueous phase (water and methanol) with three-stage separation and a rate of 25 mL, followed by a separation by settling after each extraction. After pure methyl esters recovered at the end of the first separation is added to that extracted by hexane, the mixture was concentrated in the rotatory evaporator. Finally, the resulting methyl esters were weighed to determine the yield and the efficiency of this transesterification. Crystallization with Urea The aim of this second step is to crystallize methyl oleate and methyl linoleate by four successive crystallizations. During the preparation of methyl oleate, the major

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portion of the product is either located in the filtrate or in the crystalline phase. And for the crystallization of methyl linoleate, this second product will be concentrated only in the filtered phase for the four separations, depending on the conditions of separation and dilution. • Crystallization of methyl oleate – 100 g of methyl esters with 100 g of urea were dissolved in 1 L of methanol, under gentle heating to promote solubilization. – The solution is cooled at 4  C overnight and then filtered on a Büchner filter, under vacuum of 1 bar maintained with a vacuum pump, to obtain two phases: filtrate F1 and crystalline C1. – In the filtrate F1, 200 g of urea was added while heating. The mixture was cooled to 4  C, to separate the crystals C2 that will be washed, and diluted with 1.5 L of methanol and filtered under the vacuum at 22  C (room temperature). – The product of the second separation F3 is mixed with 120 g of urea at a temperature of 4  C. The solution is rested overnight until the recovery of the crystalline phase C4. • Crystallization of methyl linoleate – 100 g of methyl esters with the same amount of urea were dissolved in 1 L of methanol under heating to dissolve the urea in methanol. After allowing the solution to stand overnight at 5  C, two phases F0 1 and C0 1 have been obtained. – After vacuum filtration and washing with cold methanol of the solid precipitate C0 1, the washing liquid was recovered into the filtrate phase F0 1. – The phase F0 1 was mixed with 100 g of urea under heating for solubilization, and after a night of separations at T = 5  C, two new phases C0 2 and F0 2 will be formed. – Even for the first separation, C0 2 was washed with 200 mL of methanol with recovering the washing liquid in the phase filtrate F0 2. – For a third separation, and in the same conditions of the temperature, 50 g of urea was mixed with F0 2 to have C0 3 and F0 3. – Finally, and under the same temperature of separation (5  C), 150 g of urea was mixed F0 3 and added to the washing liquid of C0 3, to give the final phase C0 4 concentrated in methyl linoleate. • Samples analysis by gas chromatography (GC) During separation, samples were analyzed using a gas chromatography VARIANT 304 CX whose characteristics are the following: Name: VARIANT 304 CX Column length: 50 m Stationary phase: Silica Carrier gas: Helium

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Helium flow rate: 1 ml/min Injection volume split: 0.5 μL T column: 210  C Injector temp: 190  C Detector: FID (flame ionization detector) In a round-bottom flask, 5 g of oil is weighed and mixed with 50 mL of methanol and 1 mL of a methanolic solution of potassium hydroxide 5% as catalyst. Afterwards, the refluxing operation is conducted until disappearance of the oil drop. 200 g of cold water is then added. In a separatory funnel, 40 mL of hexane was mixed with the flask contents, and after 30 min of settling, the upper layer (methyl esters) was recovered and washed three times, these esters will be filtered on a filter containing the anhydrous sodium sulfate. As for analysis, we inject 0.5 μL of the filtrate on CPG.

Purification of Final Phases To separate the two acids, all traces of urea must be eliminated, according to the following steps: • A hydrochloric acid solution (1%) was added, followed by a natural decantation. • The methyl ester rich phase is then separated with hexane and washed three times with pure water to remove the excess of acid. • The phase must be dried after by anhydrous sodium sulfate, and for the solvent, it will be evaporated under vacuum. • The resulting methyl esters were saponified, and the fatty acids were released by the addition of hydrochloric acid.

Production of Soap and Glycerin In this work, a biodegradable soap was produced from the oil separated from OMW; in order to do so: • 100 g of this oil was mixed with 100 mL of ethanol in the presence of 30% sodium hydroxide; the flask containing the reaction mixture was fit with a reflux condenser and heated for 4 h at 50  C. • The mixture was poured into a saturated salt solution of sodium chloride (NaCl). The aim of this step is the recovery of glycerin and removal of excess sodium hydroxide. • After a night of separation, the solid phase was filtered under vacuum using Büchner funnel through a filter cloth (40 μm). The soap was washed two times with 50 mL of distilled water. The resulting aqueous solution of the saponification and transesterification contains a significant amount of glycerin C3H8O3; this compound is a fairly high-value commercial chemical, historically valued at $0.60–$0.90/lb., that is primarily used in

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the manufacture of various foods and beverages, pharmaceuticals, cosmetics, and other personal care products. After recovery of the resulting aqueous phase of the saponification and transesterification, it is neutralized with concentrated hydrochloric acid (30%), and then by evaporation followed by simple distillation, glycerol or glycerin is then recovered.

Extraction of Polyphenols from the Aqueous Phase To ensure a total delipidation of the OMW, the aqueous phase represents a raw material, to recover an important amount of polyphenols, as follows: • 250 mL of this phase was mixed with 400 mL of hexane for 10 min and then separated by decantation allowing us to obtain two phases: an upper phase of hexane and lower defatted phase. • 250 mL of ethyl acetate was added to the defatted phase, the mixture was vigorously stirred for 20 min, and then placed into a separatory funnel for decantation. Two phases were obtained: An organic phase representing the top layer contains the ethyl acetate, and the second phase located at the bottom contains the rest of the OMW. • In order to remove the solvent from the upper phase and recover the pure polyphenols, the separated solution will be concentrated in a rotary evaporator under vacuum at 37  C. • The extraction was repeated successively four times for maximum recovery of phenolic compounds. Figure 1 shows a diagram of valuation of OMW including production of soap, glycerin and extraction of two expensive acids in the international market, and the recovery of polyphenols. After the separation of the oily phase remaining in the OMW, 50 L of sample gave 15.4 L of oil – this yield of 30.8% should not be neglected – proving that despite its treatment throughout the extraction process, this olive mill wastewater still contains a very important oily residue. Crude OMWW

Centrifugation

Organic phase

Soap

Glycerol

Oleic acid

Fig. 1 Diagram of valuation of OMW

Aqueous phase

Linoleic acid

Polyphenols

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The relative weight percentages of the lipid classes during the crystallization steps are presented in Tables 3 and 4. After four successive crystallizations, we are able to recover two final phases C4 and C0 4 rich in methyl oleate and linoleate with a purity of 97.86 and 95.71, respectively. The initial solution had an oleic acid purity of 68.22% and 20.31% purity of linoleic acid; the four crystallizations were carried out on the four filtered phases F0 1, F0 2, F0 3, and F0 4. Their concentrations in linoleic acid are respectively, 35.12%, 56.12%, 74.19%, and 95.71%, and for the preparation of oleic acid, the evolution of the concentration is 78.19%, 84.61%, 86.09%, and 97.86% for the four phases F1, C2, F3, and C4, respectively. For the crystallization of the two acids, the first separation has eliminated a large amount of saturated acids, Palmitic and Stearic acid; the addition of urea to the filtrate F1 and F0 1 caused the crystallization of unsaturated acids, oleic acid, and the rest of the saturated acids. The crystalline phase C2 and after dilution with the methanol produced F3 (86.09%) and C3(79.21%) rich in oleic acid; for the separation of linoleic acid, the addition of urea to the phase F0 2 produced a third phase that has higher concentration of linoleic acid F0 3 (74.19%). In the final separation, the addition of urea, and cooling at 4  C caused the formation of a final phase C4 with 97.86% purity of oleic acid, with traces of other unsaturated acids. The same goes for linoleic acid, which would be obtained at the final phase F0 4 with a purity of 95.71%. Our method is considered to be a very effective extraction of two valuable acids from OMW, even though the study was conducted with oils extracted from OMW which have low concentrations of oleic acid and linoleic. After calculating the yield of recovering, based on the results shown in Figs. 2 and 3, 59.07% and 52.21% are, respectively, the quantities of oleic and linoleic acid recovered. It also must be mentioned that the four phases have yields greater than 50%. This confirms the effectiveness of the proposed technique for recovery of oleic and linoleic acids from OMW with high purities and yields. • Polyphenols from aqueous phase To determine the content of polyphenols in the sample of the aqueous phase extracted by ethyl acetate, the official spectrophotometric using the reagent FolinCiocalteu was carried out according to the procedures described by several works (Folin and Ciocalteu 1927; Atanassova et al. 2005; Singleton and Rossi 1965). The content of phenolic compounds in OMW used in this study is 2.1 g/L, lower than that reported by other authors (Adhoum et al. 2004; Tsioulpas et al. 2002). This is due to the fact that the samples of the aqueous phase were only centrifuged once, nor filtered before the determination of polyphenols concentration. Several techniques have showed that the amount of polyphenols under acidic conditions is higher than that reported in crude OMW.

Fatty acid Palmitic acid Palmitoleic acid Stearic acid Oleic acid Linoleic acid Linolenic acid

Structure C16:0 C16:1 C18:0 C18:1 C18:2 C18:3

Composition 7.96 0.69 1.29 68.22 20.31 1.53

F1 0.36 0.05 – 78.19 20.37 1.03

Table 3 Fatty acid composition during the crystallization of oleic acid C1 33.88 – 14.26 36.46 7.35 8.05

F2 1.52 1.67 1.02 41.22 50.51 4.06

C2 3.83 2.46 0.24 84.61 5.73 3.13

F3 2.15 3.11 – 86.09 6.79 1.86

C3 9.17 4.19 – 79.21 5.76 1.67

F4 1.86 4.89 – 24.12 68.02 1.11

C4 0.88 0.13 – 97.86 1.07 0.06

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Fatty acid Palmitic acid Palmitoleic acid Stearic acid Oleic acid Linoleic acid Linolenic acid

Structure C16:0 C16:1 C18:0 C18:1 C18:2 C18:3

Composition 7.96 0.69 1.29 68.22 20.31 1.53

F0 1 3.81 2.49 0.57 55.34 35.12 2.67

C0 1 32.62 – 19.17 32.16 12.59 3.46

Table 4 Fatty acid composition during the crystallization of linoleic acid F0 2 2.11 2.05 – 36.05 56.12 3.67

C0 2 4.18 2.93 – 86.64 3.94 2.31

F0 3 0.88 1.09 – 23.84 74.19 –

C0 3 9.83 4.84 – 78.72 4.32 2.29

F0 4 – 1.10 – 2.84 95.71 0.35

C0 4 3.61 2.37 – 88.92 3.29 1.81

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Fig. 2 Diagram of evolution esters mass and content of methyl oleate during crystallization

100 80

%

60 40 Mass of esters Methyl oleate

20 0 Initial mixture

F'1

F'2

F'3

F'4

Crystallization phases Fig. 3 Diagram of evolution esters mass and content of methyl linoleate during crystallization

Mass of esters

100

Methyl linoleate 80

%

60 40 20 0 Initial mixture

F'1

F'2

F'3

F'4

Crystallization phases

Since the OMW are released into the environment, without any serious pretreatment, it is preferable to extract polyphenols from raw sample of OMW, in order to have a high efficiency of technique. The final residue has a concentration of 1407 mg/L, giving a yield of 67%. • Characteristics of resulting soap and glycerin From 100 g of separated oil, 101.67 g of soap was produced; the yield of saponification can be calculated based on the reaction of hydrolysis, 98.55%. The saponification of residual oil from OMW provided soap with a very good quality and similar to the soap made from pomace olive (Elkacmi et al. 2016a, b).

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The alkaline aqueous solution produced during the saponification and transesterification processes contains more than 70% of glycerol and traces of soap, which will be eliminated by neutralization with HCl followed by filtration. In the final phase, the solution was concentrated by evaporation and distillation to produce an extra pure glycerin (99.5%).

Conclusion The olive mill wastewater (OMW) produced by the olive oil extraction process is the main harmful waste of this industry. The purpose of this study is to treat these effluents on one hand to reduce their toxicity by electrocoagulation technique using aluminum electrodes and on the other hand to the recovery of valuable products. A variable order kinetic (VOK) derived from the Langmuir-Freundlich equation was developed to simulate the kinetics of the detoxification of OMW with electrocoagulation using bipolar aluminum electrodes in STR. The results showed good agreement between the predictive equation and the experimental data. The originality of the second part lies in the separation by a centrifugation of oil from olive mill wastewater which cannot be intended for food because of its high acidity; the fractional crystallization with urea was developed as a quick, easy, and effective method to extract oleic and linoleic fatty acids as a product with a very important commercial value and preparation of biodegradable soap and glycerol as a coproduct. On the other hand, the aqueous part will be used for the recovery of polyphenols.

Cross-References ▶ New Techniques for Treatment and Recovery of Valuable Products from Olive Mill Wastewater

References Achak M, Mandi L, Ouazzani N (2009) Removal of organic pollutants and nutrients from olive mill wastewater by a sand filter. J Environ Manag 90(8):2771–2779 Adhoum N, Monser L (2004) Decolourization and removal of phenolic compounds from olive mill wastewater by electrocoagulation. Chem Eng Process Process Intensif 43(10):1281–1287 Antonacci R, Brunetti A, Rozzi A, Santori M (1981) Trattamento Anaerobica di Acque di Vegetazione di Frantoio Risultati Preliminari. Ingegneria Sanitaria 6:357–363 Atanassova D, Kefalas P, Psillakis E (2005) Measuring the antioxidant activity of olive oil mill wastewater using chemiluminescence. Environ Int 31(2):275–280 Aviani I, Laor Y, Medina S, Krassnovsky A, Raviv M (2010) Co-composting of solid and liquid olive mill wastes: management aspects and the horticultural value of the resulting composts. Bioresour Technol 101(17):6699–6706

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Matteson MJ, Dobson RL, Glenn RW, Kukunoor NS, Waits WH, Clayfield EJ (1995) Electrocoagulation and separation of aqueous suspensions of ultrafine particles. Colloids Surf A Physicochem Eng Asp 104(1):101–109 Meunier N, Drogui P, Montané C, Hausler R, Mercier G, Blais JF (2006) Comparison between electrocoagulation and chemical precipitation for metals removal from acidic soil leachate. J Hazard Mater 137(1):581–590 Morillo JA, Antizar-Ladislao B, Monteoliva-Sánchez M, Ramos-Cormenzana A, Russell NJ (2009) Bioremediation and biovalorisation of olive-mill wastes. Appl Microbiol Biotechnol 82(1):25 Mouncif M, Tamoh S, Faid M, Achkari-Begdouri A (1993) A study of chemical and microbiological characteristics of olive mill waste water in Morocco. Grasas Aceites 44(6):335–338 Murray KE (1955) Low pressure fractional distillation and its use in the investigation of lipids. Prog Chem Fats Other Lipids 3:243–273 Paraskeva P, Diamadopoulos E (2006) Technologies for olive mill wastewater (OMW) treatment: a review. J Chem Technol Biotechnol 81(9):1475–1485 Paredes C, Bernal MP, Cegarra J, Roig A (2002) Bio-degradation of olive mill wastewater sludge by its co-composting with agricultural wastes. Bioresour Technol 85(1):1–8 Razouk R, Oubella K, Chergaoui A (2012) Changes induced by olive mill wastewaters used as organic fertilizer on polyphenol content, pH and salinity of an alkaline soil. J Sci Res Rep 2(2):559–570 Roig A, Cayuela ML, Sánchez-Monedero MA (2006) An overview on olive mill wastes and their valorisation methods. Waste Manag 26(9):960–969 Sassi AB, Boularbah A, Jaouad A, Walker G, Boussaid A (2006) A comparison of olive oil mill wastewaters (OMW) from three different processes in Morocco. Process Biochem 41(1):74–78 Sayadi S, Allouche N, Jaoua M, Aloui F (2000) Detrimental effects of high molecular-mass polyphenols on olive mill wastewater biotreatment. Process Biochem 35(7):725–735 Scoma A, Varela-Corredor F, Bertin L, Gostoli C, Bandini S (2016) Recovery of VFAs from anaerobic digestion of dephenolized olive mill wastewaters by electrodialysis. Sep Purif Technol 159:81–91 Singleton VL, Rossi JA (1965) Colorimetry of total phenolics with phosphomolybdicphosphotungstic acid reagents. Am J Enol Vitic 16(3):144–158 Soto ML, Moure A, Domínguez H, Parajó JC (2011) Recovery, concentration and purification of phenolic compounds by adsorption: a review. J Food Eng 105(1):1–27 Swern D, Parker WE (1952) Application of urea complexes in the purification of fatty acids, esters, and alcohols. II. Oleic acid and methyl oleate from olive oil. J Am Oil Chem Soc 29(12):614–615 Takaç S, Karakaya A (2009) Recovery of phenolic antioxidants from olive mill wastewater. Recent Patent Chem Eng 2(3):230–237 Tsioulpas A, Dimou D, Iconomou D, Aggelis G (2002) Phenolic removal in olive oil mill wastewater by strains of Pleurotus spp. in respect to their phenol oxidase (laccase) activity. Bioresour Technol 84(3):251–257 Un UT, Koparal AS, Ogutveren UB (2009) Electrocoagulation of vegetable oil refinery wastewater using aluminum electrodes. J Environ Manag 90(1):428–433 Vlyssides AG, Loizides M, Karlis PK (2004) Integrated strategic approach for reusing olive oil extraction by-products. J Clean Prod 12(6):603–611

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Payal Garg, Geetanjali Kaushik, Jitendra Kumar Nagar, and Poonam Singhal

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Types of Earthworms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Ecological Categories of Earthworms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Earthworms and Pollution Control . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Role of Earthworms for Healthy Soil . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Role of Earthworms in Soil Erosion Control . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Vermicompost . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Nutrient Availability from Vermicompost . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Enzymatic Activities in Earthworm’s Cast and Vermicompost . . . . . . . . . . . . . . . . . . . . . . . . . . . . Plant Growth Hormones and Regulators in Vermicompost . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Use of Vermicompost for Crop Production . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Vermiwash . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Use of Earthworms in Vermiwash Production: A Liquid Manure . . . . . . . . . . . . . . . . . . . . . . . . . Effect of Vermiwash on Yield and Quality of Crops . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Use of Earthworms in Land Improvement and Reclamation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cross-References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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P. Garg · P. Singhal Centre for Rural Development and Technology, Indian Institute of Technology Delhi, New Delhi, India e-mail: [email protected]; [email protected] G. Kaushik (*) MGM’s Jawaharlal Nehru Engineering College, Mahatma Gandhi Mission, Aurangabad, Maharashtra, India e-mail: [email protected] J. K. Nagar Vallabhbhai Patel Chest Institute, University of Delhi, New Delhi, India © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_162

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Abstract

Rapidly growing Indian cities generate huge quantities of solid wastes on a daily basis. On account of enormous environmental implications the management of this organic solid waste has become a major problem. Heavy metal contamination due to dumping of contaminated solid wastes has been reported. Application of vermicomposting on a large scale for urban as well as rural waste disposal has been highlighted by research. This paper discusses the types of earthworms and their role in maintaining soil fertility, vermicomposting, soil erosion control and land reclamation. Keywords

Wastes · Management · Pollution control · Vermicomposting · Land reclamation

Introduction The role of earthworms in breaking down of dead plants and animal residues in soil was first studied by Charles Darwin in 1881. Later on many of the research workers studied the mechanisms of conversion of organic matter into humus by introducing vermiculture in the field. Earthworms can be defined as invertebrates belonging to phylem annelida, order oligochaeta, class clitellata, which live in soil. From the days of ancient greek philosophers and Darwin to the present day, these lowly creatures have been recognized as master builders of top soil and man’s fellow tillers. Aristotle has referred to earthworms as “the intestine of the earth.” Darwin (1881) remarked, “The plough is one of the most ancient and most valuable man’s inventions, but long before it existed, the land was infact regularly ploughed and still continues to be thus ploughed by earthworms.”

Types of Earthworms More than 4200 species of Oligochaetas are known in the word. Of these, 280 microderill and remaining about 3920 belongs to megadrilli (earthworms). In the Indian subcontinent earthworms also form bulk of the Oligochaete fauna. They are representing by 509 species and 67 genera, indicating a high degree of diversity in this region as compared to other areas. Earthworms form a major component of the soil biota, and they together with a large number of other organisms constitute the soil community. The chief source of food to the soil biota is the litter contributed by plants. Although the dead plant tissues constitute the bulk of the food ingested by the earthworms, living microorganisms, fungi, microfauna and mesofauna, and their dead tissue are also ingested as an important part of the diet (Parle 1963; Piearce 1978). Though earthworms are generally called as saprophages, they can be classified based on the feeding habits (Lee 1985) into detrivores and geophages.

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Detrivores feed at or near the soil surface, mainly on plant litter or dead roots and other plant debris in the organic matter–rich surface soil horizon or on mammalian dung. These worms are classified as humus formers and comprise of epigeic and anecic forms. Perinyx excavatus, Eisenia fetida, Eudrillus eugineae, Lampito mauritti, Polypheretima elongate, Octochaetona serrata, and Octochaetona surensis are a few examples of detrivores earthworms. Geophages worms feed deeper beneath the surface, ingesting large quantities of organically rich soil. The worms are generally called as humus feeders and comprise of the endogeic worms. Octochaetona thrustoni is one such earthworm commonly available in Madras. A different classification has been proposed by Bouche (1977) laying stress on ecological strategies. He classified earthworms into epigeics, anecics, and endogeics. The epigeics have no effect on the soil structure as they generally cannot dig. They are efficient agents of combination and fragmentation of leaf litter. These are broadly classified as phytophagous earthworms. The anecic feed on the leaf litter mixed with the soil of the upper horizons. They may also produce surface casts. These are called as geophytophagous earthworms.

Ecological Categories of Earthworms Based on the nature and their vertical distribution in soil, earthworms can be distinguished into three different ecological categories (Bouche 1977), (i) Epigeic – In nature epigeic worms live in the top soil and duff layer on the soil surface. These small, deeply pigmented worms have a poor burrowing ability, preferring instead an environment of loose organic litter or loose topsoil rich in organic matter to deeper soils. Epigeic species feed in organic surface debris and have adapted beautifully to the rapidly shifting, dynamic environment of the soil surface. They are tolerant to some disturbance, have moderate to high rate of cocoons production, and have short life cycle. (ii) Endogeic – Endogeic worms build complex lateral burrow systems through all layers of the upper mineral soil. These worms rarely come to the surface, instead of spending their lives in these burrow systems where they feed on decayed organic matter and bits of mineral soil. They are the only category of worm which actually eats significant volumes of soil and not strictly the organic component. Endogeic worms tend to be medium sized, tolerant to some disturbance, have moderate to high rate of cocoon production, lifecycle not very long, and are pale colored. (iii) Anecic – Anecic worms (like the common night crawler Lumbricus terrestris) build permanent, vertical burrows that extend from the soil surface down through the upper mineral soil layer. These worm species coat their burrows with mucous which stabilizes the burrow so it does not collapse, and build little mounds (called middens) of stone and castings outside the burrow opening.

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They also tend to be very large worms and have bellies with less pigmentation than their backs. These worms are intolerant to disturbance, and have low rate of cocoons production, and long life cycle (Julka 1986). It is reported that mostly epigeic and some endogeic species are suitable for vermicomposting.

Earthworms and Pollution Control Management of organic solid waste has become a major problem due to environmental implications, thus attracting the attention of researchers (Patil et al. 2016; Patil and Kaushik 2016). Many publications related to solid waste generation, problems and disposal practices have emerged throughout the world to generate environmental awareness. The urgency for recycling and composting of solid wastes is felt as countless dumping areas in and around urban and rural settlements causes environmental hazards (Chaturvedi 1994; Chaturvedi et al. 2008; Bandela et al. 2016a, b). The long-term consequences of organic wastes dumping near the agricultural fields in Hyderabad have been reported (Rao and Shantaram 1994). It is found that soil up to 30-cm depth was contaminated with heavy metals due to dumping of contaminated solid wastes. Unprecedented increase in the cost of chemical fertilizers hit the small and marginal farmers badly. An obvious way to mitigate their difficulty is generation of fertilizers at the village level through recycling of wastes. Bhiday (1994) asserted that the feasibility of using vermiculture on large scale for urban as well as rural waste disposal is very high and it is a right kind of technology for recycling such wastes. Earthworms encourage growth of microorganisms in their gut providing ideal conditions there in (Bhawalkar 1989). Werner and Cuevas (1996) reported that Cuba has more than 170 vermicomposting centers that are engaged in producing earthworm casting for use as fertilizers in tobacco farming. Mani (1997) suggested that vermicomposting of animal and agricultural waste results in higher nutrient content within 6 weeks. Singh and Rai (1997) asserted that earthworm farming and vermicomposting is a boon for sustainable agriculture in India. Status report on solid waste management by vermicomposting in India has been prepared by Vasudevan (2001) (Bhawalkar and Bhawalkar 1993). Santra and Bowmik (2001) studied that vermicomposting is attaining a special significance for the abatement of pollution hazards created by large amount of organic wastes in our country and also reduce the demand for chemical fertilizers.

Role of Earthworms for Healthy Soil Earthworms, soil’s intimate friends and benefactor, has since long been helping soil in respiration, nutrition, excretion, and various other vital activities. Through its characteristic functions of breaking, grinding, churning, assimilation, and tunneling, earthworm has proved to be soils mouth, stomach, and intestine. The earthworm is

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nature’s marvelous control machine as it eats practically almost any matter except rubber, plastic, glass, and metal. Alvares (1984) reported that 20 million worms could handle 80 metric tons of pulp sludge daily or about 5 gm of sludge per worm per day and whatever worm eats get transformed into organic fertilizer. The worm’s castings are rich in nitrates, phosphates, and potash in other words a rich source of organic fertilizer. Earthworms, the major secondary decomposer macrofauna have a very major role in hasting up of the rate of decomposition and also in improving the structural properties of soil. They serve as agent for pollution control and for amelioration of the soil. The lumbricid earthworms are dominantly distributed in the temperate soil as megascolecid earthworms are in the tropical and subtropical soils. Roles of earthworms in soil fertility are very high basically due to life activities of earthworms in soil and their importance in utilization of earthworms for human welfare, viz., vermiculturing and vermicomposting. Earthworms establish well in lands receiving organic manure. Role of earthworms in the breakdown of organic debris on soil surface and in soil turn over process was first highlighted by Darwin (1881). The involvement of earthworms in the composting process decreases the time of stabilization of the waste and produces an efficient organic pool with energy reserves as vermicompost. The sludge from both agro-based industries and domestic sewage plants can be a food source for composting earthworms with suitable organic amendments such as plant litter or animal waste. Earthworms in nature promote infiltration of water in ground as they make soil porous and promote drainage. Earthworms increase natural fertility of soil. They would help in decreasing the chemical fertilizer use and increase the fertilizer efficiency in biological and natural way. Chemical fertilizers are well known to affect soil chemistry, deplenish soil micronutrients, and cause water pollution.

Suitable Earthworm Species for Vermicomposting Species which are identified as potentially useful species to break down organic wastes include E.fetida, Dendrobaena venta, and Lumbricus rubellus, from temperate areas and Eudrilus eugeiae and perionyx excavatus from the tropics. In different parts of India, degradation of organic waste was done successfully using a number of species, viz., E.fetida, Aporrectodea caliginosa, Eudrilus eugeniae, Perionyx arboricola, etc. (Goswami and Kalita 2000). E.fetida (Savigny) is abundant worldwide for vermicomposting and a large literature is available on its performance in converting organic waste into quality vermicompost and its high growth rate on various organic substrates (Kaviraj and Sharma 2003; Kaushik and Garg 2003; Thimmaiah 2001). It is an epigeic earthworm species which lives in organic wastes and require high moisture content, adequate amount of suitable organic material, and dark conditions for proper growth and development (Gundai and Edwards 2003; Chaudhari and Bhattacharjee 2002). E.fetida proved to be a suitable species for vermicomposting application, because of its rapid growth rate, reproductive potentials, and occurrence in rich organic substrates in nature (Neuhauser et al. 1980).

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Influence of Earthworm Activities on Soil Properties Earthworm activities are direct action of feeding and burrowing along with related biological activities. Earthworm eats its way through soil and organic humus, etc. Earthworms improve the soil fertility in following ways: Influence on Soil pH The pH of the intestinal contents of earthworms is remarkably stable around neutral to slightly alkaline. This can have a profound effect on the overall level of soil pH and on the course of organic decomposition. In neutral or slightly alkaline conditions, bacterial activity is favored, leading to more complete breakdown of organic compounds and a multi-type humus. Physical Decomposition The passage of organic material through the earthworm gut results in the physical decomposition due to the muscular grinding action of gizzard, aided by ingestion of silica granules. This provides considerably enhanced surface area for microbial decomposition. Humus Formation The process of humus formation is often characterized by the selective breakdown of cellulose. The end product is a complex mixture of various organic acids, amino acids, polyphenols, and sugars such as glucose, galactose, mannose, arabinose, and xylose. Lignin fibers are present in raw humus and peat but are degraded to polyphenols in well-decomposed humus. Improvement of Soil Structure The physical communication of organic particles, the amelioration of soil pH, the enhancement of microbial decomposition activity, all these results of earthworm activity contribute to soil fertility. Burrowing of earthworms brings about tillage of soil up to 3 m without adversely affecting plants in any manner excepting some species in special situations. This also accompanies breakdown of soil particles and mixing of soil nutrients and bacteria in digestive process as well as with deposit of casts of various levels. These affect automatic conversion of organic wastes. These micronized soil particles lead increase of particle surface area which leads increased moisture absorption and holding, air circulation, etc. This also increases microbial action. Casts is fine biofertilizer having up to 1000 times more microbes than in surrounding soil. Increased porosity eventually increase percolation of water generally referred as charging of sub-soil water. This in turn led maintenance of soil temperature which adds to toleration in soil faunal and floral components. These reduce severity of soil fluctuations essential for plant growth. Soil Enrichment Application of vermicompost significantly improved the physical properties of all the soil types under study. Earthworms physically mix the contents of the deeper layers and make the soils loose and porous. Their body exudates improve the water holding capacity of soil and promote establishment of microorganisms (Kale 1994). Rani and Srivastava (1997) conducted an experiment on rice with full dose of nitrogen replaced by one-third and quarter of N as vermicompost.

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Compared with N fertilizer alone, vermicompost application showed increase in grain yield and yield components of rice. It also improves the availability of phosphorous and potassium as well as micronutrients.

Role of Earthworms in Soil Erosion Control Soil aggregates are formed by the adhesion of mineral and organic particles. The shape and physical packing of which influence the aeration, infiltration of water, water holding capacity and surface area, etc. Formation of aggregates makes the soil well aerated and drained. These aggregates are mineral granules joined together in such a way that they can resist wetting, erosion, or compaction and remain loose when the soil is dry or wet. Most of the workers agree that earthworm casts contain more stable aggregates than the surrounding soil. In an experiment, the percentage of aggregates in soil to which earthworms are added was compared with that in soil without earthworms. In general, earthworm burrows and structural aggregation due to their casting activities promote water entry into the soil and therefore reduce surface runoff. In addition, polysaccharide gums are produced by the bacteria by the soil, as the soil passes through the gut of the earthworm (Dash 1978). The production of the polysaccharides gums is enhanced by the presence of the organic components in the ingested material. Another possibility of the cause of stability is that the soil particles are cemented by calcium humate which is derived from the interaction of the ingested organic matter and calcite excreted by the earthworm’s calciferous glands.

Vermicompost Roles of earthworms in soil fertility are very high basically due to life activities of earthworms in soil and their importance in utilization of earthworms for human welfare, viz., vermiculturing and vermicomposting. Earthworms establish well in lands receiving organic manure. Role of earthworms in the breakdown of organic debris on soil surface and in soil turn over process was first highlighted by Darwin (1881). The involvement of earthworms in the composting process decreases the time of stabilization of the waste and produces an efficient organic pool with energy reserves as vermicompost. The sludge from both agro -based industries and domestic sewage plants can be a food source for composting earthworms with suitable organic amendments such as plant litter or animal waste.

Nutrient Availability from Vermicompost About 5–10% of the ingested material is absorbed into the tissue of earthworms for their growth and metabolic activity and rest is excreted as cast or vermicompost. Vermicomposting are a highly enriched kind of biofertilizer. It is more chemically

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Table 1 Nutrient content Nutrient of vermicompost Organic carbon Total nitrogen Available phosphorous Available potassium Available sodium (ppm) Copper Iron Zinc Available sulfur

Percentage 9.15–17.98 0.5–1.5 0.1–0.3 0.15–0.56 0.06–0.03 2.0–9.5a 2.0–9.3a 5.7–11.5a 128–548a

Source: Kale 1995 a Values in ppm

neutral than the surrounding soil. The nutrient levels in the vermicompost depend on the nature of the organic waste used as food source for earthworms. It is found that a heterogeneous waste mix will have balanced level of nutrients than from any one particular waste. Vermicompost contains most nutrients in plant available forms such as nitrates, phosphates, exchangeable calcium, soluble potassium, etc. (Edward 1998) and large surface area that provide many micro sites for microbial activity and for the strong retention of nutrients. The nutrient status of vermicompost is given in Table 1 (Kale 1995). The vermicompost is considered an excellent product since it is homogenous, has reduced level of contaminants, and tends to hold more nutrients over a longer period without impacting the environment (Ndegwa et al. 2000). Earthworm cast typically have high N content which suggests that they would be good sources of plant N (Parmelee and Crossley 1988; Ruz-Jerez et al. 1992). In addition to increased N availability, C, P, K, Ca, and Mg availability in the casts is also greater than initial feed material (Orozco et al. 1996; Daniel and Anderson 1992; Lavelle and Martin 1992; Basker et al. 1993). The beneficial effect of earthworm cast has been observed in both horticultural plants (Hidalgo 1999; Saciragic and Dzelilovic 1986) and in agronomic crops (Pashanasi et al. 1996). More than the regular macronutrients, vermicompost contributes to the supply of micronutrients essential for crops. The stimulatory effect of vermicompost for nutrient uptake, growth, and yield of crops is linked to the secretion of earthworms and the associated microbes mixed with the cast.

Enzymatic Activities in Earthworm’s Cast and Vermicompost Enzyme activities have been postulated as indicators of the decomposition process (Diaz-Burgos et al. 1992; Garcia et al. 1993). Vermicompost contains enzymes such as proteases, amylases, lipases, cellulases, and chitinases, which continue to disintegrate organic matter even after earthworms have been excreted and hence, vermicompost is believed to have additional attributes of providing enzymes and

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hormones which stimulate plant growth (Abbasi and Ramasamy 1999; Atiyeh et al. 2001;Chaoui et al. 2003). In several studies, higher enzyme activities were measured in worm casts than in surface soil. Earthworm casts have been shown to have enhanced carbohydrase, protease (Ross and Cairns 1982), phosphatase, and dehydrogenase activities than the surrounding un-ingested soil (Tiwari et al. 1989; Mulongoy and Bedoret 1989). Ranganathan and Vinotha (1998) showed enhanced enzyme activity in the gut of E.eugeniae when it was fed with press mud.

Plant Growth Hormones and Regulators in Vermicompost Vermicompost contain plant growth regulators and other plant growth influencing materials (Tomati et al. 1988; Grappelli et al. 1987; Atiyeh et al. 2002). The beneficial influence of worm cast has been related to the biological factors like gibberellin, cytokinins, and auxins released due to microbial activity of the microbes harbored in the cast (Brown 1995). It is reported that certain metabolites produced by earthworms may be responsible for stimulating plant growth. It is considered that earthworms release into the soil certain vitamins and similar substances which may be B group vitamins or some pro vitamin D or free amino acids. Vermicompost also contain large amounts of humic substances (Senesi et al. 1992; Masciandaro et al. 1997) and some of the effects of these substances on plant growth have been shown to be very similar to the effects of soil-applied plant growth regulators or hormones (Muscolo et al. 1999). The antibacterial activity of coelomic fluid of earthworms is accounted and this activity is only directed against the highly pathogenic soil bacteria. Thus it could be deducted that earthworms apart from encouraging the establishment of beneficial microorganisms to some extent can also inhibit the soil borne pathogens. Some of the organic acids that are isolated from the body fluid of earthworms and their cast have similar response as that of plant growth promoter substances. The nutrient value of worm castings is not high compared to chemical fertilizers. The key factor is microbial activity. Microbial activity is 10–20 times higher than in the soil and organic matter that the worm ingests. The most important effect of earthworms may be the stimulation of microbial activity that occurs in casts. This enhances the transformation of soluble nitrogen into microbial protein, preventing their loss by leaching to the lower horizons of the soil. Vermicompost is better than chemical fertilizers in economical and ecological aspects. Replacing costly yet deadly chemicals with cheap yet friendly vermicompost will ensure sustainable food production. Use of vermicompost as manure has multifold benefits. They are: (i) (ii) (iii) (iv)

Healthy soil with soil organisms Limited external inputs Cost effective farming practices and healthy food Problems of leaching and mineralization of nutrients are reduced

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Use of Vermicompost for Crop Production Vermicompost is one such material which is being experimented as a growth media and carried for biofertilizer. The use of vermicompost as soil amendments can have many positive effects on soil physical characteristics following high rates of application. It is rich in nitrogen, phosphorous, potassium, carbon, and organic matter all of which are essential for the growth of microbes. High levels of organic humic matter soil amendment in the form of compost improve soil structure by increasing porosity and reducing the bulk density of an amended soil. Polysaccharides and other polymeric substances present in organic matter act as aggregating compounds (Masciandaro et al. 2000) and increase micropores in the soil. The improvements to soil physical structure, soil fertility, and soil microbiological properties associated with compost application all promote plant growth, as a growth medium for transplants and a soil amendment for field crops. Several studies have evaluated the effect of vermicompost-amended potting media on plant growth greenhouse production. Generally, potting medium with 10–20% vermicompost by volume provides adequate fertilization for transplant growth (Subler et al. 1998; Atiyeh et al. 2000; Ozores-Hampton and Vavrina 2002. In one study, germination rates of greenhouse tomatoes increased up to 15% when vermicomposted pig manure was mixed with potting medium at 20%, 30%, and 40% by volume. The highest marketable yield of fruit was reported in the 20% mixture. Treatments consisting of 100% vermicompost led to smaller growth and fewer leaves than other treatments, due to high-moisture and possible phytotoxicity (Atiyeh et al. 2000). Sagar et al. (2002) compared growth of Ocimum sanctum, a crop used for essential oil, in vermicomposted cow manure, farmyard manure, urea, and a control growth medium. While oil yield was 15 g kg1 in all treatments, leaf weight was 25% higher in vermicompost treatments than control and plant weight 52% higher. Ranganathan and Cristopher (1994) reported that the vermicompost not only helped to protect fertility of the soil but also boosted productivity depending on crops, season, and other factors and also enhanced the quality of end products.

Vermiwash Vermiwash is a collection of excretory products and mucus secretions of earthworms along with nutrients from the soil organic molecules. If it collected properly is a clear and transparent, pale yellow colored fluid. The utility of vermiwash as a biocide, either singly or when mixed with botanical pesticides is still under investigation.

Use of Earthworms in Vermiwash Production: A Liquid Manure Foliar Sprays are used as a part of agronomic practices for crop production; vermiwash is one such new concept in this context. Worm worked soils have burrows formed by the earthworms. Bacteria richly inhabit these burrows, also

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called as the drilospheres. Water passing through these passages washes the nutrients from these burrows to the roots to be absorbed by the plants. This principle is applied in the preparation of vermiwash. It is a collection of excretory products and mucus secretions of earthworms along with nutrients from the soil organic molecules. If it collected properly is a clear and transparent, pale yellow colored fluid. The utility of vermiwash as a biocide, either singly or when mixed with botanical pesticides is still under investigation. Following methods of vermiwash preparations are described by few scientists:

First Method Ismail (1997) has suggested a method of preparation of vermiwash using a barrel in which the layers of cattle dung and hay were placed on the top of the layer of soil (Fig. 1). The epigeic earthworms were introduced to produce compost at faster rate and anecics were put to produce a large number of drilospheres. The vermiwash was collected from the bottom of the barrel after sprinkling water through perforated mud from the above. It was suggested to use the vermiwash for foliar spray either as such or dilution with water or 10% cow’s urine. The physicochemical characteristics of vermiwash are as follows (Table 2). Second Method A method described by Kale (1998) consisted of an outer and an inner vessel. The inner vessel would have an outlet at the lower side of the vessel. The inner vessel was filled with decomposing organic matter and about 1–2 kg earthworms were accommodated in 12–16 L capacity vessel. (Fig. 2). As the earthworms were accommodated in waste, water was slowly added into vessel in excess. The excess water flowing out through the outlet as thick syrupy fluid was collected in the outer vessel. The fluid so collected was siphoned out and after diluting was used as foliar spray to

Fig. 1 Method of vermiwash preparation

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Parameter pH Chlorides (ppm) Sulfates (ppm) Inorganic phosphate (μg/l) Amonical nitrogen (ppm) Potassium (ppm) Sodium (ppm) Total hardness (ppm) Calcium hardness (ppm) Magnesium hardness (ppm) BOD (biological oxygen demand) (ppm) COD (chemical oxygen demand) (ppm)

Vermiwash 6.9 110.00 177.00 50.9 Below detectable level 69.00 122.00 375.00 175.00 200.00 4.60 97.00

Source: Ismail 1997

Fig. 2 Method of vermiwash preparation

Rubber Tube Funnel Plastic Tub Vermicompost Vermiwash Inner funnel Pebbles & Sand

Vermiwash

different crops. In another method, the cement tank was built at an elevated place from which they want to collect the wash. The slope provided in the tank provided scope for excess water to flow out in drops as thick syrupy emulsion through a small outlet. This was collected in a container and stored in bottles. Before using, it was diluted and sprayed to crops.

Third Method Karuna et al. (1999) also prepared earthworm exudates and used successfully for foliar spray to Anthuriums. In this method, clitellated earthworms of Eudrillus eugeniae were collected from the culture bins. These worms were cultured on organic waste comprising mostly of leaf litter, weeds, stubble mixed with one part of cow dung by volume in the form of slurry. Earthworms weighing 1 kg were placed into a dry enamel tray for 15–20 min to clear out the cast that will be excreted due to handling. Earthworms were then carefully separated from their excreta and then added into glass plastic bowl having 500 ml of distilled water having temperature of 37–40  C (luke warm) (Fig. 3). The worms were agitated for 3–5 min and removed and added into another bowl containing 500 ml of water at room temperature

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Fig. 3 Method of vermiwash preparation

(25–27  C) to rinse them thoroughly to collect the exudates adhering to its body wall before releasing back to the culture bins. The contents in the two bowls were mixed to use as a spray. The exudates thus collected were a syrupy, light yellow fluid.

Fourth Method Giraddi (2001) evaluated a method of extraction of earthworm wash, a plant promoter substance. In order to have simple devices of collection of earthworm secretions, an equipment made of GI sheets or plastic crates is used. The equipment consists of culturing compartment, joined to collection chamber at the bottom, while water sieve is kept enclosed on the culturing chamber. Both water sieve and culturing compartment have got small holes to allow water from sieve to culturing compartment to collection chamber, and this is regulated by inserting a divider in between two chambers. Number of earthworms, quantity and quality of raw material used, duration of earthworm culturing, and time allowed for water to stagnate in the culturing compartment would decide the quality of Vermiwash to be used. Vermiwash have enzymes, secretions of earthworms which would stimulate the growth and yield of crops and even develop resistance in crops receiving this spray. Such a preparation would certainly have the soluble plant nutrients apart from some organic acids and mucus of earthworms and microbes (Shivsubramanian and Ganeshkumar 2004). But so far there are no experimental evidences to quantify the effect of such spray. Chemical composition of vermiwash varies with the types of substrates used for vermiwash (Lourduraj and Yaday 2007).

Effect of Vermiwash on Yield and Quality of Crops As in crops, to tackle the pest problem, indiscriminate use of synthetic chemical insecticides was undertaken. This had led to many serious problems like environmental contamination by way of pesticide residues, development of resistance in

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pests to pesticides, pest resurgence, destruction of natural enemies, etc. There is a need for developing pest management strategies, which are eco-friendly, and environmentally safe. Vermiwash seems to possess an inherent property of in soil fertility, is very important in maintaining balance in acting not only as a fertilizer but also as a mild biocide on ecosystem (Shuster et al. 2000). The fresh vermiwash harbors a large number of beneficial microorganisms that help in plant growth and protects it from a number of infestations. Ismail (1997) reported that vermiwash can be sprayed on plants as a foliar spray for improving quality and yields of Okra crop. Foliar application of vermiwash prepared by both the methods resulted into considerable increase in total nitrogen, phosphorous, and potassium uptake by seasonal chrysanthemum, marigold, and china aster in comparison to NPK alone gave indication of quick absorption of the above nutrients through foliage for better nourishment of these flowering plants (Todkari and Talashilkar 2001).

Use of Earthworms in Land Improvement and Reclamation It is now well established that the introduction of earthworms into soils that have no earthworms or have only low natural earthworm population is usually beneficial in terms of plant growth and crop yield. Stockdill and Cossens (1966) have successfully introduced earthworms into pastures in NewZealand that lacked native earthworms. Following establishment of the earthworm populations, the organic matter at the base of the grass disappeared and compacted soils with poor structure were transformed into deep friable top soils. Some of the methods of earthworm inoculation for the improvement and reclamation of land tried by various workers are described below:

Direct Release of Worms A large number of deep working species of earthworms are introduced by sufficiently moistening and liming of acidic soils. If food is a limiting factor, addition of organic material such as municipal sludge or animal wastes allow their earlier establishment to facilitate stabilization of temperature and moisture conditions. The mature worms should be placed at the rate of 150/m2. Addition of earthworms to soil seems particularly promising in reclaiming flooded areas that are subsequently drained and put into cultivation with tree crops. In such cases, earthworms can be introduced at the rate of 100–180 worms per tree in the basis around the base of tree and mulched with cattle dung and leaf litter. As Soil Blocks Earthworms can be transferred to new localities by moving blocks of soil cut from the surface at localities where target species are common. Such Blocks of soil are commonly introduced into small areas and supplied with organic matter as food.

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As Turf Pieces In this method, turf pieces with earthworms are cut and placed on the ground surface at 10 m spacing between them. To cut turf pieces, sod machines are developed in New Zealand. This method is useful for introduction of earthworms into pasture devoid of earthworm before. As Vermicomposting Surface application of vermicompost to degrade soil can improve its structure and fertility. Such lands may be planted with grass cover for the establishment of earthworms. Barley and Keleing (1964) successfully introduced Aporrectodea caliginosa and the megascolecid species, Microscolex dubius into newly sown, irrigated pasture on sandy loam soil in Australia with corresponding significant improvement is soil structure, loss of organic materials, and increased productivity. Ghilarov and Mamajev (1966) inoculated earthworms into reclaimed irrigated land in Uzbekistan and reported considerable improvement in soil structure and fertility. Van Rhee (1969, 1971) introduced earthworms into polders that had been drained and reclaimed from the sea in the Netherlands and reported that this accelerated the development of normal soil structure and facilitated the return of the polders to productive use. He found that grass yields of these polders increased up to four times and clover yields up to ten times, after inoculation with earthworms. The dry matter yield of spring wheat was doubled by inoculation with earthworms and fruit trees were established much more readily than in uninoculated soil.

Conclusion With rapidly growing Indian cities becoming powerhouses of commercial and industrial activity, huge quantities of wastes are expected to be generated every day. Indiscriminate disposal of organic wastes is associated with various forms of environmental hazards. In this background, earthworm through application of vemicomposting offers a ray of hope for waste management and pollution control.

Cross-References ▶ Application of Novel Microbial Consortia for Environmental Site Remediation and Hazardous Waste Management Toward Low- and High-Density Polyethylene and Prioritizing the Cost-Effective, Eco-friendly, and Sustainable Biotechnological Intervention ▶ Management of Hazardous Paper Mill Wastes for Sustainable Agriculture ▶ Soil Pollution and Remediation

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Todkari AA, Talashilkar SC (2001) Effect of vermiwash prepared by two methods on growth characteristics, yield and nutrition of three following plants. Souvenir and abstracts. A paper presented in silver jubilee celebrations of the Indian Society of Soil Biology and Ecology and VII national symposium on soil biology and ecology, U.A.S, Bangalore, 7–9 Nov 2001, pp 97 Tomati U, Grappelli A, Galli E (1988) The hormone-like effect of earthworm casts on plant growth. Biol Fertil Soils 5:288–294 Van Rhee JA (1969) Inoculation of earthworms in a newly drained polder. Pedobiologia 9:128–132 Van Rhee JA (1971) Some aspects of the productivity of orchards in relation to earthworms activity. Ann Zool Ecol 4:99–108 Vasudevan P (2001) Status reports on solid waste management by vermicomposting, UNICEF sponsored project Werner M, Cuevas R (1996) Vermiculture in Cuba. Biocycle, vol 37, JG Press, Emmaus, pp 61–62

The Application of Membrane Bioreactors (MBR) for the Removal of Organic Matter, Nutrients, and Heavy Metals from Landfill Leachate

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Susan Hayeri Yazdi, Ali Vosoogh, and Alireza Bazargan

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . MBR and Its Application in BOD5, COD Removal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . MBR and Its Application in Nutrient (Phosphorous and Nitrogen) Removal . . . . . . . . . . . . . . . . MBR and Its Application in Heavy Metal Removal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AnMBR or Anaerobic Membrane Bioreactor . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . The Effect of pH and Recycling on Efficiency of MBR Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

Generally, the disposal of waste to landfills is a serious environmental concern. The leachate produced in landfill sites may infiltrate into the soil and contaminate surface and/or groundwater. Because of the high concentration of pollutants, landfill leachate (LFL) is very difficult to treat using conventional biological processes. Since the turn of the century, the application of membrane bioreactors (MBRs) has proven to be a promising alternative to conventional treatment methods.

S. Hayeri Yazdi Gas Turbine Power Plant Division, Monenco Iran Company, Tehran, Iran e-mail: [email protected] A. Vosoogh Department of Civil Engineering, Iran University of Science and Technology, Tehran, Iran e-mail: [email protected] A. Bazargan (*) Department of Civil Engineering, K. N. Toosi University of Technology, Tehran, Iran e-mail: [email protected] © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_168

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MBRs are essentially composed of two main sections, the biological unit or bioreactor responsible for the biodegradation of the waste compounds and the membrane module for the separation of the treated water from biosolids/microorganisms. MBR systems commonly use ultrafiltration (UF) or microfiltration (MF) membranes in hollow fiber, plate and frame, flat sheet, or tubular configuration. The microorganisms are of importance because the capacity of MBR systems to eliminate organic matter depends on the biological activity in the active sludge. Also, pretreatment and posttreatment can improve MBR efficiency. In the following chapter, a review of the studies pertaining to landfill leachate treatment with membrane bioreactors has been provided. The results show that excellent biological oxygen demand (BOD5) and ammonia removals of 90% or higher are achievable with a much shorter hydraulic residence time (HRT) and much larger organic loading rate (OLR) in comparison to conventional biological systems. MBR systems also allow for excellent chemical oxygen demand (COD) removal (higher than 75% and in some cases even exceeding 90%), even with old LFL under optimized conditions. Heavy metal concentrations could be reduced by more than 99%, and for NH4-N removal, percentages of more than 97% have been reported. MBRs have also been effectively used to remove micropollutants. Furthermore, recent developments such as anaerobic MBR and PAC-amended (powdered activated carbonamended) MBR have shown great potential in LFL treatment. However, one of the most important problems with the application of MBR for landfill leachate treatment is the occurrence of biofouling on the membrane surface, leading to diminished flux and the requirement of cleaning processes. Keywords

Membrane bioreactor · Nutrients removal · Anaerobic bioreactor · Metal removal · MBR efficiency

Introduction Commercial and industrial developments, alongside improving living standards, have resulted in rapid end-of-life for products and the present tendency to produce wastes in modern society. This has resulted in ever-increasing production of solid wastes. One of the most routine methods for waste management is land filling. Today, in most countries, disposal by landfilling is the basic fate of most wastes which are collected and transported to a landfill site directly. Landfilling differs from mere dumping of waste and can be defined as the disposal of solid wastes on or within the earth in a sanitary way without creating public health hazards or inconvenience (Kreith 1999). Landfill leachate (LFL) is generated during the percolation of rainwater and waste moisture and the decomposition of waste from the landfill area. So leachate contains

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a large amount of inorganic and organic compounds, and their concentrations depend on the landfill environment, the type of waste, the age of the landfill site, and filling technique (Campagna et al. 2013). Based on the age of the LFL, it is categorized as young, intermediate, or old/stabilized (Sahinkaya et al. 2013). For the purposes of this chapter, young LFL is less than 5 years old, intermediate is between 5 and 10 years, and stabilized LFL is aged 10 years or more. In the case of acidity, young LFL is usually acidic with pH of less than 6.5, intermediate LFL is more neutral and the pH is between 6.5 and 7.5, and the pH of stabilized or old LFL is usually more than 7.5. Over time, the biodegradability of leachate decreases, with the highest degradability associated with the young leachate. Also, the Kjeldahl nitrogen concentration of young leachate is between 0.1 and 0.2 g/l and decreases with time. But ammonia nitrogen has a different story, starting with values of less than 400 mg/l for young leachate and possibly increasing at a later stage. The TOC/COD proportion is less than 0.3 for young leachate and increases to more than 0.5 for old leachate. Heavy metals have a low concentration for various types of leachate, but they are important because of toxicity. BOD5/COD is another important proportion which is reported at 0.5 to 1 for young leachate, between 0.1 and 0.5 for intermediate leachate, and less than 0.1 in old leachate. This shows that BOD5/COD decreases as the leachate stabilizes. COD is also a very important parameter and decreases to less than 4000 mg/l for old leachate from values of over 10,000 mg/l when the leachate is young (He et al. 2015). Landfills are identified as primary threats for groundwater resources due to disposal of waste containing metals. So leachate from landfills presents a potential health risk to both surrounding ecosystems and human populations (Mor et al. 2006; Al Sabahi et al. 2009). Due to the lack of landfills drainage systems, the leachate containing threatening and toxic substance (such as heavy metals) may penetrate into groundwater tables (Baun and Christensen 2004). Several researchers from around the world have studied LFL treatment by different methods. One of the effective and reliable treatment technologies that can be used for LFL treatment is the application of MBR. In comparison to conventional treatment technologies, this method has many advantages such as flexibility in operation, complete retention of solids, and reduced sludge production (Boonnorat et al. 2016). MBRs essentially consist of two primary parts: first the biodegradation of the waste compounds in the biological unit or bioreactor and, next, the separation of the treated water from microorganisms via the membrane module (Ahmed and Lan 2014). It seems that reverse osmosis (RO) is another very efficient method for landfill leachate treatment. Removal percentages of more than 98% and 99% have been reported for COD and heavy metals, respectively. Ultrafilters (UF) are usually used as a pretreatment for removing the larger molecular weight components of leachate that may cause fouling on the reverse osmosis membranes and subsequent operational problems. Nanofiltration (NF) has a high rejection rate for divalent and multivalent constituents in the liquid such as sulfate ions but a low rejection rate for chlorides and

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sodium. Conveniently, most of the heavy metals are multivalent cations and are rejected. On the other hand, the monovalent cations are comparatively harmless species (such as potassium and sodium) and pass through the NF. In one study, for example, the rejection of cadmium, zinc, lead, and chromium was higher than 70%, but the rejection of sodium and potassium was less than 10% (Peng et al. 2014). This chapter aims to review the research pertaining to the application of MBR systems to remove organic matter, nutrients, and heavy metals from LFL.

MBR and Its Application in BOD5, COD Removal The ability to treat difficult polluted leachate and handle high organic loadings is some of the advantages of the MBR system which make it attractive. BOD5 removals between 90% and 99% were achieved in most studies, depending on the operating conditions and leachate age. According to Insel et al. (2013), who used a screen for MBR pretreatment, followed by NF/RO and MBR as treatment of young leachate, approximately 90% of the COD was removed. The feed COD was about 19 g/L which was reduced to approximately 2 g/L after the MBR. The conductivity was unchanged before and after the MBR. After application of NF and RO, the values of the COD could be decreased to 300 and 180 mg/L, respectively. The experiments related to young landfill leachate reported that 17% of the total COD in the influent can be regarded as the soluble inert COD. In addition, the COD which is biodegradable made up approximately 70% of the total COD. This value was higher than the measured BOD5, which equaled to less than 50% of the total COD (Fig. 1). Likewise, MBR technology has been used to treat compost leachate. Although most studies have focused on MBR independently, an Iranian study used an advanced process of pre-anaerobic/aerobic treated compost leachate, reducing BOD5 and total COD to below the required guideline values. In the research carried out by Hashemi (2015), the reactor was a combination of MBR and SBR systems and was fed with treated leachate with overall 70–1360 mg/l chemical oxygen demand (COD). The process operation was divided into 5 phases: feeding

Fig. 1 The application of MBR followed by NF and/or RO for landfill leachate treatment (Insel et al. 2013)

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(15 min), aeration (12–22 h), settling (1 hour), filtration (30 min), and membrane relaxation (15 min). Posttreatment processes, such as advanced oxidation processes (AOPs), reverse osmosis (RO), and nanofiltration (NF), were proposed as possible routes for polishing the MBR filtrate. The removal efficiency of COD in experiments reached and even exceeded 80%. Also, up to 99% of the solids, which include colloidal solids, were removed due to the microporous structure of the membranes. As expected, there was no significant difference between TDS concentrations in the feed and filtrate. Therefore, TDS values of the filtrate were higher than the permitted limit. In addition, there weren’t any important differences in the quality of product as the feed concentrations varied. Nevertheless, under high loads, membrane fouling led to filtrate flux loss and an increase in the frequency of membrane cleaning and replacement. The acceptable performance under different conditions observed shows the promising capability of a full-scale, on-site MBR system. Thanh et al. (2013) worked on a low flux submerged membrane bioreactor to treat leachate. The organic loading rate (OLR) varied in the range of 2–10 kg COD/m3. day. The submerged membrane bioreactor (S-MBR) showed good COD removal (more than 90%). When pretreatment prior to the membranes is done via an anaerobic process, the COD removal will already be 70–80%. The efficiency of overall COD removal was always more than 90%. The highest removal efficiencies of ammonia and TN were 92  1.52% and 88  1.8% at a flux of 3.8 LMH, respectively. Also, the fouling rates of 0.075, 0.121, 3.186, and 6.374 kPa/day were observed when fluxes equaled 1.2, 2.4, 3.8, and 5.1 LMH, respectively. When treating high-strength leachate from a solid waste transfer station, a flux of lower than 2.4 LMH was able to be sustained. Syron et al. (2015) worked on performance analysis of a pilot-scale membrane aerated biofilm reactor with a volume of 60 liters. During a year of research, the average hydraulic retention time was about 5 days, and the value for the total suspended solids (TSS) was typically around 300 mg/L. This pilot study showed that landfill leachate full of ammonium could be effectively nitrified using a membrane aerated biofilm reactor (MABR). The pilot-scale MABR showed high levels of nitrification when hydraulic retention times of more than 4 days were applied, as compared with 40 days in full-scale SBRs at the landfill treatment site. Analysis of the operational data demonstrated that the MABR was not limited with the delivery of oxygen but rather by transportation of ammonium to the biofilm-covered membranes. The predictions of a multi-species AQUASIM model supported this observation. Also, the model suggested that higher activities of biofilm are possible. Shorter hydraulic retention times could also be possible with better module design and adequate nutrition. The COD concentrations in influent ranged from 1 to 3 g/L and were reduced to approximately 200–500 mg/L. The oxygen transfer rates achieved during the study were as high as 35 g O2/m2/day. Without any negative impact on oxygen transfer rates, high oxygen transfer efficiencies were achieved by operating at low gas flow rates. It was suggested that the biofilm was not limited by oxygen. Furthermore, according to the results, it is obvious that, with process optimization, a lower-energy usage can ensue and that MBR technology presents a low-energy option to treat leachate effectively. Finally, considering the aeration

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efficiency analysis of the MABR, SAEs in excess of 10 kg O2/kWh are feasible. This means that with this technology, there is an opportunity to reduce the costs of operation for aeration by an order of magnitude over other bubble-based aeration technologies. In other research, Wang et al. (2014) studied the role of anoxic/aerobic granular active carbon in assisting MBR. This was integrated with reverse osmosis and nanofiltration for advanced treatment of municipal landfill leachate (Fig. 2). The researchers had two lab-scale MBR systems with and without granular activated carbon (GAC). The feed water was old landfill leachate. This study investigated the impact of GAC performance on membrane permeability and sludge flock characteristics. Results showed that the presence of GAC can improve the removal of heavy metals (Cd, Cu, and Cr) and COD effectively. It is very important that after adding GAC, a decrease in membrane fouling was observed, and it was proposed that this was mainly due to the improvement of compactability of flocs and settleability, as well as an increase in particle size. It was found that the high-valence metal ions attached more easily on the sludge flocs in comparison with low-valence ions in both MBR systems. Although the total salt rejection by the MBR was understandably low, the NF membrane demonstrated P

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Fig. 2 Diagram of A/OMBR followed by NF/RO, (a) with and (b) without granular activated carbon addition (GAC) (Wang et al. 2014)

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excellent removal efficiencies of organic contaminants and color. NF was an excellent option, but the RO membranes not only showed better salt rejection but also held a higher and steadier membrane flux in comparison with other cases reported in landfill leachate treatment. The average removal of COD in the anoxic bioreactor was approximately 25%. Integration of the effluent of the anoxic bioreactor with the aerobic MBR raised the average NH3-N and COD removal efficiency to 93.2 and 81.5%, respectively. To compare with regular MBR processes, here, a higher removal of COD was achieved, showing that the addition of GAC was useful in the activity of microorganisms for the degradation of organic pollutants and perhaps the adsorption of contaminants. In summary, adding GAC not only increased the removal of heavy metals and hazardous organic pollutants but also helped bioflocculation and particle size of flocs, which greatly reduced membrane fouling severity. In 2012, Brito et al. worked on a lab-scale MBR to treat landfill leachate containing Saccharomyces cerevisiae (yeast). The particular setup that was employed showed good COD removal efficiency and less propensity to fouling. Average removal efficiencies of COD, color, and humic acid were 74.04%, 82.08%, and 67.17%, respectively. In this method, chemical cleaning by NaOCl was carried out every 20 days. He et al. (2015) worked on the effect of Fenton oxidation on biotoxicity, biodegradability, and the dissolved organic matter distribution in landfill leachate. The stages of this process consisted of a stabilization pool, screen, sedimentation, UASB reactor, A/O reactor, MBR system, UF, and NF. In this research, the effect of the Fenton process on dissolved organic matter (DOM) distribution (as fulvic acid (FA), humic acid (HA), and the hydrophilic fraction (HyI)), the chemical forms of toxic organic compounds and metals, and their biotoxicity were investigated. Apparently, the ability of DOM complexing differs with types of metal in the leachate. The biotoxicities of the DOM for luminescent bacteria were in the order of HA > FA > HyI. The results showed that the Fenton process could affect the distribution of pollutants in DOM and their biotoxicities by the breakdown of HA and FA in the leachate. Also, it was proposed that metals are the main cause of the biotoxicity to luminous bacteria. These results showed that an important factor in the selection of a treatment process for concentrated leachate after the Fenton process is the effect of metals on microorganisms. Akgul et al. (2013) have attempted to treat landfill leachate via a UASB-MBRSHARON configuration. This sequential use of UASB and MBR systems provided 90% COD and 99% BOD5 removal from young leachate. The average removal of NH4N and NO2N efficiencies reached over 75% and 90%, respectively, for the entire process after feeding the Anammox reactor with the effluent transferred from the SHARON process. The results indicated that a combined UASB-MBR-SHARON and Anammox configuration has the ability to provide effluent of adequate quality reliably. Even when the influent BOD5 concentration was over 8000 mg/L, the concentration of BOD5 in the effluent decreased to 50 mg/L by the UASB and MBR hybrid. In most cases, the COD and BOD removals were above 90%.

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In 2012, Boonyaroj et al. removed toxic organic micropollutants in a long-term operated membrane bioreactor for treatment of leachate. The operation of the MBR with two stages was continuously studied for 300 days under long sludge age conditions. AMBR utilizing inclined plate separator in the first stage (anoxic) followed by the second stage (aerobic) submerged MBR was developed and operated satisfactorily and steadily with high efficiency. The phthalic acid esters (PAEs) and phenolic compounds in the landfill leachate and the treated water after the MBR system were quantified by solid-phase extraction and gas chromatography–mass spectrometry, showing a removal range of 77–96%. Also, it was discerned that biodegradation is the main mechanism for removing BPA and BHT, whereas DEHP was mostly removed via adsorption onto sludge particles. BHT was highly retained in comparison with BPA and DEHP. This research showed the possibility of the application of membrane bioreactors to remove low concentrations of organic micropollutants in landfill leachate under high biomass concentration and long sludge retention times. It showed that the removal of organic micropollutants in landfill leachate was improved under longer sludge age conditions and higher biomass concentration. The experimental results showed that a two-stage MBR system can provide more than 85% removal for COD, BOD, and NH3N. It was found that the sludge volume index (SVI), which is a measure of settling properties, was relatively constant between 20 and 30 during the first stage of operation, meaning that settling is good. Analysis of particle size showed that biological flocs in the aerobic reactor were at first smaller and became larger during the first 100 days from about 30 μm to more than 100 μm. With time, despite the increase in SVI, the MBR sludge still indicated satisfactory settleability. Reverse osmosis with aerobic S-MBR systems enhanced by Fenton oxidation for advanced treatment of old municipal landfill leachate (older than 10 years) was conducted by Zhang et al. in 2013. The Fenton process was designed to degrade the non-biodegradable organic pollutants in the leachate. The combined process of Fenton–S-MBR–RO which is new not only achieved a high-quality effluent (NH3-N less than 15 mg/L and COD less than and 8 mg/L) meeting the wastewater reuse standards but also reduced membrane fouling notably. The permeability of the membrane was increased approximately threefold according to the decreased zeta potential of the colloids, the reduction in the EPS content, and the increase in particle size. Moreover, the fact three kinds of RO membranes were tested, all keeping a constant high flux (less than 23% attenuation), meant that many of the species that contribute to fouling were removed via the combined S-MBR Fenton process. Boonnorat et al. (2014, 2016) used a membrane bioreactor (two-stage) for long duration treatment of highly contaminated municipal solid waste leachate. The experiments were conducted for a long duration of 500 days. Phenolic and phthalate were highly removed (>95%). In addition, the removal of bisphenol A (BPA), 2,6-di-tert-butylphenol (2,6-DTBP), 2,6-di-tert-butyl-4methylphenol(BHT), di-ethyl phthalate (DEP), di-butyl phthalate (DBP), and bis (2-ethylhexyl) phthalate (DEHP) was detected in the membrane bioreactor (MBR with two-stage) by operating under no sludge wastage condition. The removals of the aforementioned contaminants were 99.5%, 99.0%, 99.5%, 97.9%, 96.8%, and 95.7%, respectively,

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under long-term operation in the system. The key factor that affected the removal efficiency of emerging contaminants appeared to be the chemical properties of the species themselves. The six compounds, with relative hydrophilicity and high solubility in water, were found to have higher biodegradation. Also, DEHP’s molecular structure is more complex, suggesting that it could initially be removed via adsorption, followed by subsequent biodegradation. The following equation was applied for determination of the rate constant for the removal of emerging contaminants by the combined adsorption and biodegradation mechanism. This was done for the active sludge case: ln

qt ¼ kt t q0

where q0(g/l) and qt(g/l) are the concentrations of contaminants in the solution containing water at time = 0 and time = t (hours) and kt(1/h) is the first-order rate constant for the removal of the contaminants. The following equations were applied for modeling the adsorption for emerging contaminants onto the MBR sludge: dqt ¼ k1 ð qe  qt Þ dt dqt ¼ k2 ðqe  qt Þ2 dt where qe(g/g) is the solid-phase concentration of the emerging contaminants at equilibrium, representing the maximum adsorption capacity of the sludge, qt(g/g) is the solid-phase concentration at contact time t, and k1 and k2 are the pseudo-firstorder and pseudo-second-order rate constant, respectively. The average influent concentrations for the BOD and COD in the study were 7402 (SD = 1203) mg/l and 13,531 (SD = 1861) mg/l, and the food to microorganisms (F/M) ratio in the system was between 0.4 and 0.7 1/d. After the treatment, the BOD and COD concentrations in the effluent averaged 36 (SD = 20) mg/l and 648 (SD = 310) mg/l, respectively. This meant that the applied treatment resulted in average removal efficiencies for BOD and COD of 99.5% and 95.2%. The authors showed that the majority of the amount of organic removal (about 65%) took place in the anoxic reactor. Other researchers have previously used pretreatment and posttreatment to improve COD removal such as Hasar et al. (2009), who used coagulation and ammonium stripping as pretreatment, followed by MBR and finally RO. The COD in the feedwater was between 8500 and 19,200 mg/l, and it was reduced to 4 mg/L after the RO. A forward osmosis membrane system as the posttreatment for MBR-treated landfill leachate was used by Dong et al. (2014). During long-term application, the FO could remove about 98.6% of the COD and 96.6% of the TP. In comparison between NF/RO, FO has several advantages. For example, there is no need for hydraulic pressure, high rejection of pollutants is achieved, and there is low

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membrane fouling propensity. In the past decade, FO has been used for the treatment of industrial and municipal wastewater and oil and gas wastewater, but its application for landfill leachate treatment has been limited. The FO membrane used in this study was made from cellulose triacetate (CTA) with embedded polyester mesh. The mesh does not take part in the treatment, as is used as a mechanical support. With the following equation, the rejection of pollutants in the FO system can be calculated: R¼

VC  100% V 0C0

where V0 and V are the initial and the final volumes of the feed solution and C0 and C are the initial and final concentrations of the pollutants measured in the feed solution, respectively. Amaral et al. (2015) studied the application of NF following MBR treatment of landfill leachate. The pilot plant in the study had a treatment capacity of 3m3/d. In their study, the suggested landfill treatment configuration includes an MBR, followed by air stripping, and NF. In the end, the treated leachate has reached reuse-quality for applications such as in earthworks for construction work. The NF was very important to make certain of the high quality of the product. The reduction of permeability in the NF membrane was relatively low meaning that the effluent of the MBR had satisfactory characteristics. The low fouling was also due to the fact that the operation was performed under subcritical conditions, which was important to ensure low demand for chemical cleaning procedures, which is also important to ensure longer membrane life. This system showed excellent leachate treatment performance, especially regarding the removal of chemical oxygen demand (80–96%) and color (98–99.9%). In 2013, Brown et al. used MBR technology to treat compost leachate. The setup was operated for 39 days after which it reached steady-state operation. Many waterquality parameters were detected during the experimental period. The COD of the effluent was reduced successfully by more than 99%, and this concentration was 116 g/ L in initial leachate. Other species which exhibited successful reductions were ammonia, caffeine, and many of the metals studied. This research showed the applicability of MBR for treating compost leachate. A continuous supply of air was introduced to the tank under the membranes at 1100 L/h. On the final day of the experiment (day 39), the COD had fallen to 400 mg/L which shows a 99.7% reduction compared to the initial COD. Sludge removal and membrane cleaning can help to improve COD removal and help the system performance in large-scale applications. Xue et al. (2015) have compared between MBR and other systems for the treatment of waste leachate in submerged and recirculated membrane bioreactors as seen in Fig. 3 (S-MBR and R-MBR). In both S-MBR and R-MBR, the removal efficiency of COD increased to more than 93% in 100 days of operation. The fouling rate of the R-MBR was lower than in S-MBR system. This study compares clearly the performances of the R-MBR and S-MBR for treatment of fresh leachate. The efficiency of treatment for both reactors was quite similar, and their removal efficiency for COD and NH4-N was about 90%.

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Substances similar to protein are easily degraded by both reactors; in addition, both systems are efficient at removing sulfur and straight-chain alkanes. The permeate volume of the R-MBR was much higher than that of the S-MBR. The combined facts that R-MBR had both a higher cumulative permeate and lower fouling mean that it had superior performance. The membrane fouling mechanism indicated that the cake layer has more resistance in the S-MBR, but the cake layer accounted for about one-third of the membrane fouling resistance in the R-MBR. With microbial community analysis of the cake layers for the two MBRs, it was discerned that Bacteroidetes and Proteobacteria were the main groups of bacteria in both reactors. The diversity of microbes in the R-MBR cake layer was higher than that of other layers. The S-MBR does not have any circulation pumps during operation, so its energy consumption is low, but the membrane is more easily contaminated. In the R-MBR, the permeate flux is more than S-MBR, and the fouled membrane can be easily cleaned; but a circulating pump is somewhat responsible for its high flow, so the energy consumption is one of its major disadvantages. The recirculated MBR was operated in continuous operation at a constant transmembrane pressure (TMP) of 30 kPa. When the flux of the R-MBR had decreased by 70%, the membrane was cleaned. The membranes were soaked in a 0.1% NaOH solution for 120 min for the chemical cleaning, and then a 0.5% citric acid solution (pH ~ 2) for 120 min was used. The COD of the feed solution was between 700 and 1000 mg/l, and the BOD/COD was between 0.32 and 0.39. COD removal in the S-MBR and R-MBR exceeded 93% in both systems. In 2014, El-Fadel et al. worked on a comparative inspection of SBR and MBR performance to treat high-strength landfill leachate. The MBR showed excellent performance with removal efficiencies of more than 95% for BOD5. The removal efficiency of SBR almost reached 85% for BOD5 and not quite 50% for COD. Nivya and Minimol Pieus (2016) compared the combination of MBR and the photoelectro-Fenton process (PEF) for the treatment of landfill leachate. The inlet concentrations of COD and BOD5 were 19800–24,000 mg/l and 4250–5100.

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When only PEF treatment was applied, the percentage removals of TSS, BOD5, and COD from landfill leachate were 89.3%, 71.9%, and 83.6%, respectively. Using MBR as a follow-up to the PEF process resulted in the percentage removal of the same pollutant parameters to become 95.5%, 90.2%, and 96.2%, respectively. Moravia et al. (2013) also studied an advanced oxidative process (AOP), namely, the combination of Fenton’s reagent with membrane separation (MF/NF). Fenton’s reagent (AOP/Fenton), similar to all other AOP processes, produces highly reactive hydroxyl radicals. The radicals can destroy almost every organic pollutant. In this research, the minimum COD and minimum color in the inlet were 2220 ppm and 420 μH, respectively. The removal efficiency for various substances ranged from 50% to over 70%.

MBR and Its Application in Nutrient (Phosphorous and Nitrogen) Removal Alongside removal of BOD and COD, the removal of nutrients such as phosphorous and nitrogen has also been of interest to researchers. For example, in a study by Insel et al. (2013), the TKN and organic nitrogen contents of the feedwater were 2075 and 8308 mg/L, respectively, which were reduced to 310 and 105 mg/L after MBR treatment followed by NF/RO. In 2013, Akgul et al. worked on the treatment of landfill leachate by UASBMBR-SHARON–Anammox as previously mentioned. Although the main target of the researchers was to remove COD, other parameters were also studied. In this experiment, TKN removal efficiency was also over 90% in the combined processes. In spite of the fact that high concentrations of nitrogen were present in the raw leachate, the consecutive operations of SHARON and Anammox reactors were very effective in its removal. The Anammox is a process which works based on autotrophic denitrification via nitrite. In this process, Anammox bacteria combine nitrite and ammonium to form nitrogen which then exists in the system as a gas. The SHARON process involves partial nitrification of ammonium to nitrite in a system which is operated at low sludge retention time (usually about 1 day) at a temperature (35C) which is relatively high and a pH of 7–8. Since cells require phosphorous for growth and the leachate did not have adequate phosphorous content, the addition of phosphorous to the system was necessary. Hence, orthophosphoric acid was added for the required adjustment. The efficiencies of average nitrite nitrogen and ammonia nitrogen removal were 92% and 78%, respectively. As previously mentioned, MBR has also been applied for compost leachate, for example, in the study of Brown et al. (2013). Due to the high pollution load of NH3 in this kind of leachate, the treatment of compost leachate is of interest. In this study, the concentration of ammonia was reduced from 2720 mg/L in the inlet to under 0.1 mg/L in the effluent. However, this study was not successful at reducing the nitrate concentration and provided a short explanation as to why this was the case. In another study, El-Fadel et al. (2014) showed that MBR was a suitable choice for nitrogen removal, with removal efficiencies of more than 95% for TN and NH3. The study showed low levels of contaminants in the effluent.

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As stated in the previous section, the photoelectro-Fenton process (PEF) and the combination of PEF and MBR for landfill leachate treatment have been investigated (Nivya and Minimol Pieus 2016). The ammonia nitrogen removal was more or less the same (full removal) for both processes. However, phosphorous removal increased from 58% in the PEF process to above 80% in the PEF-MBR hybrid. Zhang et al. (2013) studied the aerobic S-MBR/reverse osmosis system as well as using Fenton oxidation for the advanced treatment of old municipal landfill leachate. This study also resulted in a high-quality effluent with NH3N concentrations of less than 8 mg/L. Boonnorat et al. (2014, 2016) used a membrane bioreactor with two stages for the treatment of municipal solid waste leachate which was highly contaminated. The removal efficiencies for NH3 and TKN were 94.5% and 75.2%, respectively. The particular forte of this study was its long duration of 500 days. Wastewater treatment with phosphorus recovery by a novel osmosis membrane bioreactor (OMBR) has been studied by Huang et al. (2015). The removals for PO3P, NH4-N, and TOC in the OMBR were 96%, 43%, and 100%, respectively. To reduce costs of chemical use for pH adjustment, the OMBR system was enriched with phosphate ions.

MBR and Its Application in Heavy Metal Removal Heavy metals, which seep out from the waste in landfills, can infiltrate surface and groundwater resources and subsequently can be concentrated in food chains and cause adverse effects on human health and the environment. Therefore, in order to avert damaging effects, the leachate must be treated to reach appropriate standards. In addition, proper design and operation of the landfill are required to ensure no leakage occurs. Since leachate quality varies in different landfill sites, they must be studied separately to achieve an appropriate treatment method. It is no surprise that heavy metal removal from leachate using MBR technology has also been studied. For example, Brown et al. (2013) monitored the heavy metal concentration in their studies and found that the MBR system removed more than 80% of the heavy metals (except for Copper). In 2014, Wang et al. showed that addition of GAC to the process has some advantages such as improving the removal of hazardous organic pollutants and heavy metals and enhancing bioflocculation and particle size of flocs, which greatly reduced the propensity of the membrane to foul. The results of a research carried out by Mahmoudkhani et al. (2014) indicated that MBR treatment of leachate can provide removals of Fe, Cu, and Cd. Removal of heavy metals by MBR was a function of the aeration ratio and bioaccumulation.

AnMBR or Anaerobic Membrane Bioreactor AnMBR or anaerobic membrane bioreactor is a kind of MBR which works in the absence of oxygen. AnMBR is less common than MBR for researchers because it works slowly, but it may become more appealing in the future because it consumes

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less energy (this is more or less obvious because there is no need for air blowers). AnMBR also generates methane which is very attractive for researchers who are interested in biogas. For example, Ng et al. (2015) have evaluated the microbial communities and system performance of an anaerobic membrane bioreactor for treatment of pharmaceutical wastewater. Although this study did not directly target landfill leachate, it provided some insight regarding AnMBR systems. Mahmoudkhani et al. (2011) studied the biological and physiochemical treatment characteristics of high-strength landfill leachate in Tehran municipality landfill, Iran. The feed was young leachate with an average COD of 68,000 ppm. The volume of the reactor was 175 L and used hollow fiber membranes with a pore size of 0.1 μm for filtration. The MBR was followed by RO treatment. The dissolved oxygen (DO) concentration was kept at 3.2 mg/L. In addition, the solid retention times (SRTs) and hydraulic retention times (HRTs) were maintained at 55 and 15 days, respectively. The average of COD concentration in the MBR effluent was 1733 mg/ L, and its average removal efficiency was over 95%. The average of NH4N removal efficiency was about 99%. Complete nitrification was achieved during this period. Average removal efficiency of PO4P in the effluent was as high as 90%. After all treatment processes were complete, the average COD concentration in the effluent (post RO) was 335 mg/L. The concentration of PO4P in the RO effluent was less than 1 mg/L. Table 1 provides a summary of the results obtained in the study. In another study (Nuansawan et al. 2016), under the presence and absence of sludge recirculation, a two-stage MBR system (first anaerobic followed by aerobic as shown in Fig. 4) was operated at hydraulic retention times (HRT) of 5 and 2.5 days for each stage. Due to the anaerobic conditions, more than 90% of CH4 emissions from this process came from the first reactor, whereas N2O emissions were more or less equal from both reactors. Increasing the hydraulic loading through sludge recirculation and shortened HRT caused reduced CH4 emission by 17–31%, while organic and nitrogen removal efficiencies were adversely affected. Analysis of the Table 1 AnMBR efficiency in removal of some leachate components (Mahmoudkhani et al. 2011)

Parameter COD (mg/L) PO4-P (mg/L) BOD (mg/L) NH4-N (mg/L) Cl- (mg/L) SO4 (mg/L) Conductivity (μmhos/cm) Turbidity (NTU)

Average effluent from MBR 1733 12.94 270.8 8.1

MBR removal efficiency (%) 97.46 90.05 99.39 99.45

Average effluent from the RO 335 0.86 62.5 7.74

Removal efficiency of MBR + RO (%) 99.13 99.33 99.85 99.51

8620 2060 40,680

41.76 62.55 7.85

7710 758 9044

47.9 86.21 79.51

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Anaerobic reactor Aerobic reactor P

Treated water storage tank

Fig. 4 AnMBR and aerobic reactor combination in leachate treatment (Nuansawan et al. 2016)

microbes suggested that as hydraulic loading increased, there were less diversity and DNA abundance of N2O- and CH4-producing microorganisms. In 2014, Xie et al. used a dynamic anaerobic membrane bioreactor (AnDMBR) for treatment of landfill leachate. A mixture of both young and old leachate was used. The removal of COD was 62.2% with an influent COD concentration of 13,000 mg/ L. The amount of methane in the biogas (as a by-product) was between 70% and 90%, and the average yield of methane was 0.34 L/(gCOD removed) at the organic loading rate of 4.87 kg COD/(m3/d). The treatment of the organic fraction of municipal solid waste with a submerged anaerobic membrane bioreactor (SAMBR) was studied by Trzcinski and Stuckey (2010, 2016). The researchers focused on the treatment of leachate from the operation of the SAMBR for this type of high-strength wastewater with 5 days hydraulic retention time (HRT) and a 10 L/min biogas sparging rate. Under these conditions, the removal of COD was more than 90% during 4 months of operation, and there was no need for chemical cleaning of the membrane. In 2010, Zayen et al. used an anaerobic membrane bioreactor for the treatment of leachates without any pretreatment or posttreatment. Due to the lack of pretreatment, it is understandable that fluxes were low and frequent chemical cleaning was required. The organic loading rate (OLR) in the AnMBR increased slowly from 1 gCOD/(L.d) to above 6 gCOD/(L.d). At the highest OLR, the biogas production was more than 3 volumes of biogas per volume of the bioreactor. Under stable conditions, the efficiency of treatment was high with COD reduction of about 90% and biogas yield of 0.46 liters of biogas per gram of COD removed. Hydrocarbons and organics were efficiently eliminated. At the beginning of each phase, when the OLR was increased, there was an increase of VFA concentrations with a decrease in the efficiency of removal; but the system quickly recovered and adapted to the new conditions. The bacteria showed a slow growth rate, and the VSS reached only 3 g/L after 5 months of operation.

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Yiping et al. (2008) studied the removal of organic micropollutants with AnMBR. In this research, removals of 17 organochlorine pesticides (OCPs), 16 polycyclic aromatic hydrocarbons (PAHs), and 4-nonylphenol (4-NP) were investigated, and it was shown that the removal efficiency for these pollutants was OCPs (94%) > 4-NP (77%) > PAHs (59%). Meanwhile, removals for the more common parameters of the wastewater were BOD5 (99%) > COD (89%) > TOC (87%). Also, the majority of the PAHs in the leachate were removed.

The Effect of pH and Recycling on Efficiency of MBR Systems Many parameters can affect the operation of MBR systems, so it seems suitable to end this chapter with a summary of these effects. Boonyaroj et al. (2017) worked on enhanced biodegradation of phenolic compounds with enriched sludge with a MBR with two stages as shown in Fig. 5. During reactor operation, efficiencies of organic removal were more than 90%, and phenolic compounds (as bisphenol A (BPA) and 4-methyl-2,6-ditertbutylphenol (BHT)) were removed by 65 and 70%, respectively. It is suggested that this removal is mainly through aerobic biodegradation even at high feed concentrations of 1000 g/L for both compounds. During the MBR operation with two stages in laboratory scale, average BOD5 and COD concentrations in the landfill leachate feed were 6598 and 9273 mg/L, respectively. In 2013, Ng et al. worked on optimizing MBR systems by incorporating powdered activated carbon (PAC) in the process. This investigation was divided into two sections. First, the effect of PAC on MBR was studied; second, the effect of sludge retention time (SRT) on MBR operation with and without PAC was investigated. Although this paper did not treat landfill leachate, the novel process has been

Anoxic tank

F

Aerobic tank

Leachate preparation tank

P

F P

P

Treated water storage tank

F Air blower

Fig. 5 Anoxic MBR and aerobic reactor combination in leachate treatment (Boonyaroj et al. 2017)

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mentioned herein due to its merits. Perhaps future researchers could apply this process for landfill leachate treatment. In 2015, Sanguanpak et al. worked on the effects of acidity in mixed liquor (pH of 5.5, 6.5, 7.5, and 8.5) on fouling behavior and the landfill leachate biodegradability for 150 days. In this research, two identical lab-scale MBRs were used: one under acidic conditions (pH of 5.5 and 6.5) and the other for basic mixed liquor (pH of 7.5 and 8.5). The results from the experiments indicated that the removal of organic constituents was not highly dependent on the pH, although the microbial community did show variations. For example, dissimilar proportions of Proteobacteria between the pH of 5.5 and 8.5 influenced the transformation of dissolved organic matter (DOM). Nevertheless, a significant increase in inorganic removal was seen as the mixed liquor’s pH was increased (due to sedimentation). The fouling on the membrane was found to be most undesirable at the extremes (pH of 5.5 was worst, followed by pH of 8.5), but the fouling at pH 6.5 and 7.5 was more or less similar. Another property which was greatly influenced by pH was the sludge characteristics. In addition, protein-like substances were the main material of the foulant layer on the membrane surface at pH 5.5. As expected, inorganic precipitation and humic-like substances were predominant in the foulant layer at pH 8.5. Chemical cleaning of the membranes showed that irreversible fouling (the proportion of foulants which are irremovable) increases with the increase in mixed liquor pH. Ratanatamskul and Nilthong (2009) worked on the application of a hybrid biological powder activated carbon-membrane bioreactor (BPAC-MBR) system for the treatment of old leachate. High removal efficiencies for COD (83%), color (85%), TKN (97%), and TP (68%) were attained. The system was operated under aeration and lack-of-aeration modes, meaning that air was blown for a particular duration of time and then not blown for an equally long duration. This allowed aerobic and anaerobic conditions to prevail in sequence. Efficiencies of removal for the BPAC-MBR system in terms of organic matter, color, and nutrients at 150–150 min aerobic/anaerobic operation were higher than those at 120–120 min. Elcik et al. (2016) showed that the cake resistance was highest for UF membranes made of PVDF. When cross-flow velocity increased from 0.35 to 1.57 m/s, all the filtration resistances decreased, except for pore blocking resistance. The results from experiments indicated that the cross-flow velocity affected the fouling process more than the membrane properties. Finally, the following conclusions were drawn (Elcik et al. 2016): (1) At the beginning of each filtration period, membrane fouling due to the deposition of microalgal cells on the membrane surface was seen. (2) Microalgal cells caused both irreversible (irrecoverable) and reversible fouling for all membranes. The UH050 membrane, made from PEHS material, showed better performance in terms of reversible fouling when compared to all other membranes studied. (3) All the filtration resistances decreased when the cross-flow velocity increased.

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(4) The surface of the membranes was coated with numerous microalgae. The interior of the membrane pores was not blocked by cells because the cells have a larger diameter than the membrane pores. (5) According to the results, polysaccharides, proteins, and lipids played considerable roles in membrane fouling.

References Ahmed FN, Lan CQ (2014) Treatment of landfill leachate using membrane bioreactors: a review. Desalination 287:41–54 Akgul D et al (2013) Treatment of landfill leachate using UASB-MBR-SHARON–Anammox configuration. Biodegradation 24(3):399–412 Al Sabahi E, Rahim SA, Wan Zuhairi W, Al Nozaily F, Alshaebi F (2009) The characteristics of leachate and groundwater pollution at municipal solid waste landfill of Ibb City, Yemen. Am J Environ Sci 5(3):256–266 Amaral MCS et al (2015) Nanofiltration as post-treatment of MBR treating landfill leachate. Desalin Water Treat 53(6):1482–1491 Baun DL, Christensen TH (2004) Speciation of heavy metals in landfill leachate: a review. Waste Manag Res 22(1):3–23 Boonnorat J et al (2014) Removals of phenolic compounds and phthalic acid esters in landfill leachate by microbial sludge of two-stage membrane bioreactor. J Hazard Mater 277:93–101 Boonnorat J et al (2016) Kinetics of phenolic and phthalic acid esters biodegradation in membrane bioreactor (MBR) treating municipal landfill leachate. Chemosphere 150:639–649 Boonyaroj V et al (2012) Toxic organic micro-pollutants removal mechanisms in long-term operated membrane bioreactor treating municipal solid waste leachate. Bioresour Technol 113:174–180 Boonyaroj V et al (2017) Enhanced biodegradation of phenolic compounds in landfill leachate by enriched nitrifying membrane bioreactor sludge. J Hazard Mater 323:311–318 Brito GCB et al (2012) Treatment of landfill leachate in membranes bioreactor with yeast (Saccharomyces cerevisiae). Procedia Eng 44:934–938 Brown K et al (2013) Membrane bioreactor technology: a novel approach to the treatment of compost leachate. Waste Manag 33(11):2188–2194 Campagna M et al (2013) Molecular weight distribution of a full-scale landfill leachate treatment by membrane bioreactor and nanofiltration membrane. Waste Manag 33(4):866–870 Dong Y et al (2014) A forward osmosis membrane system for the post-treatment of MBR-treated landfill leachate. J Membr Sci 471:192–200 Elcik H, Cakmakci M, Ozkaya B (2016) The fouling effects of microalgal cells on crossflow membrane filtration. J Membr Sci 499:116–125 El-Fadel M, Hashisho J (2014) A comparative examination of MBR and SBR performance for the treatment of high-strength landfill leachate. J Air Waste Manage Assoc 64(9):1073–1084 Hasar H et al (2009) Stripping/flocculation/membrane bioreactor/reverse osmosis treatment of municipal landfill leachate. J Hazard Mater 171.1:309–317 Hashemi H (2015) Increasing of leachate quality using an integrated aerobic membrane bioreactor. J Adv Environ Health Res 3.1 He R et al (2015) Effect of Fenton oxidation on biodegradability, biotoxicity and dissolved organic matter distribution of concentrated landfill leachate derived from a membrane process. Waste Manag 38:232–239

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Insel G et al (2013) Biodegradation characteristics and size fractionation of landfill leachate for integrated membrane treatment. J Hazard Mater 260:825–832 Kreith F (1999) Handbook of solid waste management. McGRAW-HILL, California Mahmoudkhani R et al (2011) Study on high-strength anaerobic landfill leachate treatability by membrane bioreactor coupled with reverse osmosis. Int J Environ Res 6.1:129–138 Mahmoudkhani R et al (2014) Copper, cadmium and ferrous removal by membrane bioreactor. APCBEE Procedia 10:79–83 Mor S, Ravindra K, Dahiya R, Chandra A (2006) Leachate characterization and assessment of groundwater pollution near municipal solid waste landfill site. Environ Monit Assess 118(1–3):435–456 Moravia WG, Amaral MCS, Lange LC (2013) Evaluation of landfill leachate treatment by advanced oxidative process by Fenton’s reagent combined with membrane separation system. Waste Manag 33(1):89–101 Ng CA et al (2013) Optimization of membrane bioreactors by the addition of powdered activated carbon. Bioresour Technol 138:38–47 Ng KK, Shi X, Ng HY (2015) Evaluation of system performance and microbial communities of a bioaugmented anaerobic membrane bioreactor treating pharmaceutical wastewater. Water Res 81:311–324 Nivya TK, Minimol Pieus T (2016) Comparison of photo ElectroFenton process (PEF) and combination of PEF process and membrane bioreactor in the treatment of landfill leachate. Procedia Technol 24:224–231 Nuansawan N et al (2016) Effect of hydraulic retention time and sludge recirculation on greenhouse gas emission and related microbial communities in two-stage membrane bioreactor treating solid waste leachate. Bioresour Technol 210:35–42 Peng, Y (2014) Perspectives on technology for landfill leachate treatment. Arab J Chem Ratanatamskul C, Nilthong N (2009) Performance of biological powder activated carbonmembrane bioreactor (BPAC-MBR) for old-landfill leachate treatment. Int J Environ Waste Manag 4(3–4):271–281 Sahinkaya E et al (2013) Use of landfill leachate as a carbon source in a sulfidogenic fluidized-bed reactor for the treatment of synthetic acid mine drainage. Miner Eng 48:56–60 Sanguanpak S et al (2015) Influence of operating pH on biodegradation performance and fouling propensity in membrane bioreactors for landfill leachate treatment. Int Biodeter Biodegr 102:64–72 Syron E, Semmens MJ, Casey E (2015) Performance analysis of a pilot-scale membrane aerated biofilm reactor for the treatment of landfill leachate. Chem Eng J 273:120–129 Thanh BX, Dan NP, Visvanathan C (2013) Low flux submerged membrane bioreactor treating high strength leachate from a solid waste transfer station. Bioresour Technol 141:25–28 Trzcinski AP, Stuckey DC (2010) Treatment of municipal solid waste leachate using a submerged anaerobic membrane bioreactor at mesophilic and psychrophilic temperatures: analysis of recalcitrants in the permeate using GC-MS. Water Res 44(3):671–680 Trzcinski AP, Stuckey DC (2016) Effect of sparging rate on permeate quality in a submerged anaerobic membrane bioreactor (SAMBR) treating leachate from the organic fraction of municipal solid waste (OFMSW). J Environ Manag 168:67–73 Wang G et al (2014) Anoxic/aerobic granular active carbon assisted MBR integrated with nanofiltration and reverse osmosis for advanced treatment of municipal landfill leachate. Desalination 349:136–144 Xie Z et al (2014) An anaerobic dynamic membrane bioreactor (AnDMBR) for landfill leachate treatment: performance and microbial community identification. Bioresour Technol 161:29–39 Xue Y et al (2015) Comparison of the performance of waste leachate treatment in submerged and recirculated membrane bioreactors. Int Biodeter Biodegr 102:73–80

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Yiping X et al (2008) Occurrence and removal of organic micropollutants in the treatment of landfill leachate by combined anaerobic-membrane bioreactor technology. J Environ Sci 20(11):1281–1287 Zayen A et al (2010) Anaerobic membrane bioreactor for the treatment of leachates from Jebel Chakir discharge in Tunisia. J Hazard Mater 177(1):918–923 Zhang G et al (2013) Aerobic SMBR/reverse osmosis system enhanced by Fenton oxidation for advanced treatment of old municipal landfill leachate. Bioresour Technol 142:261–268

Part VIII Environmental Management of Biomaterials

Sustainable Biomedical Waste Management

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Sukanchan Palit and Chaudhery Mustansar Hussain

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . The Vision of This Treatise . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Literature Review . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sustainability Science and the March of Engineering and Technology . . . . . . . . . . . . . . . . . . . . . . . Environmental Sustainability, Industrial Wastewater Treatment, and the March of Science . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Vision of Biomedical Engineering and Waste Management . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sources of Classification of Biomedical Waste . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biomedical Waste Treatment and Disposal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Science of Solid Waste Management . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Eliminate Toxicity at Source . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Visionary Scientific Endeavor in the Field of Sustainable Waste Management . . . . . . . . . . . . . . Recent Scientific Research Pursuit in Biomedical Waste Management . . . . . . . . . . . . . . . . . . . . . . . Recent Scientific Endeavor in the Science of Sustainability . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Green Materials for Environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Green Technology, Wastewater Treatment, and Solid Waste Management . . . . . . . . . . . . . . . . . . . Future Research Trends and Future Frontiers of Waste Management . . . . . . . . . . . . . . . . . . . . . . . . The Challenge and the Vision of Sustainability and Solid Waste Management . . . . . . . . . . . . . . Conclusion and Future Scientific Perspective . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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S. Palit (*) Department of Chemical Engineering, University of Petroleum and Energy Studies, Energy Acres, Dehradun, Uttarakhand, India e-mail: [email protected]; [email protected] C. M. Hussain Department of Chemistry and Environmental Science, New Jersey Institute of Technology, Newark, NJ, USA e-mail: [email protected] © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_123

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Abstract

The domain of waste management and pollution control is today undergoing rapid and drastic changes. Human civilization and human scientific endeavor are moving toward a newer era of scientific regeneration. In the similar manner, biomedical engineering is surpassing visionary scientific frontiers. Solid waste management and biomedical engineering are the forerunners toward scientific regeneration and deep scientific vision. The present state of waste management is deeply challenged and is replete with scientific imagination. Industrial pollution today is the need of the hour. In this treatise, the authors deeply discussed the scientific success, the vast scientific potential, and the vision to move forward in the field of biomedical waste management. Human scientific research pursuit in waste management today needs to be re-envisioned and redefined. The challenge and the vision of this treatise are immense and far-reaching. The authors also in this well-researched treatise successfully elucidate the need for sustainable development in present-day human civilization. Waste management and environmental sustainability are the challenges of environmental engineering science today. This treatise opens up a new chapter in human scientific research pursuit in biomedical sciences and engineering with the sole aim and objective toward the furtherance of engineering science. Technological validation and the vast scientific ingenuity in the field of waste management are the other hallmarks of this treatise. The authors pointedly focus on the human scientific ingenuity and deep profundity in the field of biomedical engineering and waste management. Environmental sustainability is another vast area of research pursuit in this widely researched treatise. Keywords

Biomedical · Waste · Management · Sustainability · Development · Science · Engineering

Introduction Human civilization and human scientific research pursuit are the forerunners toward a greater visionary era in the field of biomedical engineering and waste management. Technology and engineering science are highly challenged today with the progress of scientific and academic rigor. Success of scientific endeavor is at stake in engineering science with the growing concerns for environmental and energy sustainability. Sustainable development is the imminent need of the hour. Medical science and biomedical engineering are the culminations of scientific endeavor today. Solid waste management in the health sector is today urging the scientific domain to gear toward newer vision and newer technologies. Scientific ingenuity, scientific forbearance, and deep scientific pragmatism are the needs of environmental engineering science today. The challenge and the vision of biomedical waste management are immense and groundbreaking. This waste management is veritably linked with sustainable development. In this treatise, the authors deeply comprehend

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the scientific intricacies and the vast scientific profundity in solid waste management. The science of biomedical engineering today stands in the midst of deep scientific vision, profundity, and forbearance. Solid waste management is the necessity of human civilization today. The entire visionary avenue of waste management needs to be re-envisioned and re-envisaged with the progress of human civilization. Industrial pollution control, wastewater treatment, and solid waste management are the visionary aisles of scientific research pursuit which need to be pondered with immense introspection and vision. Zero-discharge norms and environmental regulations are veritably challenging the scientific landscape today. This treatise deeply elucidates the scientific success, the vast scientific regeneration, and the scientific ingenuity behind the world of challenges in solid waste management in biomedical waste management. Scientific and academic rigor in solid waste management today is highly challenged, and technology and engineering science are in the midst of cross-boundary research. Interdisciplinary research is the veritable backbone of human scientific progress today. Chemical process engineering and environmental engineering science are today in the path of newer scientific rejuvenation. In this treatise, the authors rigorously focus on the vast scientific rigor in sustainable biomedical waste management with the sole objective and mission toward greater emancipation of environmental sustainability.

The Vision of This Treatise Human scientific research pursuit today stands in the midst of deep scientific hindrances and vast scientific introspection. Environmental pollution control and environmental sustainability are the scientific truth of today’s scientific research pursuit. Sustainable waste management and biomedical science and engineering are the hallmark of this research treatise. The vision of Dr. Gro Harlem Brundtland, the former Prime Minister of Norway, on the subject of sustainability needs to be redefined and re-envisioned. In this treatise, the authors pointedly focus on visionary scientific endeavor in sustainable waste management, biomedical waste management, and sustainability science. Technological and scientific validation are the cornerstones and pillars of this well presented treatise. Today human mankind is in a state of immense rejuvenation as well as scientific hindrances. In this treatise the challenge and vision of biomedical waste management and biomedical science are deeply investigated. The vision of science, the scientific cognizance, and the futuristic vision of biomedical engineering and its waste disposal issues are investigated in details in this treatise.

Literature Review Scientific vision and scientific ingenuity today stand in the midst of deep introspection and vast subtleties. The aisles of scientific endeavor in solid waste management are opening up new windows of innovation and instinct in decades to come.

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Technology and engineering science are highly challenged in today’s scientific paradigm. In this treatise, the author pointedly focuses on the vast potential and vision of engineering science of biomedical engineering with the sole mission and aim toward furtherance of science and technology. The profundity of science and technology today is evergrowing today. This section on literature review redefines and re-envisages the vast scientific aura and scientific candor in solid waste management and industrial pollution control in details. Technological vision and deep acuity are the hallmarks of scientific research pursuit in sustainable biomedical waste management today. Wilson et al. (2013) deeply discussed with lucid details integrated sustainable waste management in developing countries (Wilson et al. 2013). The vast technological vision, the futuristic vision, and the scientific success of sustainability science are presented in lucid details in this paper. This treatise uses the veritable lens of “integrated sustainable waste management” to examine how cities in developing countries have been tackling their immense solid waste issues (Wilson et al. 2013). Solid waste management and sustainable development today are opposite sides of the visionary coin in both developed as well as developing countries. Integrated sustainable waste management examines both the physical components (collection, disposal, and recycling) and the vast governance aspects which include financial sustainability and the implementation of proactive policies (Wilson et al. 2013). Tiwari et al. (Tiwari and Kadu 2013) in a lucid review delineated biomedical waste management practices in a developing country like India. Economic development of India in last two decades has resulted in immense environmental pollution and waste generation in immense quantity in India. In this treatise, an attempt is made to study the classification, legislation, and management practices in relation with biomedical waste in India (Tiwari and Kadu 2013). Human scientific endeavor and research forays are the pillars of modern science today. Solid waste management is a serious issue in a developing country context (Tiwari and Kadu 2013). This treatise upholds the scientific success, the deep scientific vision, and the technological challenges in the roads to scientific provenance and scientific destiny. Sustainable waste management is a vexing issue in developed world also. Sweden is a highly developed country which has formulated a comprehensive waste management paradigm (Tiwari and Kadu 2013). At present, waste management is far more resourceefficient and has less effect to the environmental pollution control issues of a nation. The techniques and tools envisaged since 1990s to achieve more resource-efficient use of waste have yielded immense and pragmatic results (Tiwari and Kadu 2013). Swedish Environmental Protection Agency Report (2005) deeply comprehended a deep strategy for sustainable waste management as a long-term waste plan. The authors deeply discussed targets and measures for sustainable waste management and the vast domain of Sweden’s waste plan (Swedish Environmental Protection Agency Report 2005). Human scientific endeavor and deep scientific farsightedness are the pillars of science and engineering today. Sustainability and solid waste management are today the utmost needs of the human society. Waste management in Sweden has undergone vast changes with the progress of scientific and academic rigor. Producers have been made responsible for the proper classification and total elimination of waste. European Union membership has deeply meant more detailed commitment governing hazardous wastes, landfill, and incineration (Swedish Environmental Protection Agency Report 2005).

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This report opens up new windows and new avenues of vision in the field of environmental sustainability in decades to come. According to the report, waste management is an environmental issue where waste is a resource as well as a problem. Technological vision and the science of environmental pollution control are today leading a long and visionary way in the true realization of environmental sustainability (Swedish Environmental Protection Agency Report 2005).

Sustainability Science and the March of Engineering and Technology Energy and environmental sustainability are challenging the scientific landscape as human civilization and human scientific endeavor march toward a newer visionary realm. Technology and engineering science of waste management in the similar manner are surpassing visionary scientific frontiers. Today, human civilization stands in the midst of deep scientific vision and scientific forbearance. The world of scientific challenges, the need for successful environmental engineering tools, and the futuristic vision of science will all lead a long and visionary way in the true emancipation and true realization of environmental sustainability today. In this treatise, the author rigorously points toward the research pursuit in the field of interface between environmental sustainability and solid waste management. Environmental and energy sustainability are today in a state of deep disaster as frequent environmental catastrophes and industrial water pollution are causes of immense concern to human civilization and human scientific endeavor. Throughout the world today, sustainable development is the hallmark of progress of scientific and academic rigor. The visionary words of Dr. Gro Harlem Brundtland, former Prime Minister of Norway, on sustainability science need to be redefined and re-envisaged with every step of human scientific rigor. Human vision, the vast scientific prowess of human mankind, and the veritable needs for progress of human civilization will go a long and visionary way in the true realization and the true emancipation of environmental sustainability today. In this treatise, the authors pointedly focus on the scientific success, the vast scientific vision, and the deep scientific truth behind waste management and the successful application of sustainability to human society.

Environmental Sustainability, Industrial Wastewater Treatment, and the March of Science Environmental sustainability and industrial wastewater treatment are today linked to each other by an unsevered umbilical cord. Technological advancements, the provision of basic human needs, and the answers to sustainability will all go a long and visionary way in the true realization of sustainability science today. Solid waste management and industrial wastewater treatment are the forerunners to the application of sustainability today. The march of science and technology is veritably linked to the scientific vision of sustainability as human civilization trudges forward toward

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a newer visionary realm. Human mankind and its scientific girth and determination are veritably challenging the scientific landscape today. Frequent environmental disasters, the vast technological profundity, and the scientific needs of human society are the torchbearers toward a newer visionary eon in the domain of both environmental engineering science and chemical process engineering. The authors in this treatise deeply comprehend the vast need of environmental sustainability in the future progress of human mankind. The march of engineering science today is evolving into new knowledge dimensions in the field of sustainability science. The future of application of environmental sustainability and the intricate and vast domain of industrial wastewater treatment needs to be envisioned and re-envisaged with the passage of scientific history and time.

Vision of Biomedical Engineering and Waste Management Biomedical engineering and waste management are the vexing issues which confronts human scientific endeavor today. Mankind’s immense scientific prowess, human civilization’s deep vision, and the global research and development initiatives will all lead a long and visionary way toward the true realization of science and technology today. Waste management is the need of the hour. Solid waste management in health sector globally is the utmost need of the hour. Scientific vision needs to be revamped and re-envisaged as industrial pollution control, solid waste management, and the true vision of industrial wastewater treatment enter into a newer visionary realm. Medical science and health science today stand in the midst of deep scientific discernment and vast scientific fortitude. Technological farsightedness and scientific motivation in the field of biomedical engineering are challenging the scientific domain. The vast vision of solid waste management is today witnessing drastic and dramatic challenges. In this treatise, the author rigorously focuses on the scientific success, the deep scientific profundity, and the vast scientific challenges in the research pursuit in waste management. Medical science is today in a state of immense revamping and deep scientific vision. In the similar manner, biomedical engineering science and solid waste management today stand in the midst of deep scientific comprehension and fortitude. In this treatise, the author deeply elucidates the immense scientific success, the scientific genesis, and the vast scientific profundity in the field of sustainable biomedical waste management. The vast scientific prowess of human mankind and the technological vision behind environmental pollution control are deeply discussed in lucid details in this paper.

Sources of Classification of Biomedical Waste Biomedical wastes disposal are vexing issues facing scientific domain today. Technology and engineering science are highly challenged today as human civilization and human scientific endeavor enter into a newer visionary arena. Medical

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science and medical technology are in the threshold of a newer scientific regeneration and vast scientific vision. In this treatise, the authors deeply comprehend the diverse areas of biomedical waste and mainly the vast world of solid waste management in biomedical engineering domain. Medical and healthcare wastes have alarmingly increased in recent decades due to the immense proliferation of human population, number and size of healthcare facilities, as well as the use of disposal medical products. According to the US Environmental Protection Agency (USEPA), medical wastes contain all waste materials generated by healthcare facilities, such as hospitals, clinics, physician’s offices, dental practices, blood banks, and veterinary hospitals/clinics, as well as at medical research facilities and dedicated laboratories that can include a wide range of materials, such as used needles and syringes, soiled dressings, body parts, diagnostic samples, blood, chemicals, pharmaceuticals, medical devices, and radioactive materials.

Biomedical Waste Treatment and Disposal Biomedical waste elimination today stands in the midst of deep scientific comprehension and scientific hindrances. Biomedical waste management and the vast domain of sustainable development are veritably relevant areas of scientific research pursuit today. In this section, the author rigorously points toward the immense scientific success, the technological profundity, and the innumerable challenges facing solid waste management today (www.google.com; www.wikipedia.com). The following are the different biomedical waste management techniques: • • • • • •

Incineration technology Non-incineration technology Autoclaving Microwave irradiation Chemical methods Plasma pyrolysis

The biomedical waste treatment processes encompass (1) chemical processes, (2) thermal processes, (3) mechanical processes, (4) irradiation processes, and (5) biological processes. Chemical processes involve processes that use chemicals that act as disinfectants (www.google.com; www.wikipedia.com). Sodium hypochlorite, dissolved chlorine dioxide, peracetic acid, hydrogen peroxide, dry inorganic chemical, and ozone are some of the many examples. Thermal processes are those processes which use heat as disinfectant. Depending on the temperature they operate, it has been grouped into categories, which are low-heat systems and highheat systems. Low-heat systems operate between 90  C and 177  C and use steam, hot water, or electromagnetic radiation to heat and eliminate wastes (www.google. com; www.wikipedia.com). Examples are autoclaving and microwaving. High-heat systems operate between 540  C and 8300  C (www.google.com; www.wikipedia. com). These techniques employ combustion and high-temperature plasma to

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decontaminate and permanently destroy wastes. Examples are hydroclaving and incineration. Mechanical processes involve compaction and shredding. Irradiation processes expose wastes to ultraviolet or ionizing radiation in an enclosed chamber (www.google.com; www.wikipedia.com). Biological processes use biological enzymes for treating medical wastes. The science of biomedical waste treatment is today witnessing drastic changes and vast challenges. Biomedical waste management needs to be sustainable as science and engineering moves forward toward a newer vision and newer innovations. Environmental sustainability and the success of environmental engineering tools are changing the face of human science and veritably crossing scientific frontiers (www.google.com; www.wikipedia.com).

Science of Solid Waste Management Solid waste management today stands in the midst of deep scientific vision and vast scientific emancipation. Environmental pollution control, industrial wastewater treatment, and the vast scientific success behind it are the scientific imperatives of today. Medical and health services arena are the challenges of solid waste management scenario today. Biomedical engineering and medical science are veritably in the threshold of a newer eon and a newer scientific regeneration. The challenge and the vision of biomedical engineering science are far-reaching and surpassing innovative scientific frontiers. The vision of science of solid waste management is today changing the scientific landscape of environmental engineering science and environmental pollution control. The vast domain of solid waste management according to international organizations needs to be deeply comprehended and re-envisioned. Biomedical waste or biowastes are those potential hazardous waste materials, consisting of solids, liquids, sharps, and laboratory wastes. Today the technology of biomedical engineering and solid waste management is vast, versatile, and far-reaching. This area of environmental engineering science needs to be vastly re-envisioned with the progress of science and technology (www.google.com; www.wikipedia.com).

Eliminate Toxicity at Source Elimination of toxicity at source is of utmost importance in the path toward progress in scientific and academic rigor in medical and biomedical science. Science, technology, and engineering are today ushering in a new era of scientific regeneration and deep scientific rejuvenation. Human scientific endeavor and scientific regeneration are in a state of immense difficulties and scientific hindrances. Biomedical engineering in the same manner are greatly in a path of scientific vision and regeneration. Toxicity is an alarming problem of biomedical waste management. Technology and engineering science of biomedical engineering are highly challenged and surpassing vast scientific boundaries. Medical science and biomedical engineering today are in the path of scientific rejuvenation and deep scientific

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introspection. In this treatise, the authors deeply comprehend the vast scientific farsightedness in today’s scientific forays in biomedical engineering (www.google. com; www.wikipedia.com).

Visionary Scientific Endeavor in the Field of Sustainable Waste Management Waste management paradigm is today in a state of immense scientific comprehension and scientific regeneration. Human civilization and human scientific research pursuit need to be re-envisioned and revamped with the progress of scientific rigor in medical science and biomedical engineering. Solid waste management and the deep scientific and academic rigor associated with it are highly challenged today. Waste management and environmental pollution control are the utmost needs of human civilization today. Environmental sustainability and industrial wastewater treatment are today the forerunners toward a greater emancipation of science and engineering today. In this section, the author pointedly focuses on the scientific success, the scientific vision, and the deep scientific fortitude in the application of sustainability and waste management in the furtherance of engineering science and technology (www.google.com; www.wikipedia.com). Solid waste is the unwanted or useless solid materials generated from combined residential, industrial, and commercial activities in a given locality. It may be categorized according to its origin (domestic, industrial, commercial, construction, or institutional), according to its contents (organic material, glass, metal, plastic paper, etc.), or according to its hazard potential (toxic, nontoxic, flammable, radioactive, infectious, etc.) (www.google.com; www.wikipedia.com). The vast challenge of technology and engineering science of solid waste management is delineated in minute details in this vastly researched treatise. Human scientific endeavor and human civilization are in the paved path of newer scientific rejuvenation and newer eon. Solid waste management and environmental engineering techniques are detailed in deep details in this treatise. Singh et al. (2014) discussed with deep and cogent foresight the sources, collection, transportation, and recycling of solid waste management. Solid wastes may be delineated as useless, unused, unwanted, or discarded material available in solid form. Semisolid food wastes and municipal sludge may be included in municipal solid wastes. Today solid waste management is a vexing issue globally. In the United States, over 180 million tons of municipal solid waste were generated in 1988 (Singh et al. 2014). To present before the reader the concept of solid waste management, an overview of municipal solid waste problems, sources, collection, resource recovery, and disposal methods is deeply comprehended in this paper (Singh et al. 2014). Technological acuity, scientific fortitude, and deep scientific enlivening are the hallmarks of this well-researched treatise (Singh et al. 2014). In this treatise, the author gives greater emphasis to the design and operation of municipal sanitary landfills, regulations governing land disposal and leachate generation, and treatment methods (Singh et al. 2014).

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Annepu (2012) in a well-researched treatise elucidated in minute details sustainable solid waste management in a developing country like India. This study rigorously examined the present status of waste management in India, its effect on public health and the environment, the concept of hygiene, and the vast prospects of introducing improved techniques of disposing solid waste in India. Technological validation, the vast scientific objectives, and the futuristic vision of environmental pollution control are the veritable forerunners toward a newer visionary eon in the field of sustainable waste management. The needs of human civilization and human scientific research pursuit are vast and visionary (Annepu 2012). The systems and techniques discussed are informal and formal recycling, aerobic composting and mechanical-biological treatment, small-scale biomethanation, refuse-derived fuel, waste-to-energy combustion, and landfill mining (Annepu 2012). Science is today moving in a definite direction as regards validation of engineering science and technology. The primary objective of this study was to find ways in which enormous quantity of solid wastes currently disposed off on land can be reduced by recovering materials and energy from wastes in a vastly environmental friendly manner (Annepu 2012). The scientific success, the deep scientific motivation, and acuity in solid waste management are the cornerstones of this vastly visionary scientific research pursuit. Morrissey et al. (Morrissey and Browne 2004) discussed with lucid details waste management models and their application to sustainable waste management. The vision of this paper is to review the types of models that are currently in use in the area of municipal waste management and also highlighted some major shortcomings of these models (Morrissey and Browne 2004). Decision support systems and its vast and varied applications are the veritable pillars of this well-researched paper. Technology and engineering science of mathematical modeling and waste management is highly advanced today as science and engineering moves toward a newer eon. A model is veritably a representation of an object, system, or idea in some form, other than that of reality itself. Scientific forbearance and technological farsightedness are the cornerstones of this research endeavor (Morrissey and Browne 2004). Many of the models identified are decision support models, using a variety of methods and techniques, such as risk assessment, environmental impact assessment, and the vast domain of life cycle analysis (Morrissey and Browne 2004). Life cycle assessment and its applications are the deep visionary avenues of this endeavor (Morrissey and Browne 2004). The authors rigorously point toward the scientific success, the vast scientific profundity, and the immense challenges behind mathematical modeling and waste management (Morrissey and Browne 2004). Seadon (2010) deeply comprehended on the vast scientific vision and scientific cognizance of sustainable waste management systems. Human scientific research pursuit, the futuristic vision of engineering science, and the technological fortitude will all lead a long and visionary way in the true emancipation and true realization of sustainability and waste management (Seadon 2010). Waste management is widely emancipated as part of a generation, collection, and disposal system. Sustainable practice and vast environmental sustainability are the veritable pillars of this treatise. According to this treatise, the forays and moves toward a sustainable society start

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with the success of managing waste. The challenge and the vast vision of sustainability science are slowly gearing forward toward a newer vision and newer innovations. The pillars of science and technology are today far-reaching, and solid waste management is an enjoinder toward its future. Sustainable solid waste management practically opens up a new eon in environmental engineering science (Seadon 2010). Human mankind’s scientific prowess and deep scientific girth are the needs of environmental sustainability and waste management (Seadon 2010). The salient features of sustainable development are the cornerstones of this research pursuit (Seadon 2010). Agamuthu (Agamuthu et al. 2009) discussed with lucid details drivers of sustainable waste management in Asia. South Asia and Southeast Asia are in the threshold of newer scientific regeneration and deep rejuvenation of technology (Agamuthu et al. 2009). In this review, four groups of drivers of sustainable management, specifically in Asia, are investigated. The four groups of drivers consist of three human elements (human, economic, and institutional) and the environment as a single entity (Agamuthu et al. 2009).

Recent Scientific Research Pursuit in Biomedical Waste Management Scientific advancement in the field of waste management and environmental pollution is reaching immense heights as scientific and academic rigor progresses forward. The world of science and technology needs to be re-envisioned and re-envisaged with the passage of scientific history and visionary timeframe. Sustainable biomedical waste management is an immensely large area of scientific endeavor. Human scientific rigor and research pursuit in solid waste management are the forerunners toward a greater visionary emancipation of environmental pollution control today. In this treatise, the authors repeatedly point out toward the scientific vision and the scientific forbearance toward the application areas of solid waste management. Sarsour et al. (2014) delineated with lucid details in a pilot study assessment of medical waste management within selected hospitals in a developing country. The present study aims to provide information about the effective management, segregation, storage, and disposal of medical wastes in public and private hospitals in Gaza Strip Palestine (Sarsour et al. 2014). This study is an eye-opener toward the greater scientific vision in the techniques applied in biomedical waste management in the developing world. A cross-sectional study was employed, and simple random sample techniques were used to distribute a semistructured questionnaire among 164 health workers at two hospitals in Gaza Strip with 100 respondents from government hospital and 64 respondents from private hospital (Sarsour et al. 2014). The results disclosed that healthcare facilities whether private or government hospitals still suffer from disorganized and inappropriate biomedical waste management which have veritably not received sufficient concern according to 60% of participants who pointed out that hazardous and medical wastes are still handled and

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disposed together with domestic wastes and segregation was applied only for sharp waste which is collected in special sharp boxes at the beginning after usage (Sarsour et al. 2014). Medical and healthcare wastes have sharply increased in recent decades due to the increased population, number and size of healthcare facilities, as well as the use of disposal medical facilities. The vision and the challenge of science in biomedical waste disposal are surpassing visionary scientific frontiers today (Sarsour et al. 2014). According to the US Environmental Protection Agency, medical wastes contain all waste materials generated by healthcare facilities, such as hospitals, clinics, physician’s offices, dental practices, blood banks, and veterinary hospitals/ clinics, as well as at medical research facilities and laboratories (Sarsour et al. 2014). Health care and medical science are entering into a newer scientific paradigm and an era of newer scientific innovation. The authors in this treatise discussed lucidly and gave deep insights into assessment of medical waste management in Gaza Strip Palestine, as well as to assess if there are any control measures for healthcare safety and medical waste disposal (Sarsour et al. 2014). Human civilization and human scientific progress are in the threshold of newer scientific regeneration and vast scientific advancements. The authors also discussed background information about the study of respondents, training programs about healthcare waste management, segregation and container issues, medical waste storage issues, medical waste transportation issues, and the deep scientific realization of solid waste management (Sarsour et al. 2014). Arora (2013) deeply discussed with cogent insight the management and handling of hospital wastes. Scientific vision and the challenge and fortitude of science are reaching glorious heights as environmental pollution control and solid waste management enter into a newer visionary eon. Medical wastes have been deeply identified by US Environmental Agency as the third largest known source of dioxin air emission and contributor of about 10% of mercury emissions to the environment from human activities (Arora 2013). This treatise will urge the reader about the impacts of improper waste management (Arora 2013). The challenge and the targets of science and engineering are today leading a long and visionary way in the true realization of environmental sustainability and ecological biodiversity (Arora 2013). Environmental pollution control and waste management are today in the threshold of a new era of scientific vision and scientific regeneration (Arora 2013). Today, the main bottleneck of sound hospital waste management is the lack of proper training and appropriate skills, insufficient resource allocation, and lack of adequate equipment (Arora 2013). This treatise gives a broad idea of hospital waste management practices and will veritably equip the readers with enough skills for effectively managing hospital waste, safeguarding themselves and the community against adverse health issues (Arora 2013). The vast rules for management and handling of hospital wastes are also summarized, giving the categories of different wastes, suggested storage containers including color coding, and treatment options. Scientific discernment, scientific cognizance, and scientific thoughts are the hallmarks of this well-researched treatise (Arora 2013). Hirani et al. (2014) discussed with cogent insight the management of biomedical wastes. Technological vision, scientific profundity, and the vast challenges are the

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scientific drivers of the greater realization of waste management and sustainability today (Hirani et al. 2014). Sustainable development whether it is energy or environment today stands tall with scientific might and vision as engineering and science moves forward. Biomedical wastes are those potential hazardous waste materials, comprising of solids, liquids, and laboratory wastes which pose a grave danger to the human health as well as in other living organisms (Hirani et al. 2014). The targets and vision of science are challenging the human landscape and ushering in a new era in vision and fortitude. The authors in this treatise present with immense vision the scientific subtleties and the deep scientific vision behind biomedical engineering applications and waste management principles. The world of challenges, the visionary greatness of biomedical science, and the futuristic vision will all lead a long and effective way in the true emancipation of environmental pollution control and environmental sustainability today. Sustainable development according to United Nations is ushering in a new era in the field of engineering and science. Approximately 75–90% of the biomedical waste is nonhazardous and as harmless as any other municipal wastes (Hirani et al. 2014). The rest 10–25%, though mixed with nonhazardous wastes, can be injurious to humans and or animals and veritably deleterious to environment and human health. Human scientific vision and human scientific forbearance are today in the state of immense scientific regeneration. Sustainable biomedical waste management is the veritable cornerstones toward a visionary era in environmental engineering science today (Hirani et al. 2014). Sharma (2010) deeply discussed with vast foresight awareness biomedical waste management among healthcare personnel of some important medical institutions in Agra, India. The vision and the challenge of science are far-reaching and surpassing vast scientific frontiers (Sharma 2010). Technological advancements in biomedical engineering are witnessing immense challenges and scientific fortitude. The proper handling of biomedical wastes is immensely imperative as science and technology marches forward. There are fixed set of rules and regulations for handling biomedical wastes throughout the world. Technological and scientific challenges are immense and groundbreaking today as medical science and engineering moves toward a newer arena of scientific regeneration and scientific revamping. Biomedical waste removal and effective disposal are of utmost importance as medical science and bioengineering surpasses visionary boundaries. Human civilization and human scientific research pursuit today stand in the midst of high scientific hopes and scientific determination (Sharma 2010). Yelebe et al. (2015) in a detailed review deeply discuss biomedical waste treatment in a case study of some selected hospitals in Bayelsa State, Southern Nigeria. Technological vision and scientific motivation are the cornerstones of this vast scientific research pursuit. Human scientific endeavor is deeply challenged and today stands in the midst of deep introspection (Yelebe et al. 2015). The treatment and disposal of solid medical waste from hospitals in Nigeria have been of growing concern and immense caution in recent times. This is due to the hazardous nature of the wastes and the potential threat to spread deadly diseases to human beings and other living organisms (Yelebe et al. 2015). To characterize and quantify these wastes, a study was carried out to ascertain the generation of biomedical wastes

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from ten hospitals in Bayelsa State, Southern Nigeria. Technology and engineering science of biomedical engineering and biomedical science are surpassing vast and versatile scientific frontiers. Medical science today is in a state of immense scientific regeneration. In this study, the hospitals were characterized into tertiary, secondary, and primary health institutions and grouped into public- and private-owned facility. Scientific evolution, scientific temperament, and deep scientific candor are today in the path of newer regeneration. This treatise widely presents the scientific success, the scientific vision, and the challenges behind biomedical waste removal and veritable disposal with a strong vision toward the emancipation and true realization of science and engineering. Human mankind’s immense scientific prowess, the deep scientific revelation, and the futuristic vision of engineering science will all lead a long and visionary way in the opening of a new era in solid waste management. Solid waste management and the visionary aisles of technological validation are changing the scientific landscape. While waste management has become a critical issue which has taken a central place in the national health policies of developed nations and is attracting considerable interest, in most developing countries like Nigeria, the handling and treatment of municipal solid wastes or domestic waste have not received sufficient attention (Yelebe et al. 2015). In most developing countries like Nigeria, the management, treatment, and the handling of medical waste are often in a poor condition as medical wastes are still disposed together with municipal solid waste in landfills and/or open municipal dumpsites at various locations within cities. The vision and the challenge of science are today opening new vistas in the areas of solid waste management and environmental pollution control. Environmental sustainability is another area of vast scientific imperative (Yelebe et al. 2015). Arshad et al. (2011) reviewed lucidly hospital waste disposal. Human civilization and human scientific endeavor are today in the path of vast scientific regeneration and scientific forbearance. This is a review paper targeting the domain of survey of hospitals and research studies (Arshad et al. 2011). Hospital waste management in the world is a formal and vast discipline and does veritably occupy a critical place in the management of the evergrowing healthcare sector. The management of hospital waste veritably requires its removal and disposal from the healthcare establishments as hygienically and economically as possible by methods that all decreases the risk to public health and human environment (Arshad et al. 2011). Human scientific endeavor and vast research pursuit in medical science today are opening new doors of scientific instinct and scientific innovation in decades to come (Arshad et al. 2011). Technology and engineering science of biomedical engineering are today ushering in newer vision and newer scientific methodologies. To present and analyze the situation, analysis of medical waste management system was deeply performed to discern the various handling and disposal procedures, the vast knowledge awareness of individuals involved in waste generation, handling and disposal, and the potential impacts of waste stream on both human health and the natural environment. Technological vision and advancements are today in deep turmoil with the evergrowing concerns for environmental catastrophes and environmental sustainability. Technology and engineering science are in a state of immense distress (Arshad et al. 2011). The methods employed in the present study were literature

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review and survey method. The data collection was extensively done through questionnaire (data collection form), informal interviews, and site visits. It was found that a variety of methods were used by the medical facilities to dispose their wastes including burning burial, entombing, selling, dumping, and removal by municipal bins. The waste elimination practices were found to be very unsafe, and both clinical and nonclinical wastes were found to be thrown together. The vision of waste disposal is today found to be highly undeveloped, yet the science of biomedical engineering surpasses scientific imagination. Here arise the questions of environmental sustainability and deep vision. There was absolute insufficient awareness of the magnitude of the medical wastes issued by concerned individuals at different levels from director or divisional heads or ragpickers. The challenge and the vision of biomedical science go beyond scientific foresight and scientific imagination. This study opens up new avenues of scientific forbearance and vast scientific vision in proper waste management strategy which is veritably needed to ensure health and environmental safety. Technological enshrinement and scientific foresight today are in a state of immense revamping and vision. This challenge needs to be surpassed as human mankind moves forward.

Recent Scientific Endeavor in the Science of Sustainability Sustainability and sustainable development today are ushering in a new era in the field of science and technology. Human scientific endeavor and progress of academic and scientific rigor today stands in the midst of vision and deep scientific profundity. Scientific challenges, technological motivation, and the vast futuristic vision of environmental sustainability will all lead a long and visionary way toward the true emancipation of successful sustainable development today. In this section, the author pinpoints the global forays in science and engineering of sustainability. Technology and engineering science of biomedical engineering are today surpassing vast and versatile scientific frontiers. Sustainability science needs to be readdressed and re-envisioned with the passage of scientific history, scientific vision, and timeframe. Water science and water technology are of equal importance in the futuristic vision and future scientific research pursuit. Today true realization of sustainable development also involves water science, water technology, and environmental pollution control which encompass solid waste management. The vision of Dr. Gro Harlem Brundtland, former Prime Minister of Norway, needs to be re-envisioned and revisited. In this section, the author rigorously points out toward the interlinked domain of sustainability (environmental, energy, and infrastructural) and the vast domain of waste management. Human scientific forays and research validation of sustainability science are at deep stake. This paper gives vast and glorious insight into the success of energy and environmental sustainability to the scientific progress of human society. Human civilization and human society are today in the midst of an evergrowing sustainability crisis. The challenges and the vast vision are enormous and groundbreaking. In this vastly visionary treatise, the deep scientific prowess and the success

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of environmental engineering science are brought into the forefront toward the avid scientific reader and the scientific domain as a whole. Technological vision and scientific acuity are the pillars of this research work on sustainability and waste management. The challenge needs to be re-enlivened and re-justified as science and engineering of biomedical sciences and waste management surges forward. Canadian Institute of Actuaries Report (2015) vastly elucidated on the subject of climate change and resource sustainability. The report gave a vast overview for actuaries. Technology and engineering science of environmental and energy sustainability are today ushering in a new era and a newer vision. The topics of global warming and climate change are vastly researched and well presented in minute details in today’s chapter in human history (Canadian Institute of Actuaries Report 2015). Topics and research forays of increasing concentration of greenhouse gases, carbon dioxide emissions caused by fossil fuel resources, reduction of emissions of pollutant gases, renewable energy generation, and sustainability of earth’s resources are veritably changing the scientific landscape. Validation of science and technological vision is the utmost need of the hour today (Canadian Institute of Actuaries Report 2015). Climate change is widely supposed to influence the work done by actuaries. This vision and the challenge need to be readdressed and redefined with the progress of scientific history, scientific vision, and time. The vision of this paper is to provide some deep background on the science of climate change, its vast impacts, key ways to reduce the damage, and the roles that the actuarial profession can play in addressing the risks. Sustainability and actuary science are today linked by an unsevered umbilical cord. Actuaries are becoming more aware of the combined impact of climate change and limitations of resources – two vastly significant and separate issues – putting at risk the sustainability of the current socioeconomic situation of the entire planet (Canadian Institute of Actuaries Report 2015). Climate change is more vulnerable than global warming. The rise in average temperature is only one visionary indicator of vast changes also translating to extreme temperature, drought, flooding, storms, rising sea levels, food scarcities, and rise of infectious diseases (Canadian Institute of Actuaries Report 2015). Technology and science has few answers to this vital area of scientific concern. Mitigating resource scarcity entails adopting new approaches such as “circular economy” or restorative economy. The science of actuary thus opens up into new vistas of scientific evolution and innovation in decades to come. It aims to rely on renewable energy; favors recycling; minimizes, tracks, and hopefully eradicates the use of toxic wastes; and eliminates wastes through careful design. The questions toward effective implementation of environmental and energy sustainability are many and are veritably surpassing scientific boundaries. The potential impact on actuarial methods, especially future growth expectations, is pervasive in the work of actuaries and affects health and pensions areas, investment practices, and the broad domain of enterprise risk management (Canadian Institute of Actuaries Report 2015). Human mankind’s immense scientific prowess, the vast technological vision, and the deep scientific and academic rigor will all lead a long and visionary way in the true emancipation of environmental sustainability today (Canadian Institute of Actuaries Report 2015). Rusko et al. (Rusko and Prochazkova 2011) deeply discussed and comprehended solutions to the problems of the sustainable development management. This treatise

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shows that environment is one of the basic public assets of a human system and it must be veritably protected. According to the present knowledge, the sustainability is absolutely necessary for all human systems, and it is of immense necessity to invoke the sustainable development principles in all human system assets (Rusko and Prochazkova 2011). Human civilization and human scientific endeavor today stand in the midst of deep scientific comprehension and wide vision. In this treatise, the authors summarize the conditions for sustainable development; tools, methods, and techniques to solve the environmental issues; and the vast tasks of executive governance in the environmental domain. This treatise broadly explains the scientific success, the scientific revelation, and the deep scientific profundity behind true emancipation of sustainable development management (Rusko and Prochazkova 2011). Federal Ministry of Education and Research Germany Report (2009) discussed with immense lucidity research for sustainable development. The scientific status, the vast scientific vision, and the technological motivation are the forerunners toward a newer visionary era in sustainability science today. This treatise gives a wider glimpse on sustainability with specific importance on Germany (Federal Ministry of Education and Research Report, Germany 2009). Germany has an overwhelming and pioneering role in climate research and climate modeling, in the development and realization of strategies to adapt to climate change, and in climate services, not only in Europe but globally. Human scientific research pursuit today stands in the midst of deep comprehension and vast scientific introspection. German government has deeply with immense vision and farsightedness framed the concept of “Research for Sustainable Development.” The central fields of action are (1) global responsibility, (2) earth system and geotechnologies, (3) climate and energy, (4) sustainability and resources, and (5) social development. Scientific vision, scientific forbearance, and sustainable development are today interlinked. This report envisages the framework program “Research for Sustainable Development” (Federal Ministry of Education and Research Report, Germany 2009). The vision of sustainability today is truly opening up new knowledge dimensions and a new era in the field of environmental engineering and energy engineering. According to the report, three new funding initiatives are of immense necessities: (1) eye-level cooperation with third world countries, (2) research partnership with newly industrialized and developed nations in the field of climate protection, and (3) thorough understanding of the earth’s systems. The evergrowing loss of biodiversity, energy problems, scarce resources, and climate change as well as the social impacts are veritably interlinked at the global and local level. Thus the need of an integrated conceptual network on sustainability and sustainable development. Germany as one of the major developed nations and an innovative society assumes immense responsibility to the vast challenges of global sustainability. Key technologies which need attention today are materials sciences, nano- and biotechnology, environmental technologies, and the diverse domains of applied engineering sciences. This treatise opens up new scientific discernment and vast scientific acuity in the sustainability program of a highly developed country like Germany (Federal Ministry of Education and Research Report, Germany 2009). Wewerinke et al. (Wewerinke and Yu 2010) revisited and redefined a comprehensive treatise on climate change through sustainable development and the vast

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promotion of human rights. Human rights are a vital component toward successful implementation of sustainability today. Human scientific regeneration and human scientific rejuvenation are today witnessing turmoil and introspection as sustainable development gains immense importance. Climate poses immense risk to the human rights of millions of people such as rights to life, health, food, and water. Sustainability implies provision of basic human needs such as food, electricity, and water. This treatise addresses the success and vision of science with regard to sustainability. This paper sets out the immense relevance of human rights obligations in light of scientific and technological hindrances climate change poses to human development in developing countries (Wewerinke and Yu 2010). The success of human civilization and the success of vast scientific vision need to be re-envisioned and re-envisaged as human mankind moves forward (Wewerinke and Yu 2010). This paper demonstrates that the framework provided by United Nations Framework Convention for Climate Change (UNFCCC) is veritably suitable to facilitate rights-based cooperation in accordance with the principle of affirmative action and broad realization of human rights. The immense scientific fortitude, the vision, and the challenges of environmental and energy sustainability are opening new vistas of engineering research in decades to come (Wewerinke and Yu 2010). Kuhlman et al. (Kuhlman and Farrington 2010) redefined the concept of sustainability. Sustainability as a veritable policy concept has its origin in the Brundtland Report of 1987. The vision and the challenge of human civilization and human scientific endeavor are today in a state of immense comprehension and forbearance. Environmental and energy sustainability are the cornerstones of scientific research pursuit today. Human vision at the same time is highly challenged. The Brundtland Report was concerned with the tension between the aspirations of human civilization toward a better life and the vast limitations attached to it. In the course of scientific history and visionary timeframe, this concept encompasses and re-envisioned the three avenues: (1) economic, (2) social, and (3) environmental (Kuhlman and Farrington 2010). This treatise vociferously argues that the redefinition obscures the real contradiction between the aims of welfare for all and the vast domain of environmental conservation (Kuhlman and Farrington 2010). Here comes the necessity of an in-depth analysis. It is now proposed to deeply return to the previous definition of sustainability. The vast domain of sustainability thus is invariably and veritably linked with the scientific and academic rigor of human mankind (Kuhlman and Farrington 2010). United Nations Development Programme Report (United Nations Development Programme, Human Development Reports 2011) deeply comprehended with lucid details the vast vision of sustainability and inequality in human development. Science and technology in this age of sustainability needs to be re-envisioned and revamped with human scientific progress. This paper widely analyzes the theoretical and empirical links between inequality in human development in one hand and sustainability in the other. Technological and scientific validation are today in the midst of deep introspection and vision. This report specifically looks at the causality in both directions of inequality and sustainability (United Nations Development Programme, Human Development Reports 2011). Human knowledge dimensions need to be re-envisioned and reorganized as sustainability science gears toward a

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newer eon. The Human Development Research Paper Series is veritably a medium for sharing recent research commissioned to inform the global Human Development Report which aims at the furtherance of knowledge dimensions in sustainability science. Technology and engineering science of sustainability are today ushering in a new avenue in the field of human development, inequality, and sustainability (United Nations Development Programme, Human Development Reports 2011). This treatise widens the human thought process in these directions. Kates et al. (2005) discussed with deep and cogent insight the concept of sustainable development and its goals, indicators, values, and practice. Sustainability and the holistic domain of sustainable development are today in the process of newer scientific regeneration. The most widely accepted definition of sustainability encompasses the following concept. Humanity has the supreme ability to make development sustainable – to veritably ensure that it meets the needs of the present without compromising the ability of future generations to meet their own needs (Kates et al. 2005). This is the basic definition of sustainability. In this treatise, the author pointedly focuses on the concept, along with the differences and the common ground in definitions, goals, indicators, values, and practices of the vast domain of sustainability. In the last half of the twentieth century, as the world of science and technology progressed, four key themes emerged from collective concerns and deep aspirations of the citizens: world peace, freedom, development, and environment (Kates et al. 2005). Thus it arose the deepest concerns for sustainable development, and the United Nations World Commission on Environment and Development was initiated in 1987 (Kates et al. 2005). It resulted and ushered in a new era in the peace-loving nations of the earth. Some of the concrete challenges facing human civilization and sustainable development are at least as heterogeneous and complex as the diversity of human races throughout the world. Human scientific rejuvenation after 1987 paved a definite and visionary way in the true emancipation of environmental and energy sustainability. Turner (2014) vastly comprehended on the imminent global collapse and compared the limits of economic growth with historical data. Scientific vision, technological acuity, and the progress of academic rigor are today changing the scientific landscape of sustainability and moving toward a global reform. Sharpness of scientific research pursuit stands as a major imperative toward the progress of human civilization today. In this treatise, a deep scientific vision is envisaged and enlivened in the area of sustainable development and economic growth across nations (Turner 2014). The challenge and the immense vision of science and technology are today changing the global sustainable development scenario. The Limits to Growth is a well-researched treatise and veritably targets scientific sustainable development across nations. The vast concepts of sustainability have been over the years re-envisioned and restructured as regards the progress of economic growth and the progress of sustainability over the years (Turner 2014). As science progressed over the second half of the twentieth century, critical environmental issues, resource constraints, and the crisis of depletion of fossil fuel resources have urged the scientific domain to newer knowledge dimensions in science of sustainable development. The author in this treatise enlivens the ongoing debate and the ongoing tryst with scientific destiny between sustainable development and economic progress of a nation.

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DANIDA/Ministry of Foreign Affairs of Denmark Report (Ministry of Foreign Affairs, Denmark Report, DANIDA, International Development Corporation 2012) comprehended with deep insight and elucidated on the challenges, opportunities, policies, and partnerships in inclusive and sustainable development. Environmental engineering science and environmental sustainability are the utmost need of human development and economic growth of a nation today (Ministry of Foreign Affairs, Denmark Report, DANIDA, International Development Corporation 2012). The areas of research in this treatise target the challenges in sustainable development and the policy and institutional response to development challenges. The new developmental challenge, the long-term vision and objectives, and the futuristic vision of sustainability will lead a long and visionary way in the true realization and true emancipation of economic progress of a developing and developed nation today (Ministry of Foreign Affairs, Denmark Report, DANIDA, International Development Corporation 2012). Scientific discernment, scientific enlivening, and deep scientific vision are the necessities of environmental sustainability, environmental engineering science, and waste management. Sustainable waste management today stands tall as a major pillar toward scientific and economic progress of human society. Human development, human scientific rigor, and the futuristic vision of biomedical waste management are all the forerunners of environmental pollution control and the success of environmental sustainability. In this treatise, the author rigorously points toward the scientific success, the scientific provenance, and the immense scientific cognizance behind waste management and environmental sustainability.

Green Materials for Environment Green materials and green sustainability are today in the threshold of newer vision and newer scientific rejuvenation. Nanomaterials and green technology are the utmost need of the hour. Human scientific vision and human scientific endeavor are veritably challenged today and needs to be re-envisioned and revamped with the progress of nanotechnology and nanomaterials. In this treatise, the author repeatedly and rigorously focuses on the vast scientific success and the scientific regeneration in the field of green materials and green nanotechnology. Research and development initiatives in the field of green materials and green nanomaterials need to be vastly highlighted as science and engineering gears forward toward a newer eon (www. google.com; www.wikipedia.com).

Green Technology, Wastewater Treatment, and Solid Waste Management Green technology and solid waste management today are two opposite sides of the visionary coin. Wastewater treatment and solid waste management are the utmost need of the hour with the progress of scientific rigor. Solid waste management of

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medical domain is the vexing issues in the vast academic and scientific rigor of environmental pollution control. Industrial wastewater treatment today stands in the midst of deep scientific introspection and vast technological vision. The author in this treatise pointedly focuses on the success of science and engineering in waste management and industrial pollution control. Human scientific instinct in medical science is being highly challenged with the passage of deep scientific history and time (www.google.com; www.wikipedia.com).

Future Research Trends and Future Frontiers of Waste Management Research trends in waste management today are vast and versatile. The futuristic vision of waste management, the scientific prowess, and the technological profundity are the forerunners toward a newer era in biomedical science and engineering. Human scientific research pursuit in solid waste management and industrial wastewater treatment is the vision and the challenges of science today. Medical science is a huge colossus with a vast scientific vision and scientific profundity of its own. Technological advancements in medical science and biomedical engineering are ushering in a new era in science and engineering today. Research focus and research challenges should be targeted toward zero-discharge norms. Future frontiers of waste management need to be re-envisioned and redefined as science and technology moves forward. Technology and engineering science of biomedical engineering are highly advanced today. The vast scientific vision and the deep scientific prowess of present-day scientific civilization need to be restructured and re-envisaged with the passage of scientific history and time. Future frontiers of waste management are vast and versatile as science and technology moves forward. Research trends and scientific farsightedness should be targeted toward newer techniques in elimination of biomedical waste generation. Today, biomedical engineering stands in the midst of scientific comprehension and deep scientific introspection. In this paper, the author pointedly focuses on the vast research potential, the technological validation, and the scientific motivation in treating biomedical wastes. The scientific vision and the scientific forbearance of research forays are presented in minute details in this treatise (www.google.com; www.wikipedia.com).

The Challenge and the Vision of Sustainability and Solid Waste Management Solid waste management and industrial pollution today are in the path of newer scientific regeneration and deep technological profundity. The challenge and the wide vision of sustainability science need to be revamped and readdressed as human civilization and human scientific endeavor move from one visionary scientific paradigm toward another. The vast vision of Dr. Gro Harlem Brundtland, the former Prime Minister of Norway, on the subject of sustainable development and

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sustainability needs to be redefined and restructured as science and engineering gears toward a newer visionary realm. The challenges of solid waste management, industrial wastewater treatment, and drinking water treatment are immense and groundbreaking. Environmental and energy sustainability are the hallmarks of newer scientific innovation of tomorrow. Global water crisis and global water hiatus are changing the scientific mind-set of environmental engineers and chemical process engineers. The vast success, the targets, and the immense potential are changing the scientific paradigm of environmental engineering science and solid waste management. Solid waste management and biomedical engineering are the veritable challenges of today. Solid waste management and environmental engineering science are veritably linked by an unsevered umbilical cord.

Conclusion and Future Scientific Perspective Human civilization and human scientific endeavor today veritably stand in the midst of vast scientific vision and deep fortitude. In this treatise, the authors rigorously point out toward the recent research endeavor in the field of sustainable biomedical waste management. Future perspectives and futuristic vision of biomedical engineering and environmental engineering science need to be re-envisioned and revamped as science and engineering marches forward. Environmental concerns, the vast scientific prowess, and the scientific determination will all lead a long and visionary way in the true emancipation of biomedical engineering and waste management. Scientific research pursuit in engineering science is changing the face of profundity of science and technology today. This treatise rigorously pursues the scientific intricacies and scientific hindrances in the quest for effective solid waste management. Biomedical engineering and waste management need to be re-envisioned and revamped with immense vigor and emancipation. Environmental pollution control and environmental engineering science are the utmost need of the hour as science and engineering progresses with immense vigor and might. Human scientific regeneration and deep scientific vision are the forerunners toward a newer emancipation of waste management and the holistic domain of environmental sustainability. Solid waste management and successful sustainable development are the two opposite sides of the visionary coin today. Human mankind’s immense scientific prowess, scientific adjudication, and the vast scientific landscape will all lead a long and visionary way in the true realization of scientific genesis of sustainable waste management today.

References Agamuthu P, Khidzir KM, Hamid FS (2009) Drivers of sustainable waste management in Asia. Waste Manage Res 27:625–633 Annepu RK (2012) Sustainable waste management in India. Master of Science in Earth Resources and Engineering Thesis, Columbia University in the City of New York

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Arora M (2013) Hospital waste: management and handling. Int J Adv Res Technol 2(11):238–245 Arshad N, Nayyar S, Amin F, Mahmood KT (2011) Hospital waste disposal: a review article. J Pharm Sci Res 3(8):1412–1419 Canadian Institute of Actuaries Report (2015) Climate change and resource sustainability- An overview for actuaries, August, 2015, Ottawa, Canada Federal Ministry of Education and Research Report, Germany (2009) Research for sustainable development. Framework Programme of the German Federal Ministry of Education and Research, Berlin, Germany Hirani DP, Villaitramani KR, Kumbhar SJ (2014) Biomedical waste: an introduction to its management. Int J Innovative Res Adv Eng 1(8):82–87 Kates RW, Parris TM, Leiserowitz AA (2005) What is sustainable development? Goals, indicators, values and practice. Sci Policy Sustain Dev 47(3):8–21 Kuhlman T, Farrington J (2010) What is sustainability? Sustainability 2:3436–3448 Ministry of Foreign Affairs, Denmark, Report, DANIDA, International Development Corporation (2012) Inclusive and sustainable development: Challenges, opportunities, policies and partnerships, Overseas Development Institute and DANIDA, 2012, United Kingdom Morrissey AJ, Browne J (2004) Waste management models and their application to sustainable waste management. Waste Manag 24:297–308 Rusko M, Prochazkova D (2011) Solution to the problems of the sustainable development management. In: Research Papers, Faculty of Materials Science and Technology in Trnava, Slovak University of Technology in Bratislava, Slovakia. no 31, pp 77–84 Sarsour A, Ayoub A, Lubbad I, Omran A, Shahrour I (2014) Assessment of medical waste management within selected hospitals in Gaza strip Palestine: a pilot study. Int J Sci Res Environ Sci 2(5):164–173 Seadon JK (2010) Sustainable waste management systems. J Clean Prod 18:1639–1651 Sharma S (2010) Awareness about bio-medical waste management among health care personnel of some important medical centers in Agra. Int J Environ Sci Dev 1(3):251–255 Singh GK, Gupta K, Chaudhary S (2014) Solid waste management: its sources, collection, transportation and recycling. Int J Environ Sci Dev 5(4):347–351 Swedish Environmental Protection Agency Report (2005) A strategy for Sustainable Waste Management, Sweden’s Waste Plan, Stockholm, Sweden, September, 2005.(ISBN: 91-620-1249-5) Tiwari AV, Kadu PA (2013) Biomedical waste management practices in India-a review. Int J Curr Eng Technol 3(5):2030–2033 Turner G (2014) Is global collapse imminent? An updated comparison of the limits to growth with historical data. Melbourne Sustainable Society Institute, University of Melbourne, Research Paper No. 4 United Nations Development Programme, Human Development Programme, Human Development Reports (2011) Sustainability and equity: a better future for all, Palgrave Macmillan, New York, USA Wewerinka. M., and V.P. Yu III (2010) Addressing climate change through sustainable development and the promotion of human rights, South Centre, Geneva, Switzerland Wilson DC, Velis CA, Rodic L (2013) Integrated sustainable waste management in developing countries. Proc Inst Civil Eng Water Resour Manage 166(2):52–68 Yelebe ZR, Samuel RJ, Yelebe BZ (2015) Biomedical waste treatment: a case study of some selected hospitals in Bayelsa state, South-South, Nigeria. Am J Eng Res 4(6):160–164

Agrowaste Materials as Composites for Biomedical Engineering

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Geetanjali Kaushik, Poonam Singhal, and Arvind Chel

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Natural Fibers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Natural (Plant) Fibers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Natural Fibers as Biocomposites . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biomaterials . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Plant-Based Fibers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Nanoparticle-Reinforced Biocomposites . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cross-References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

Presently steel and plastic are widely used in the manufacture of various products such as doors, false ceilings, toys, boxes for agricultural use, rims, and mobile panels. However, it is evident that both these materials are neither economical nor eco-friendly, and their presence poses serious impacts for the user and the environment. Efforts are underway for research and development in agro-waste fibers, which have shown immense potential as alternative to conventional man-made materials. Agro-waste fibers such as bagasse, rice husk, coconut, banana, and sisal fibers hold significant potential as “Natural Green Composite” due to their high strength, environment-friendly nature, low cost, availability, and G. Kaushik (*) · A. Chel MGM’s Jawaharlal Nehru Engineering College, Mahatma Gandhi Mission, Aurangabad, Maharashtra, India P. Singhal Centre for Rural Development and Technology, Indian Institute of Technology Delhi, New Delhi, India Indian Institute of Management, Calcutta, India © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_154

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sustainability. It is also important to note that until now only 10% of natural fibers derived from Agro-wastes have found application as raw materials for several industries such as biocomposites, automotive component, biomedical, and others. It is concluded that these composites will save the cost involved in manufacturing, processing, and disposing to a significant extent as well as preserve the environment. Keywords

Agro-waste · Composites · Biomedical engineering

Introduction For the past few decades composite materials of high strength fibers, such as carbon, glass, and aramid, and low strength polymeric matrix are in focus in the aerospace, leisure, automotive, construction, and sporting industries due to their unique advantages of high strength to weight ratio, noncorrosive property, and high fracture toughness. However, recent times it is also seen that because of the serious drawbacks of these polymer-based materials, there is a shift in the research and engineering interest from the polymer and synthetic composites to natural or biocomposites. They are (i) nonrenewable, (ii) nonrecyclable, (iii) high energy consumption in the manufacturing process, (iv) health risk when inhaled, and (v) nonbiodegradable. Although glass fiber-reinforced composites have been widely used for many years to provide solutions to many structural problems as they are low cost and are of moderate strength, most of the Western countries are now concerned about the use of these materials as they induce serious environmental problems. Recently, because of strong emphasis on environmental awareness worldwide, it has brought much attention to the development of recyclable and environmentally friendly and sustainable composite materials. Environment supporting composites are biodegradable as they are broken down to smaller compounds by the action of living organisms leading to changes in physical properties. This is a desirable property involving microorganisms to degrade biomaterials (Cheung et al. 2009).

Natural Fibers Natural fibers have found their usage in composite applications over the past few decades (Aziz and Ansell 2004). Composite fibers have many advantages as compared to synthetic fibers in terms of their low tool wear, low density, cheaper cost, availability, and biodegradability (Nishino et al. 2003; Wambua et al. 2003). Most common plants used in for the production of natural fibers are bast fibers such as hemp, jute, flax, kenaf, and sisal (Kozlowski et al. 1999). One of the main reasons for such a growing interest in these fibers is that they have a higher specific strength than glass fiber and a similar specific modulus

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(Bledzki and Gassan 1999). These natural fibers therefore offer desirable specific strengths and modulus at a lower cost (Huda et al. 2006).

Natural (Plant) Fibers Natural fibers are further classified based on their origin, that is, whether they are derived from plants, animals, or minerals as shown in Fig. 1. According to studies, plant fibers are the most popular natural fibers used as reinforcement in the category of fiber-reinforced composites. Plant fibers are made from different parts of the plant such as bast (or stem, soft, or sclerenchyma) fibers, leaf or hard fibers, seed, fruit, wood, cereal straw, and other grass fibers. The chemical composition as well as the structure of plant fibers is illustrated in Table 1. Most plant fibers, except for cotton, are composed of cellulose, hemicelluloses, lignin, waxes, and several water-soluble compounds, and out of these constituents cellulose, hemicelluloses, and lignin are the major constituents and rest are present in less proportion. The chemical nature of these constituents also varies as cellulose is rigid and crystalline, lignin is in the form of microfibrilreinforced amorphous, and hemicelluloses are present in the form of a matrix. Cellulose is the main constituent of any plant fiber (Akil et al. 2011; Chawla 1998). Cellulose is the natural homopolymer or a polysaccharide with the same compound, where D-glucopyranose rings are connected to each other with β-(1,4)glycosidic linkages, as can be seen in Fig. 2. Cellulose relatively has high modulus and is the main fibril component of many naturally occurring composites, such as wood, and is often found with lignin. Cellulose is mainly composed of three elements, C, H, and O, with a general formula of C6H10O5, and is crystalline in nature (Chawla 1998). Lignin and other noncellulosic substances are also present in the cell walls, and their presence modifies the properties of the fiber. The noncellulose material is difficult to separate and is an expensive process and therefore cannot be removed completely from these fibers. Natural fibers Vegetable (cellulose or lignocellulose)

Seed Fruit Cotton Coir Kapok Milkweed

Bast Leaf Wood (or stem) (or hard) Flax Hemp Jute Ramie Kenaf

Pineapple (PALF) Abaca (Manila-hemp) Henequen Sisal

Animal (protein)

Stalk

Cane, grass & reed fibers

Wheat Maize Barley Rye Oat Rice

Bamboo Bagasse Esparto Sabei Phragmites Communis

Wool/hair Lamb’s wool Goat hair Angora wool Cashmere Yak Horsehair etc.

Fig. 1 Classification of natural fibers (Source: Akil et al. 2011)

Mineral fibers

Silk Tussah silk Mulberry silk

Asbestos Fibrous brucite Wollastonite

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Table 1 Chemical composition, moisture content, and microfibrillar angle of plant fibers

Fibers Flax Hemp Jute Kenaf Ramie Nettle Sisal Henequen PALF Banana Abaca Oil palm EFB Oil palm mesocarp Cotton Coir Cereal straw

Lignin Cellulose (%) Hemicellulose (%) (wt.%)

Pectin (wt.%)

71 70–74 61–71.5 45–57 68.6–76.2 86 66–78 77.6 70–82 63–64 56–63 65

2.3 0.9 0.2 3–5 1.9

18.6–20.6 17.9–22.4 13.6–20.4 21.5 13.1–16.7

2.2 3.7–5.7 12–13 8–13 0.6–0.7

10–14 4–8

10–14 10 13.1 5–12.7 5 12–131 1 19

10

60

Moisture content (wt.%)

Microfibrillar Waxes (%) angle ( )

8–12 6.2–12 12.5–13.7

1.7 0.8 0.5

5–10 2–6.2 8

7.5–17 11–17 10–22

0.3

7.5

2

10–22

11.8 10–12 5–10

14

42

11

85–90 32–43 38–45

5.7 0.15–0.25 15–31

46 0–1 3–4 8

40–45 12–20

7.85–8.5 8

– 30–49

0.6

Source: Bismarck et al. (2005) and Akil et al. (2011)

H

OH

CH2OH H

OH

H

H

H H

OH

H

O

OH

O

O

O CH2OH

CH2OH H

O H

OH

O

O

H

H

CH2OH

H

OH

n

Fig. 2 Chemical structure of cellulose (Akil et al. 2011)

An important property of cellulose is that it is hygroscopic in nature and is able to absorb moisture from the atmosphere in comparatively large quantities (Chawla 1998) because of which most polymeric fibers swell. This absorption leads to changes in weights and dimensions, as well as in strengths and stiffness. Some of the disadvantages with plant fibers are that they are exposed to biological decay as a result of which they darken and weaken with age on exposure to light. Plant fibers are not as durable as synthetic polymeric fibers. They are all easily attacked by a

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variety of microorganisms in a conducive environment like high humidity and temperature, leading to rot and mildew. Therefore, plant fibers are considered as renewable resources and they do not exacerbate problem of the CO2 emissions (Akil et al. 2011).

Properties and Characteristics of Natural (Plant) Fibers The essential properties and characteristics like density, diameter, and mechanical properties of both natural (plant) and synthetic fibers are summarized in Table 2 (Bismarck et al. 2005). Natural (plant) fibers are nonabrasive toward mixing and molding equipment, which can contribute to significant equipment maintenance cost reductions. They also present safer handling and working conditions compared to synthetic reinforcements, such as glass fibers. Their processing is environmental friendly, offering better working conditions and, therefore, a reduction in risk of dermal or respiratory problems. The most interesting aspect of natural (plant) fibers is their positive environmental impact. Natural (plant) fibers are renewable resources, where they are biodegradable and their production requires little energy (Mohanty et al. 2002). Major drawbacks of natural (plant) fibers compared to synthetic fibers are their nonuniformity, variety of dimensions, and their mechanical Table 2 Characteristic values for the density, diameter, and mechanical properties of (natural) plant and synthetic fibers Fibers Flax Hemp Jute Kenaf Ramie Nettle Sisal Henequen PALF Abaca Oil palm EFB Oil palm mesocarp Cotton Coir E-glass Kevlar Carbon

Density (g cm3) 1.5 1.47 1.3–1.49

Diameter (μm) 40–600 25–500 25–200

1.55 1.45

50–200 20–80

0.7–1.55

1.5–1.6 1.15–1.46 2.55 1.44 1.78

150–500

12–38 100–460 Pm (15) α α 1 LPðPm þET m Þα

Ch ¼ 1 if ET m  Pm Q(t) = xQ(t1) + 0.001(1  x) A SRm (16)

Notation Csr: actual runoff coefficient, Cwp: weighted potential runoff coefficient, RCD: regional consecutive dryness level, AImp: percentage of impervious surface per grid cell, Cper: runoff coefficient for permeable areas, CImp: runoff coefficient of the impervious area, n: Manning’s roughness coefficient, θW: volumetric soil water content, S: land surface slope in percentage, w1, w2, and w3: weights, Ch: coefficient representing soil moisture, α:parameter, Pm:rainfall [mm/month], Q(t): volumetric runoff[m3/month], x: delay factor, Q(t  1) volumetric runoff of previous month[m3/month]:, A:, AImp:, P24 : average daily rainfall in rainy days, ETm: potential evapotranspiration [mm/month], LP: parameter

(Csr) is an actual runoff coefficient and the second one (Ch) is a coefficient representing soil moisture conditions. SRm ¼ Csr Ch ðPm  I m Þ

(10)

To perform calculations, the model divides each grid cell into impervious and permeable surfaces fractions. Next, it will obtain a weighted potential runoff coefficient out of their combination. Then, the effect of the depth of rainfall will be taken into account to convert potential runoff to the actual runoff coefficient.

Evapotranspiration To compute grid cell evapotranspiration, WetSpass adds up actual evapotranspiration obtained from impervious surface (ETi), open water (ETO), bare soil (ETs), and the vegetated (ETV) fractions. Actual evapotranspiration is calculated using monthly potential evaporation and vegetation coefficients. To calculate the reference transpiration from the potential evapotranspiration (ETP), a vegetation coefficient is needed: γ 1þ Δ

c¼ γ rc 1þ 1þ Δ ra 

(17)

where γ is the psychrometric constant [kPa/ C], rc (bulk) surface resistance [s m1], and ra aerodynamic resistance [s m1]

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ra ¼



2 1 Za  Zd ln 0:168  U a Za Z0

1973

(18)

where the constant 0.168 is the square of von Karman coefficient (0.41), Ua [m/s] is the wind velocity at elevation Za[m] and Z0 is the aerodynamic roughness height [m], and Zd is zero displacement elevation [m]. Transpiration Trv is calculated as T rv ¼ c ET p

(19)

For vegetated regions where the root zone is above the ground water level, the modified Trv is:   T v ¼ 1  a1 w=T rv T rv

(20)

where a1 is a calibrated parameter and w is the available water for transpiration:   w ¼ Pm þ θfc  θpwp Rd

(21)

where θfcθpwp is the plant available water content per time step and Rd is the rooting depth. The total actual evapotranspiration (ETm [mm/month]) in this model is calculated by: ET m ¼ av ET V þ as ET s þ aO ET O þ ai ET i

(22)

Where aV is the vegetated area fraction; ETV is the evapotranspiration for vegetated area. The bare soil and its evapotranspiration are expressed as aS and ETS, respectively; the open water is referred to with aO and the evapotranspiration for that area is ETO; and ai is the impervious surface and ETi is the evapotranspiration from the impervious surface (Batelaan and De Smedt 2001, 2007).

Recharge and Base-Flow The primary purpose of developing the WetSpass methodology was to simulate long-term average spatial patterns of recharge (Batelaan and De Smedt 2001). In WetSpass-M, monthly recharge (Rm [mm/month]) is calculated as the residual term of the water balance: Rm ¼ Pm  SRm  ET m

(23)

The storage parameter (β) relates the previous base-flow (Qb(t1)) and current base-flow (Qb(t)): QbðtÞ ¼ βQbðt1Þ þ 0:001 N m ð1  βÞ∅Rm

(24)

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Where Nm is the number of days per month and ∅ [m2/day] is the contribution parameter to base-flow.

Applications WetSpass has been applied for numerous applications. For example, to analyze the relation between a number of hydrological processes such as distributed land use change for the Grote Nete basin, Belgium (Batelaan et al. 2002, 2003). Applications have also focused on past, present, and future climate scenarios (Mogheir and Ajjur 2013); recharge estimation (Batelaan and De Smedt 2007; Al Kuisi and El-Naga 2013; Melki et al. 2017; Mustafa et al. 2017); or other components of the water balance (Abu-Saleem et al. 2010; Gebreyohannes et al. 2013; Abdollahi et al. 2017).

References Abdollahi K (2015) Basin scale water balance modeling for variable hydrological regimes and temporal scales. PhD Dissertation, Department of Hydrology and Hydraulic Engineering, Vrije Universiteit Brussel, Brussels, p 173 Abdollahi K, Bashir I, Verbeiren B, Harouna MR, Van Griensven A, Huysmans M, Batelaan O (2017) A distributed monthly water balance model: formulation and application on Black Volta Basin. Environ Earth Sci 76(5):198 Abu-Saleem A, Al-Zubi Y, Rimawi O, Al-Zubi J, Alouran N (2010) Estimation of water balance components in the Hasa basin with GIS based WetSpass model. Agron J 9(3):119–125 Al Kuisi M, El-Naqa A (2013) GIS based spatial groundwater recharge estimation in the Jafr basin, Jordan–application of WetSpass models for arid regions. Rev Mex Cienc Geol 30(1):96–109 Arnold JG, Muttiah RS, Srinivasan R, Allen PM (2000) Regional estimation of base-flow and groundwater recharge in the Upper Mississippi river basin. J Hydrol 227(1):21–40 Ampe EM, Vanhamel I, Salvadore E, Dams J, Bashir I, Demarchi L, Batelaan O (2012) Impact of urban land-cover classification on groundwater recharge uncertainty. IEEE J Sel Topics Appl Earth Observ Remote Sens 99:1–9 Asheesh M (2007) Allocating gaps of shared water resources (scarcity index): case study on Palestine-Israel. In: Water resources in the Middle East. Springer, Berlin/Heidelberg, pp 241–248 Balek J (1989) Analysis and synthesis of the water balance components. In: Groundwater resources 281 assessment. Developments in water science, vol 38. Elsevier, pp 61–89 Batelaan O, Smedt FD (2001) WetSpass: a flexible, GIS based, distributed recharge methodology for regional groundwater modeling. In: Proceedings of a symposium held during the Sixth IAHS Scientific Assembly at Maastricht, July 2001, IAHS Publication 269, pp 11–17 Batelaan O, De Smedt F (2007) GIS-based recharge estimation by coupling surface–subsurface water balances. J Hydrol 337(3):337–355 Batelaan O, Wang Zhong-Min, De Smedt F (1996) An adaptive GIS toolbox for hydrological modelling. In: Kovar K, Nachtnebel HP (eds) Application of geographic information systems in hydrology and water resources management. IAHS publication, vol 235. IAHS Press, Wallingford, pp 3–9 Batelaan O, De Smedt F, Triest L (2002) A methodology for mapping regional groundwater discharge dependent ecosystems. In: Schmitz GH (ed) Proceedings of third international

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conference on water resources and environment research, Vol II. Dresden, pp 311–315, 22–25 July Batelaan O, De Smedt F, Triest L (2003) Regional groundwater discharge: phreatophyte mapping, groundwater modelling and impact analysis of land-use change. J Hydrol 275(1–2):86–108 Blöschl G, Ardoin-Bardin S, Bonell M, Dorninger M, Goodrich D, Gutknecht D, Matamoros D, Merz B, Shand P, Szolgay J (2007) At what scales do climate variability and land cover change impact on flooding and low flows? Hydrol Process 21(9):1241–1247 Botter G, Basso S, Rodriguez-Iturbe I, Rinaldo A (2013) Resilience of river flow regimes. Proc Natl Acad Sci 110(32):12925–12930 Chow VT, Maidment DR, Mays LW (1988) Applied hydrology. McGraw-Hill, 570 pp De Groen MM (2002) Modelling interception and transpiration at monthly time steps introducing daily variability through Markov chains. PhD dissertation, IHE-Delft, Swets and Zeitlinger, Lisse, 211 pp De Groen MM, Savenije HH (2006) A monthly interception equation based on the statistical characteristics of daily rainfall. Water Resour Res 42(12):W12417. https://doi.org/10.1029/ 2006WR005013 Gebreyohannes T, De Smedt F, Walraevens K, Gebresilassie S, Hussien A, Hagos M, Gebrehiwot K (2013) Application of a spatially distributed water balance model for assessing surface water and groundwater resources in the Geba basin, Tigray, Ethiopia. J Hydrol 499:110–123 Hoogeveen J, Faures JM, Peiser L, Van de Giesen NC, Burke J (2015) GlobWat–a global water balance model to assess water use in irrigated agriculture (discussion paper). Hydrol Earth Syst Sci Discuss 12:2015 Kneis D (2015) A lightweight framework for rapid development of object-based hydrological model engines. Environ Model Softw 68:110–121 Korzoun VI, Sokolov AA (1978) World water balance and water resources of the earth. Water Development, Supply and Management, United Nations Educational, Scientific and Cultural Organization, 75 – Paris (France). International Hydrological Decade, Moscow (USSR). USSR National Committee Luu TNM, Garnier J, Billen G, Orange D, Némery J, Le TPQ, . . . Le LA (2010) Hydrological regime and water budget of the Red River Delta (Northern Vietnam). J Asian Earth Sci 37(3):219–228 Lytle DA, Poff NL (2004) Adaptation to natural flow regimes. Trends Ecol Evol 19(2):94–100 Manfreda S, Fiorentino M, Iacobellis V (2005) DREAM: a distributed model for runoff, evapotranspiration, and antecedent soil moisture simulation. Adv Geosci 2(2):31–39 Melki A, Abdollahi K, Fatahi R, Abida H (2017) Groundwater recharge estimation under semi arid climate: case of Northern Gafsa watershed, Tunisia. J Afr Earth Sci 132:37–46 Merritt DM, Wohl EE (2002) Processes governing hydrochory along rivers: hydraulics, hydrology, and dispersal phenology. Ecol Appl 12(4):1071–1087 Merz SK (2006) Stocktake and analysis of Australia’s water accounting practice final report to Department of Agriculture, Fisheries and Forestry Middelkoop H, Daamen K, Gellens D, Grabs W, Kwadijk JC, Lang H, . . . Wilke K (2001) Impact of climate change on hydrological regimes and water resources management in the Rhine basin. Clim Chang 49(1):105–128 Mogheir Y, Ajjur S (2013) Effects of climate change on groundwater resources (Gaza strip case study). Int J Sustain Energy Environ 1:136–149 Murdoch PS, Shanley JB (2006) Flow-specific trends in river-water quality resulting from the effects of the clean air act in three mesoscale, forested river basins in the northeastern United States through 2002. Environ Monit Assess 120(1):1–25 Mustafa SMT, Abdollahi K, Verbeiren B, Huysmans M (2017) Identification of the influencing factors on groundwater drought and depletion in north-western Bangladesh. Hydrogeol J 25(5):1357–1375 Oki T, Kanae S (2006) Global hydrological cycles and world water resources. Science 313(5790):1068–1072

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Pengra B (2012) The drying of Iran’s Lake Urmia and its environmental consequences. UNEPGRID, Sioux Falls, UNEP Global Environmental Alert Service (GEAS) Porporato A, Daly E, Rodriguez-Iturbe I (2004) Soil water balance and ecosystem response to climate change. Am Nat 164(5):625–632 Shiklomanov IA (1998) World water resources: a new appraisal and assessment for the 21st century: a summary of the monograph World water resources. Unesco, Paris Smakhtin V (2004) Taking into account environmental water requirements in global-scale water resources assessments, vol 2. IWMI, Colombo Szilagyi J (2013) Recent updates of the calibration-free evapotranspiration mapping (CREMAP) method. In: Evapotranspiration-an overview. InTech, Rijeka Thompson SE, Harman CJ, Troch PA, Brooks PD, Sivapalan M (2011) Spatial scale dependence of ecohydrologically mediated water balance partitioning: a synthesis framework for catchment ecohydrology. Water Resour Res 47(10):W00J03 Wang X, Pullar D (2005) Describing dynamic modelling for landscapes with vector map algebra in GIS. Comput Geosci 31(8):956–967 Wang ZM, Batelaan O, De Smedt F (1996) A distributed model for water and energy transfer between soil, plants and atmosphere (WetSpa). Phys Chem Earth 21(3):189–193 Wisser D, Frolking S, Douglas EM, Fekete BM, Vörösmarty CJ, Schumann AH (2008) Global irrigation water demand: variability and uncertainties arising from agricultural and climate data sets. Geophys Res Lett 35(24):1–5 Zeinoddini M, Bakhtiari A, Ehteshami M (2015) Long-term impacts from damming and water level manipulation on flow and salinity regimes in Lake Urmia, Iran. Water Environ J 29(1):71–87

Modeling the Feasibility of Employing Solar Energy for Water Distillation

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Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Water Distillation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Solar Irradiation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Solar-Distillation Cost Analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Problem Formulation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . General Assumptions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Specific Assumptions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Mathematical Models and Equations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Results and Discussion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Effect of Solar Radiation on Water Temperature . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Effect of Radiation Time on Water Temperature . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Effect of Position (Horizontal-Axis, z) on Water Temperature . . . . . . . . . . . . . . . . . . . . . . . . . . . . Maximum Water Temperature . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Solar Radiation Versus Radiation Time . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

1978 1979 1984 1985 1985 1986 1987 1988 1993 1993 1994 1994 1995 1996 1998 1999

Abstract

The world’s demand on drinking water has been increasing enormously during the last few decades. Solar distillation may become a potential alternative for water treatment. The feasibility of using solar energy for water distillation was identified by modeling the water temperature profile in a rectangular stainlesssteel evaporator tank. Five different heat transfer scenarios were investigated to determine the maximum water temperature that could be achieved. Studied heat transfer modes included either conduction, convection, or radiation under steady or unsteady state conditions. A comparison between the five studied scenarios H. A. Maddah (*) Department of Chemical Engineering, King Abdulaziz University, Rabigh, Saudi Arabia e-mail: [email protected] © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_120

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showed that there was a noticeable rise in water temperature in all scenarios. The average temperature rise of the five scenarios was approximately 28  C and was equivalent to a final water temperature of 53  C which was kept initially at 25  C. Results confirmed that the increase in both radiation energy and radiation time would boost-up the water temperature quickly and more efficiently. Keywords

Water distillation · Solar energy · Radiation · Modeling · Mathematical models · Water treatment · Water purification · Heat transfer · Conduction · Convection · Radiation · Evaporator tank · Temperature · Heat · Energy · Treatment · Solar feasibility · Desalination · Sun energy · Radiation time · Radiation energy · Heat scenario · Transfer mode · Drinking water · Fresh water

Introduction Fresh water is the most vital element for human existence. Water makes up to 70% of human total body weight, and it is very necessary to increase life quality (Ogbonmwan 2011). However, natural fresh water resources and supplies, such as rainfalls and lakes, are limited in many countries. For example, there is a serious water problem in deserted countries like Saudi Arabia and Jordan (Middle East Area) in which the water demand is exceeding their water supplies due to the limited varieties of natural water resources (Muslih et al. 2010; Jaber and Mohsen 2001). Speaking of water demand world widely, 20% of the world’s current population lacks safe drinking water, and it is expected that by 2025, there will be around two billion people facing difficulties in finding fresh water and will live in a scarcity of water. The previous scenario will occur because of the rapid growth in world’s population, which consequently will increase the worldwide demand on fresh water (Maddah and Chogle 2017). Hence, there is a growing demand to improve the performance of current water purification technologies and to utilize renewable energies such as solar energy to compensate for the shortage in fresh water supply (Maddah and Chogle 2017; Maddah et al. 2017). Water treatment or purification process is basically a separation process known as desalination. In a desalination process, salts and other minerals (concentrated brine) are rejected from seawater or brackish water flow in order to have a freshwater stream containing a low concentration of dissolved salts. There is a specific amount of energy that is required by the desalination process in order to be able to separate the excess amounts of salts from saline water. Hence, researchers and scholars are constantly devoting so much time and effort to develop today’s available and commercial separation technologies. The common goal is to have an improved separation technology for water desalination with a lower water production cost. Water production costs can be minimized either by decreasing capital costs or by decreasing operation and maintenance (O&M) costs for such a water desalination plant, keeping in mind that energy costs are under O&M category since given energy is usually utilized for plant operation (Khawaji et al. 2008).

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The use of direct solar energy for desalinating seawater has been investigated quite extensively in the last few decades. The main reason behind using solar energy for water desalination is that it is free, available, renewable, environmental-friendly, and harmless. Solar desalination or solar distillation process mimics what happens to water in the natural hydrologic cycle: meaning that saline water (seawater or brackish water) is exposed to the sun’s rays and gets heated. Produced water vapor comes in contact with a cool surface that is placed within a solar still unit or a water condenser unit. As a result of the condensation process, fresh water droplets accumulate and get collected as a final product. Examples of solar distillation include the greenhouse solar still unit, the traditional boiler/condenser distillation system powered with solar energy, and the solar pond system (basin) (Hamed et al. 1993; El-Nashar 1993; Buros 2000). Solar stills may not be the optimal choice for solar water distillation because they have a high capital cost and they are vulnerable to weather-related damage. Moreover, variations in the overall efficiency between different solar still designs have been noted in previous studies. Different solar still designs share difficulties in the requirement of a large solar collection area (e.g., 25 hectares land/l000 m3 of product water/day); and thereafter reducing the feasibility of producing water (Khawaji et al. 2008). Evaluation of the whole desalination process in such a water distillation plant is important and should be considered carefully to check how the idea is feasible commercially. The following topics are some of the important research areas under the evaluation process of a desalination plant (Khawaji et al. 2008): • Improvements in the desalination process for salt rejection • Assessment of the composition of the effluents • Assessment of the environmental impacts of the effluents This chapter (study) will focus on evaluating a proposed solar-distillation system for small-scale and mid-scale production of potable water. The objective was to identify the feasibility of using solar energy for water distillation that is known as solar distillation system. In other words, the water temperature profile in a rectangular evaporator tank (assumed geometry) was modeled theoretically for different heat transfer scenarios in order to determine the maximum water temperature that could be reached from solar radiation; knowing the water temperature profile would help us visualizing the feasibility of solar energy for water distillation as well as determining the efficiency of the system.

Water Distillation Distillation appears to be one of the best practical and economical techniques for water treatment. Distillation technology is a great choice for governments and industries when a mass production of fresh water from high saline water like seawater is required to meet the demand (Saidur et al. 2011).

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The fundamental principle of any water distillation technique is based on the vaporization/condensation process. Vaporizing (boiling) saline water at 100  C (or even below) allows water vapors and dissolved gasses to become volatile and to be condensed in a later condensation stage. Minerals and dissolved salts in water are not capable to evaporate easily unless for boiling temperatures above 300  C. Thus, only pure water vapor (fresh water produced from condensation) will come out while salts and minerals will be left behind (Baker 2000; Johnson et al. 2012). There are several distillation methods, which were developed for water desalination. Simple traditional distillation method, as shown in Fig. 1, is the most wellknown, simple, and basic method to purify water. In this method, the saline water in the boiling flask is boiled by a heat source (electrical heater, natural gas stove, oil stove, solar plat, or direct solar energy) and then evaporated to go toward the condensing unit. The condensing unit is cooled to separate the coming water vapors as pure droplets; the water vapor loses some of its latent heat and changes from being in a vapor phase to be in a liquid phase, which will be collected in the collected flask to have the fresh water product. Traditional distillation method is commonly used in chemical laboratories as well as domestic applications (Harwood and Moody 1989; Cengel and Boles 2002). However, distilling water by the traditional method is very expensive in large-scale applications because traditional distillation method requires a large amount of energy to evaporate the saline water due to the high latent heat of vaporization of water that is about 2257 kJ/kg at 100  C (Cengel and Boles 2002). Another distillation method which differs in simplicity, cost, and applications is known as single stage distillation as shown in Fig. 2. The system consists of an evaporator tank, a heating steam coil (heat exchanger), a condenser tank, and a storage tank plus piping connections. Single stage distillation process is somehow similar to the traditional distillation technique. However, single stage distillation

Cooling water

Condensing unit Boiling flask Water vapour

Distilled water

Saline water Collecting flask Electrical heater

Fig. 1 Simple traditional distillation process (Adapted from Harwood and Moody 1989)

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Vapour Vent Evaporator tank Cooling water Saline water In

Condenser tank

Heating steam Storage tank Condensate to boiler

Fresh water

Rejected brine

Fig. 2 Single stage distillation process (Adapted from Spiegler 2012)

method is usually a continuous process where simple traditional distillation method is usually in batch mode. The evaporator tank in single stage distillation method follows the same concept as the boiling flask in simple traditional distillation method. The only difference is that energy (heat) source in single stage distillation is obtained from using heating steam coils that act like a heat exchanger between inlets and outlets of both evaporator and condenser tanks (Saidur et al. 2011; Johnson et al. 2012). This method is suitable for saline water distillation in marine application and large laboratories applications. Single stage distillation is the optimal choice for compact size plants and when it is possible to exchange energy for the heating steam coils with other processes of nearby plants and factories (Rahman et al. 2003). However, multiple effect distillation (MED) system consists of four evaporators, one boiler, and one condenser as shown in Fig. 3. Boiling water in the boiler will generate hot steam that will boil the saline water in the first evaporator. Hot vapors of the first evaporator will go to the second evaporator and will be the heating medium/ source to heat and boil the saline water there (Saidur et al. 2011; Johnson et al. 2012). In multiple stage flash (MSF) method that is shown in Fig. 4, saline water is heated outside the boiling chamber and then gets evaporated in the boiling chamber by lowering pressure. The increase in seawater temperature happens due to the latent heat of condensation flowing from the condensing water vapors. Providing an external low pressure steam, by using a steam turbine, allows the saline water to be heated in the brine heater (El-Nashar 2001). The heated water is then transported into the evaporator flash chambers (stages) which are usually made of multistages and typically contain 19–28 stages in modern large MSF plants (Jambi and Wie 1989).

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2nd effect

1st effect

Salt water feed

Vapor 90°

Vapor 95°

Pressure control valve

3rd effect

4th effect

Vapor 85°

Cooling water To steam ejectior and vacuum pump

Vapor 80°

Condenser Cooling water Steam 100C @1 atm

Boiler

Valve

Valve

Valve

Pump Pump

Pump

Pump

Pump

Waste brine Fresh water

Condensate return to boiler

Fig. 3 Multiple effect distillation process (Adapted from Spiegler 2012) Vent

60 C Heater

Salt water (20C)

Steam

Distillate Condensate to boiler 100 C

60 C

Brine (60C)

Fig. 4 1-stage flash distillation process (Adapted from; Khawaji et al. 2008)

Both MED and MSF methods require an external source of heat, like crude oil or natural gas that is used to heat up the incoming saline water. MSF process differs from MED by having both heating and boiling processes happening in the same vessel. In 2002, MSF has accounted for 36.5% of the total capacity of established desalination plants for both brackish water and seawater, and it reserved the first place among seawater desalination plants making more than 5000 m3/day of fresh water (Saidur et al. 2011; Porteous 1983). Solar distillation is an attractive way for water desalination due to the utilization of a free source of energy that is coming from sun in a form of heat (Saidur et al. 2011). The history of solar distillation started as early as in the fourth century B.C. where Aristotle explored a method to evaporate contaminated water and then condense it to

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produce fresh water. In 1589, Della Porta utilized the intense heat of solar rays, as shown in Fig. 5, to evaporate water placed in wide inverted earthen pots and then collected the condensate into vases placed underneath (Tiwari et al. 2003). In 1872, a solar water treatment plant was successfully built in Chile and ran for many years to produce fresh water. During World War II, small plastic solar stills were utilized to provide potable water for life rafts floating in the ocean (Saidur et al. 2011). Solar stills mimic the nature processes for generating rainfall, which are evaporation and condensation (Tiwari et al. 2003). Solar distillation process starts by feeding saline water into a black plate located in the lower portion of the solar distiller (solar still). Water vapor condenses on a cool transparent leaning surface which is usually made of glass or plastic. Pure water droplets are formed and grow up until they become heavy enough to slide down along the leaning surface. Fresh water droplets are collected through special channels located under the leaning surface to be transferred to the product side of the still and thereafter to reach the fresh water tank (Saidur et al. 2011). Conventional solar stills have various modifications and mode of operations. Hence, solar distillation systems are classified as passive and active solar desalination technologies. Active solar stills operate with an extra-thermal energy that is fed into the passive basin in order to achieve higher evaporation rate (Tiwari et al. 2003). In other words, passive solar stills are direct solar technology while active solar stills are indirect solar technology. Direct solar desalination (passive) requires large land areas and has a relatively low productivity. However, direct solar technology is preferred over the indirect one at a small-scale and mid-scale production due to its relatively low cost and simplicity. Furthermore, the integration of solar thermal energy as a direct technology is suitable for solar desalination applications in remote areas (Qiblawey and Banat 2008). In 2016, Moh’d et al. observed that it was feasible to increase the yield of produced fresh water by developing a hybrid solar-wind water distillation system which consisted of a conventional single basin solar-still Fig. 5 Historical solar distillation apparatus (Adapted from Tiwari et al. 2003)

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and a wind-water heater operating simultaneously. The advantage of the proposed hybrid system was to operate day and night and thereafter increased the quantities of distillate water by three to four folds which could reach up to 5.5 kg/day of distillate water (Moh’d et al. 2016).

Solar Irradiation The verification of having enough solar energy from sun’s irradiation is important to utilize solar energy in water distillation applications. Hickey et al. (1980) carried out an experimental work on the calculations of solar radiation measurements from the earth radiation budget on the Nimbus 7 satellite. The registered mean value of solar radiation was around 1376.0 W/m2. Another recent study in 2007 reported a solar irradiation mean value of 1368 W/m2 (Ammann et al. 2007). Observations of solar irradiance variability have been made in a previous study in 1981 by utilizing active cavity radiometer irradiance monitor on the Solar Maximum Mission satellite. High-precision measurements of total solar irradiance indicated an average value of 1368.31 W/m2. The irradiance variability magnitude and time scale proposed that there were considerable solar energy amounts within the convection zone in solar active regions (Willson et al. 1981). However, sustaining solar power at high levels is a challengeable task due to the presence of long-term variations in solar radiation at Earth’s surface. It has been reported that there was a linear decrease in the delivered solar power to earth from about the year 1960 to 1990. However, satellite records in the year from 1983 to 2001 revealed an overall increase in solar power at a rate of 0.16 W/m2 that was approximately 0.10% per year (Pinker et al. 2005). The increase in the delivered solar power might be associated with the increase in human-made greenhouse gases and aerosols, which resulted in developing Earth’s absorbance rate (nearly about 0.15 W/m2). In other words, our planet is absorbing more energy (accumulated) from the sun than it is emitting energy to the space and making a climate change problem due to the high emissions of carbon dioxide. It is worth mentioning that there is a difference between solar power delivery (required) and solar power accumulation (unwanted). Accumulation of energy in the Earth’s atmosphere leads to an increase in the overall Earth temperature, which is related to climate change. However, ensuring the delivery of solar power can be achieved with advanced solar technologies and without energy accumulation (Glaser 1968; Hansen et al. 2005). There are recent technologies that are efficient enough to convert solar power into usable energy in a cost-effective manner. One example is the parabolic (concave) mirrors that are capable of collecting the sun’s energy over a wide area and focus it onto a smaller area on the water surface to intensively heat up the water. The concentrated solar energy on a focus point can easily raise the water temperature above 600  C when there is a full sun and that is approximately translated as about 100 W/ft2 (Hameed et al. 2013). The use of concentrated solar power (CSP) technology, which produces high temperature by concentrating solar energy in a single focal point, along with a solar

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Modeling the Feasibility of Employing Solar Energy for Water Distillation

1985

tracking system to follow the direction of the sun throughout the day was proposed in a previous project by Hameed et al. (2013) in order to easily purify water at a small-scale design and at any remote area (Hameed et al. 2013).

Solar-Distillation Cost Analysis The water production cost, expressed in $/m3 of water, depends on both the capital cost of the equipment, the O&M cost including the energy, and the operation and maintenance cost other than energy. Water production cost is obtained by dividing the sum of all costs by the total produced water quantity. Today’s industry goal is to produce desalinated water at a low cost of ¢50/m3 of water and a low cost of power at ¢2/kW h; where ¢ denotes to a US cent (Khawaji et al. 2008). The O&M cost including the cost of energy is very small or even negligible for solar stills since the energy required for treatment is delivered by solar radiation and there is no need for operating pumps and controls in batch systems (Tiwari et al. 2003). Solar distillation systems are very attractive for small-size applications. The cost of fresh water productivity in solar distillation systems may vary from one place to another depending upon the intensity of solar radiation, the sunshine hours, and the type of still (Tanaka and Nakatake 2006). However, Howe and Tleimat (1974) reported that constructing water treatment plants with a capacity less than 200 m3/day is the most economical choice when using solar distillation. The development of a solar distillation system is feasibly suitable for water treatment when the weather conditions are favorable and the demand is not too large and less than 200 m3/day (Fath 1998). Small communities located in arid environments and with no fresh water sources in the region can save over 30% in total cost by using a solar distillation system rather than transporting water from long distances. The utilization of solar insolation during hot seasons is the most convenient choice in deserted areas to achieve the highest production (Akash et al. 2000). Kumar and Tiwari (2009) analyzed the life cycle cost of a single-slope passive and a hybrid photovoltaic (PV/T) active solar stills. The estimated distilled water production cost was Rs. 0.70/kg and Rs. 1.93/kg for passive and hybrid (PV/T) active solar stills, respectively, for 30 years lifetime of the systems. Kumar’s result is much economical than the cost of producing bottled water in Indian market that is around Rs. 10/kg. The payback periods of the designed passive and active systems were estimated to be in the range of 1.1–6.2 years and 3.3–23.9 years, respectively, with a selling price of distilled water in the range of Rs. 10/kg to Rs. 2/kg. The obtained energy analysis confirmed the energy payback time (EPBT) as 2.9 and 4.7 years, respectively (Kumar and Tiwari 2009; Sampathkumar et al. 2010).

Problem Formulation A rectangular evaporator tank was considered in the formulation of the heat transfer problem. Specific dimensions and design parameters of the water tank were assigned logically and they are as shown in Table 1. A complete flow chart of the proposed

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Table 1 Design parameters of the proposed solar-distillation system and their calculated values Parameter, (Symbol) Radiation energy, (R) Radiation time, (t) Tank volume, (V)

Unit W/m2 hr m3

Value 200~ 1000 2 ~ 10 61.25

Parameter, (Symbol) Water density, (ρ)

Unit kg/m3

Value 1000

Water heat capacity, (Cp) Convective heat transfer coefficient, (ho)a Conductive heat transfer coefficient, (k)b Initial temperature, (Ti) Convective-surrounding temperature, (T11)

J/kg K W/m2 K

4182 19.36

W/m K

28.5



25 30

Cross-section area, (Ac) Surface area, (As) Tank height, (H)

m2

20.4

m2 m

51 3.5

Tank width, (W)

m

3.5

As , (α) α ¼ ρVC p

m2 K/J

Tank length, (L)

m

5



C

Tank perimeter, (P)

m

17



C

27.5

z-axis differential width, (Δz) Centered z-axis width, (z)

m

3

Water-surrounding temperature (T12) Radiation-surrounding temperature (T13) m ¼ APc k, (m)

1.99  107 25

K/W

0.03

m

1.5

m1

0.57

ffi pffiffiffiffiffiffiffiffi pffiffiffiffiffiffiffiffi qffiffiffiffiffiffi ho m ¼ hAoc Pk , ( ho m)



C C

a

Thermal convective coefficients for air and water were averaged and determined from Kurganov 2011; then ho from Eq. (6) b Thermal conductive coefficient for stainless steel was averaged and determined at 20  C (Welty et al. 2009)

solar-distillation system is shown in Fig. 6. Since Stainless Steel (SS) material is cheap, abundant, and has a low thermal conductivity, it was selected as a construction material for the water evaporator tank. SS low thermal conductivity (k) will keep the absorbed radiation in the evaporator and may reduce the heat loss of sun’s thermal energy to the atmosphere (by conduction and through tank boundary-sides).

General Assumptions The solutions of the formulated problem were carried out based on various general assumptions. The below assumptions allowed us to make the heat transfer problem easier to approach, thereby solvable analytically. 1. Water evaporator tank is in a rectangular shape and heat balance is in Cartesian coordinates. 2. One-dimensional heat transfer is considered (z-axis only). 3. Average solar radiation energy (R) is equal 200–1000 W/m2. 4. No heat generation is involved in the modeling analysis.

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Modeling the Feasibility of Employing Solar Energy for Water Distillation

1987

Fig. 6 Solar-distillation system flow chart

5. Water salinity has no effect on the rate of heat transfer and the evaporation of water. 6. Water level in the tank is just as same as the tank depth (height). The formulated problem of the heat transfer was investigated in different five scenarios. A comparison between the five scenarios was established to find out by how much solar radiation can increase the water temperature. The five studied heat transfer scenarios are shown in Table 2 and named as the following: 1. 2. 3. 4. 5.

Radiation with convection Radiation with conduction Unsteady state radiation Unsteady state radiation with convection Radiation with both convection and conduction

Specific Assumptions Other specific assumptions were assigned for each studied scenario independently to obtain the rise in water temperature easily. As a general rule of thumb, the more we include modes of heat transfer in the equation, the more the result is accurate. However, solving partial differential equations (PDEs) in both space and time would be very difficult and should be solved by advanced mathematical methods or by the aid of engineering programs such as MATLAB. Table 3 demonstrates the specific assigned assumptions for each scenario for the formulated heat transfer problem of the water evaporator tank.

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Table 2 The five heat transfer scenarios of the formulated problem Scenario # 1 2 3 4 5

Radiation √ √ √ √ √

Convection √   √ √

Conduction  √   √

Heat generation     

Unsteady state   √ √ 

Table 3 Assigned specific assumptions of the different studied scenarios Scenario # 1 2 3 4

5

Assumptions The system is under steady state conditions. No heat transfer by conduction. The system is under steady state conditions. No heat transfer by convection. No heat transfer by conduction. No heat transfer by convection. Neglect heat transfer by conduction since its effect was shown to be low from scenario #2 results. This is because air conductivity to heat is generally very low. The system is under steady state conditions.

Mathematical Models and Equations Modeling analysis and formulated heat transfer equations were determined from the general equation of energy balance, Eq. (1). The overall convective heat transfer coefficient (ho) was calculated from Eq. (6); where h1 and h2 are 19.95 W/m2 K and 650 W/m2 K which are the convective heat transfer coefficients of air and water, respectively (Kurganov 2011). Water temperature equation of every scenario was obtained in Eq. (5), Eq. (16), Eq. (22), Eq. (31), and Eq. (37) for Scenarios 1, 2, 3, 4, and 5, respectively, Table 4. Starting from the general energy balance equation (Deen 1998): Ein þ Eg ¼ Eout þ Ea

(1)

Ein ¼ Eout

(2)

qrad ¼ qconv

(3)

R As ¼ ho As ðT w  T 11 Þ

(4)

Scenario #1

5

4

3

2

# 1

Unsteady state radiation with convection Radiation with both convection and conduction

Scenario Radiation with convection Radiation with conduction Unsteady state radiation

h i   T w ¼ T i eðho α tÞ þ h1o ðR þ ho T 11 Þ  1  eðho α tÞ n o o pffiffiffiffiffi pffiffiffiffiffi pffiffiffiffiffi n pffiffiffiffiffi pffiffiffiffiffi W ho mÞ W ho mÞ z ho mÞ W ho mÞ W ho mÞ Rþho ½T 11 T 13 Reð þho eð ½T 13 T 11 eð Rþho ½T 11 T 13 Reð þho eð ½T 13 T 11 R h pffiffiffiffiffi  T w ¼ ho þ T 11 þ pffiffiffiffiffi h pffiffiffiffiffi pffiffiffiffiffi i pffiffiffiffiffi i z ho mÞ W ho mÞ W ho mÞ W ho mÞ W ho mÞ ho eð eð eð ho eð eð

R As t þ Ti T w ¼ ρVC p

T w ¼ 2RkPAzc ðW  zÞ þ T 13

T w ¼ hRo þ T 11

Temperature equation for water

Table 4 Obtained water temperature profile equations

81 Modeling the Feasibility of Employing Solar Energy for Water Distillation 1989

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H. A. Maddah

Solving the equation gives: Tw ¼

R þ T 11 ho

(5)

Where: ho ¼ 1

1    =h 1 þ 1 =h 2

(6)

Scenario #2 Now, from Eq. (2) qrad þ qz ¼ qzþΔz   dT w  dT w  þ k A ¼0 R As  k Ac c dz z dz zþΔz

(7) (8)

Where: As ¼ P Δz V ¼ Ac Δz #

 dTdzw z d dT w d2 T w zþΔz ¼ ¼ dz dz Δz dz2

(9) (10)

"dT  w lim

dz

Δz!0

d2 T w RP ¼ 2 k Ac dz

(11)

(12)

Solving the ODE gives: Tw ¼ 

R P z2 þ C1 z þ C2 2 k Ac

(13)

Applying the below boundary conditions: BC1 : z ¼ 0; T ¼ T 13 BC2 : z ¼ W; T ¼ T 13

(14)

C2 ¼ T 13 RPW C1 ¼ 2 k Ac

(15)

We get constants:

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Modeling the Feasibility of Employing Solar Energy for Water Distillation

Tw ¼

RPz ðW  zÞ þ T 13 2 k Ac

1991

(16)

Scenario #3 Now, from Eq. (2) qrad ¼ ρVCp

dT w dt

(17)

R As dT w ¼ ρVCp dt

(18)

Solving the ODE gives: R As t þ C1 ρVCp

Tw ¼

(19)

Applying the below initial condition: IC : t ¼ 0; T ¼ T i

(20)

C1 ¼ T i

(21)

Tw ¼

R As t þ Ti ρVCp

(22)

Scenario #4 Now, from Eq. (1) Ein ¼ Eout þ Ea qrad ¼ qconv þ ρVCp

(23) dT w dt

R As ¼ ho As ðT w  T 11 Þ þ ρVCp

(24) dT w dt

dT w þ ho α T ¼ α ðR þ ho T 11 Þ dt

(25) (26)

Where: As ρVCp

(27)

1 ðR þ ho T 11 Þ þ C1 eðho α tÞ ho

(28)

α¼ Solving the ODE gives: Tw ¼

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H. A. Maddah

Applying the below initial condition: IC : t ¼ 0; T ¼ T i 1 ðR þ ho T 11 Þ ho

h i 1 ð h o α t Þ Tw ¼ Tie þ ðR þ ho T 11 Þ  1  eðho α tÞ ho C1 ¼ T i 

(29) (30) (31)

Scenario #5 Now, from Eq. (2) qrad þ qz ¼ qzþΔz z þ qconv   dT w  dT w  R A s  k Ac þ k A ¼ ho As ðT w  T 11 Þ c dz z dz zþΔz "dT  #

dT w  w d dT w d2 T w dz zþΔz  dz z lim ¼ ¼ Δz!0 dz dz Δz dz2 d 2 T w ho P RP ¼ ðT w  T 11 Þ  k Ac k Ac dz2

(32) (33)

(34)

(35)

Solving the ODE in MATLAB with the given below boundary conditions: BC1 : z ¼ 0; T ¼ T 13 BC2 ¼ z ¼ W; T ¼ T 13

(36)

We get the solution:

Tw ¼

n o pffiffiffiffiffiffi pffiffiffiffiffiffi R þ ho ½T 11  T 13  ReðW ho mÞ þ ho eðW ho mÞ ½T 13  T 11 R þ T 11 þ pffiffiffiffiffiffi h pffiffiffiffiffiffi pffiffiffiffiffiffi i ho ho eðz ho mÞ eðW ho mÞ  eðW ho mÞ o pffiffiffiffiffiffi n pffiffiffiffiffiffi pffiffiffiffiffiffi eðz ho mÞ R þ ho ½T 11  T 13  ReðW ho mÞ þ ho eðW ho mÞ ½T 13  T 11 h pffiffiffiffiffiffi  pffiffiffiffiffiffi i ho eðW ho mÞ  eðW ho mÞ

(37) Where: m¼

P Ac k

(38)

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Modeling the Feasibility of Employing Solar Energy for Water Distillation

1993

Results and Discussion Modeling results allowed us to determine the behavior of the heat transferred from sun radiation into saline water in the evaporator tank. Knowing the temperature profile inside the water tank is very important to estimate the highest achievable water temperature, which will allow us to check for the feasibility of employing solar energy for water distillation. The effect of solar radiation energy, solar radiation time, and horizontal position in the evaporator tank was studied through modeling analysis. Relations between the previous parameters and the rise in water temperature were considered and discussed thoroughly in the following subsections.

Effect of Solar Radiation on Water Temperature The effect of solar radiation energy was obvious from the calculated modeling results of the five scenarios. Figure 7 shows the effect of increasing solar radiation energy on water temperature in the five studied scenarios. Obviously, modeled results were determined as expected and that water temperature increased with the increase in solar radiation which indicated a proportional relationship between radiation energy and rise in temperature. Scenario #1 reserved the highest observed temperature of 81.7  C since it included only radiation and convection (heat

90 Scenario #1 Scenario #2

80

Water Tempteration (°C)

Scenario #3 Scenario #4

70

Scenario #5

60 50 40 30 20 0

150

300

450

600

750

900

1050

Solar Radiation (W/m2)

Fig. 7 Effect of solar radiation (R ¼200 ~1000 W/m2) on water temperature rise with a radiation time of (t = 10 h)

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H. A. Maddah

transferred into the system) and neglected any heat loss by other modes such as conduction. The addition of a heat loss term by conduction decreased the maximum observed water temperature to 45.9  C, which is illustrated by Scenario #5. However, including a heat loss by a conduction mean with neglecting convection (scenario #2) showed a slight decrease in the maximum modeled water temperature that was 71.3  C. A dependent time-radiation system (Scenario #3) and timeradiation-convection system (Scenario #4) showed comparable results that furtherly decreased the observed water temperature to 32.2  C and 32.3  C, respectively.

Effect of Radiation Time on Water Temperature The effect of solar radiation time was obvious from the calculated modeling results of Scenario #3 and Scenario #4 as shown in Fig. 8. Both scenarios showed analogous results, and a proportional relationship between radiation time and rise in water temperature was confirmed. A rise of water temperate of 5.74  C and 5.79  C was observed for Scenario #3 and Scenario #4, respectively. The unsteady state situations showed that time-dependent radiation would have lower temperature rise compared to steady state scenarios.

Effect of Position (Horizontal-Axis, z) on Water Temperature The effect of a selected horizontal position (width) within the evaporator system appeared when there was a conduction term, which accounted for a heat

Fig. 8 Effect of solar radiation time (t = 2 ~10 h) on water temperature rise with a radiation of ðR ¼ 1000 W/m2)

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Modeling the Feasibility of Employing Solar Energy for Water Distillation

1995

Fig. 9 Effect of horizontal position (z = 0 ~3.5 m) on water temperature rise with a radiation energy of (R ¼ 1000 W/m2) and a radiation time of (t=10 h)

loss from the evaporator tank. It was assumed that there was no effect from the vertical axis (water depth) on the water temperature profile. In other words, the depth of water (tank height) will not affect the distribution of absorbed radiation energy (heat) inside the water tank. It was also assumed that the water depth/level in the tank is just the same as the tank height. Results of Scenario #2 and Scenario #5 in Fig. 9 showed that the peak value of water temperature was observed at the center of the evaporator tank since the heat loss by conduction occurs at the boundary-sides (from the given boundary conditions). The highest modeled water temperatures were 72.24  C and 46.29  C for Scenario #2 and Scenario #5, respectively.

Maximum Water Temperature As noted previously, the maximum observed water temperature within the evap orator tank was 81.7 C, which was the result for the case of radiation with convection (Scenario #1) with the assumption of having a radiation energy and radiation time of 1000 W/m2 and 10 h, respectively. Table 5 shows the highest modeled water temperature values that were calculated in the different studied heat  transfer scenarios. The averaged-temperature value of 53 C was calculated from the five scenarios’ results, which gave us an approximation on how much would the water temperature rise in real situations. A comparison illustration between the obtained maximum water temperatures of the different studied heat scenarios is shown in Fig. 10.

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H. A. Maddah

Table 5 Maximum observed water temperature within the evaporator tank # 1 2 3 4 5 –

Scenario Radiation with convection Radiation with conduction Unsteady state radiation Unsteady state radiation with convection Radiation with both convection and conduction Averaged-temperature



Maximum Tw ( C) 81.7 72.2 32.2 32.3 46.3 53

Fig. 10 A comparison analysis of the maximum observed water temperature

Solar Radiation Versus Radiation Time Unsteady state scenarios were analyzed to determine whether solar radiation or radiation time is the much important contributing parameter in the rise of water temperature. Specifically, Scenario #3 results in Fig. 11 showed that the increase in solar radiation from 200 W/m2 to 1000 W/m2, under a radiation time of 10 h, caused a maximum rise in water temperature that was equivalent to 5.8  C. Conversely, for the same scenario (Scenario #3), the increase in radiation time from 2 h to 10 h, under a solar radiation of 1000 W/m2, caused a maximum rise in water temperature that was equivalent to 5.7  C. Since the difference between both temperature changes was comparable and was only 0.1  C less for the latter case, it was concluded that both radiation parameters (solar energy and time) have a similar contribution in raising water temperature. The same evaluation of the two parameters was carried out for Scenario #4 results that are shown in Fig. 12. A temperature of 5.3  C was the maximum rise in water temperature due to the increase in solar radiation from 200 W/m2 to 1000 W/m2,

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Modeling the Feasibility of Employing Solar Energy for Water Distillation

1997

Fig. 11 A comparison analysis between the effect of solar radiation and radiation time (Scenario #3)

Fig. 12 A comparison analysis between the effect of solar radiation and radiation time (Scenario #4)

under a radiation time of 10 h. However, a temperature of 5.8  C was the maximum rise in water temperature due to the increase in radiation time from 2 h to 10 h, under a solar radiation of 1000 W/m2. The inclusion of a convection mode in Scenario #4 showed opposite preference to the studied parameters, meaning that radiation time parameter was much important than solar radiation parameter in terms of increasing water temperature. Thus, one may infer that the two parameters may stand little

1998

H. A. Maddah

differently depending on the studied heat transfer scenario as well as the number of heat transfer modes included in the governing equation.

Conclusion The feasibility of using a proposed solar-distillation system was obtained from the general heat equation. Modeling analysis was applied to determine the water temperature profile in a rectangular stainless-steel evaporator tank. Five different scenarios were investigated theoretically to find out the maximum water temperature that could be achieved from solar radiation. The averaged-maximum-water-temperature of the five scenarios was equivalent to 53  C for water which was kept initially at 25  C. Results confirmed that there was a proportional relationship between radiation energy and rise in temperature as well as between radiation time and rise in water temperature. The increase in both radiation energy and radiation time would boost-up the water temperature. A comparison analysis between both radiation parameters concluded that solar energy and time have a similar contribution in raising water temperature. Future studies on solar-distillation technologies have to be continued to utilize most of the benefits of solar radiation and to ensure a clean water treatment technology for the coming generations. Nomenclature

E q Ac As ho h1 h2 T z k R P V C W H L ρ Cp t α m

Energy, W Energy flow (generally by conduction unless specified with a subscript), W Cross section area, m2 Surface area, m2 Overall convective heat transfer coefficient, W/m2 K Convective heat transfer coefficient for air, W/m2 K Convective heat transfer coefficient for water, W/m2 K Temperature,  C Z-axis dimension, m Conductive heat transfer coefficient, W/m K Solar radiation, W/m2 Perimeter, m Volume, m3 Constant Width, m Height, m Length, m Water density, kg/m3 Water heat capacity, J/kg K Time, s Defined variable in Table 1, m2 K/J Defined variable in Table 1, K/W

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Modeling the Feasibility of Employing Solar Energy for Water Distillation

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Subscripts

in g out rad conv a w 11 12 12 1 2 i

into the system generation out of the system radiation convection accumulation or storage water convective surrounding water surrounding radiation surrounding first second initial

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Jambi FI, Wie JM (1989) The royal commission gas turbine/HRSG/desalination cogeneration plant, 1989 ASME COGEN-TURBO. In: 3rd international symposium on turbomachinery, combinedcycle and cogeneration. American Society of Mechanical Engineers, New York, pp 275–280 Johnson JS, Dresner L, Kraus KA (1966) Principles of desalination. In: Spiegler KS, (Ed.), pp 371–383 Khawaji AD, Kutubkhanah IK, Wie JM (2008) Advances in seawater desalination technologies. Desalination 221(1–3):47–69 Kumar S, Tiwari GN (2009) Life cycle cost analysis of single slope hybrid (PV/T) active solar still. Appl Energy 86(10):1995–2004 Kurganov VA (2011) Heat transfer coefficient. Thermopedia publisher Maddah H, Chogle A (2017) Biofouling in reverse osmosis: phenomena, monitoring, controlling and remediation. Appl Water Sci 7(6):2637–2651 Maddah HA, Alzhrani AS, Almalki AM, Bassyouni M, Abdel-Aziz MH, Zoromba M, Shihon MA (2017) Determination of the treatment efficiency of different commercial membrane modules for the treatment of groundwater 8(6):2006–2012, Kurgnov: Thermopedia publisher Moh’d AAN, Kiwan SM, Talafha S (2016) Hybrid solar-wind water distillation system. Desalination 395:33–40 Muslih IM, Abdallah SM, Husain WA (2010) Cost comparative study for new water distillation techniques by solar energy using. Appl Sol Energy 46(1):8–12 Ogbonmwan SE (2011) Water for life Ireland campaign: supporting the united nation water campaign. In: The 2nd United Nation Water day Ireland Expo 2011 Trinity Science Gallery, Dublin, Ireland Pinker RT, Zhang B, Dutton EG (2005) Do satellites detect trends in surface solar radiation? Science 308(5723):850–854 Porteous A (1983) Desalination technology: developments and practice. Applied Science Publishers, London Qiblawey HM, Banat F (2008) Solar thermal desalination technologies. Desalination 220(1–3):633–644 Rahman H, Hawlader MNA, Malek A (2003) An experiment with a single-effect submerged vertical tube evaporator in multi-effect desalination. Desalination 156(1):91–100 Saidur R, Elcevvadi ET, Mekhilef S, Safari A, Mohammed HA (2011) An overview of different distillation methods for small scale applications. Renew Sust Energ Rev 15(9):4756–4764 Sampathkumar K, Arjunan TV, Pitchandi P, Senthilkumar P (2010) Active solar distillation – a detailed review. Renew Sust Energ Rev 14(6):1503–1526 Tanaka H, Nakatake Y (2006) Theoretical analysis of a basin type solar still with internal and external reflectors. Desalination 197(1–3):205–216 Tiwari GN, Singh HN, Tripathi R (2003) Present status of solar distillation. Sol Energy 75(5):367–373 Welty JR, Wicks CE, Rorrer G, Wilson RE (2009) Fundamentals of momentum, heat, and mass transfer. Wiley, Hoboken Willson RC, Gulkis S, Janssen M, Hudson HS, Chapman G (1981) Observations of solar irradiance variability. Science 211(4483):700–702

Applications of Soft Computing Methods in Environmental Engineering

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Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Description of Soft Computing Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Artificial Neural Networks (ANN) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Fuzzy Logic (FL) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Adaptive Neuro-Fuzzy Inference System (ANFIS) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Support Vector Machines (SVM) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Implementation of Soft Computing Methods in Environmental Engineering . . . . . . . . . . . . . . . . ANN-Based Applications for Water and Wastewater Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . ANN-Based Applications for Air Quality/Pollution Control/Forecasting . . . . . . . . . . . . . . . . . FL-Based Applications for Water and Wastewater Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . FL-Based Applications for Air Quality/Pollution Control/Forecasting . . . . . . . . . . . . . . . . . . . . ANFIS-Based Applications for Water and Wastewater Treatment . . . . . . . . . . . . . . . . . . . . . . . . ANFIS-Based Applications for Air Quality/Pollution Control/Forecasting . . . . . . . . . . . . . . . SVM-Based Applications for Water and Wastewater Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . SVM-Based Applications for Air Quality/Pollution Control/Forecasting . . . . . . . . . . . . . . . . . Illustrative Soft Computing Examples for Environmental Engineers . . . . . . . . . . . . . . . . . . . . . . . . . Example 1 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Solution of Example 1 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Example 2 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Solution of Example 2 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

2002 2004 2005 2010 2015 2018 2020 2020 2021 2023 2025 2026 2027 2027 2028 2030 2030 2032 2034 2036 2040 2040

K. Yetilmezsoy (*) Department of Environmental Engineering, Faculty of Civil Engineering, Yildiz Technical University, Istanbul, Esenler, Turkey e-mail: [email protected] © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_149

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Abstract

Soft computing has been extensively studied and applied in the last three decades for scientific research and engineering computing. In environmental engineering, researchers and engineers have successfully employed different methods of soft computing for modeling of various real-life environmental problems. In this study, applications of core soft computing techniques, such as artificial neural networks (ANN), fuzzy logic (FL), adaptive neuro fuzzy inference systems (ANFIS), and support vector machines (SVM), are investigated and important mathematical aspects of these methods are highlighted. Considering the concepts and methods, this study briefly reviews applications of soft computing techniques in the field of environmental engineering, especially in water/wastewater treatment and air quality/pollution control/forecasting. A brief introduction to complexity of environmental problems and the general definition soft computing concept are presented in the first section of this chapter. The second section comprises four subsections and presents mathematical background of four different soft computing methods. Section “Implementation of Soft Computing Methods in Environmental Engineering,” which is consisted of eight subsections, reviews successful applications of soft computing-based prediction models implemented in the field of environmental engineering and summarizes the important findings obtained in these studies. At the end of the overview of the published works on soft computing applications in different environmental areas, in the last section, some special illustrative soft computing examples and the ® respective MATLAB -based solutions are presented for environmental engineers. Keywords

Adaptive neuro fuzzy inference systems (ANFIS) · Aggregation · Air quality/ pollution control/forecasting · Algorithm · Artificial neural networks (ANN) · Backpropagation · Bayesian regulation · Broyden–Fletcher–Goldfarb–Shanno (BFGS) · Centroid · Classification · Defuzzification · Early stopping · Environmental engineering · Feed-forward · Firing strength · Fletcher–Reeves · Fuzzification · Fuzzy inference system (FIS) · Fuzzy logic (FL) · Fuzzy operator · Gradient descent · Hessian matrix · Implication · Jacobian matrix · Kernel functions · Lagrange multipliers · Levenberg–Marquardt · Linguistic · ® Logarithmic sigmoid · MATLAB · Membership function · Modeling · Momentum factor · Normalized layer (N) · Polak–Ribiére · Powell-Beale · Prediction · Product layer (π) · Quasi–Newton · Scaled conjugate gradient · Soft computing · Support vector machines (SVM) · Tangent sigmoid · Training · Water and wastewater treatment

Introduction The real-life environmental problems are very complex and highly dependent on several process configurations, different influent characteristics, and various operational conditions. For a sustainable control of environmental-related problems, the proposed systems must be continuously monitored and properly controlled due to

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possible instabilities in circumstance conditions. Although statistical models may be able to establish a relationship between the input and the output variables without detailing the causes and effects in the formation of pollutants, they are not capable of capturing the inherent nonlinear nature of the environmental problems. For this reason, the complicated inter-relationships among a number of system factors in the process may be explicated through a number of attempts in developing representative and powerful prediction models allowing the investigation of the key variables in greater detail. At this point, soft computing-based control of real-time process variables may provide several potential advantages, such as protection of the system from possible risks associated with significant fluctuations in influent characteristics, optimization of the process at a reasonable cost, providing a rapid evaluation and estimation of pollutant loads and emissions on energetic basis, and also development of a continuous early-warning strategy without requiring a complex formulation and laborious parameter estimation procedures (Yetilmezsoy et al. 2011a, b, 2015). The principal soft computing technologies can be categorized as fuzzy algorithms, neural networks, supporting vector machines, evolutionary communication, machine learning, and probabilistic reasoning (Jang and Topal 2014). McCulloch and Pitts (1943) introduced an initial model of an artificial neural network (ANN), which was recognized as the first study of artificial intelligence. It has been widely accepted as an approach, which acts like a “black- box” model derived from a simplified concept of the human brain, for prediction, control systems, classification, optimization, and decision-making in various fields (Antwi et al. 2017). In 1965, fuzzy logic (FL) theory was proposed by Zadeh (1965) as a new soft computing methodology in order to address uncertainty and subjectivity (i.e., human experience and intuition) within the framework of fuzzy sets which could be described by linguistic variables and membership functions according to a fuzzy rule-based system (Assimakopoulos et al. 2013). In 1993, soft computing became a formal area of computer science and many new and hybrid algorithms, i.e., adaptive neuro fuzzy inference systems (ANFIS) (Jang 1993), were introduced with the help of advanced computer technology (Jang and Topal 2014). The SVM method, developed by Vapnik (1995), can provide an effective novel approach to overcome the inherent drawbacks such as over-fitting training, local minima, and poor generalization performance of ANN when studying with large initial data. Since SVM implements Structural Risk Minimization Principle (SRMP), instead of the Empirical Risk Minimization Principle (ERMP) like feed-forward neural networks, its process leads to better generalization than conventional methods (Yeganeh et al. 2012). The main advantage of the SVM over multilayer perceptron (MLP) or neuro-fuzzy network is its good generalization ability, acquired at relatively small number of learning data and at large number of input nodes (high dimensional problem) (Osowski and Garanty 2007; Yeganeh et al. 2012). Among soft computing techniques, ANN provides configurations made up of interconnecting artificial neurons that mimic the properties of biological neurons. It is used in a wide range of applications as a multilayer feed-forward network with back propagation learning algorithm. A typical neural network includes three layers:

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first, the input layer, second, the output layer, and third, the hidden layer or intermediate layer (Gocic et al. 2015). Other alternative methodologies have also emerged from artificial intelligence, such as FL, which is currently being tested in real environmental problems. Its success is mainly due to its closeness to human perception and reasoning, as well as its intuitive handling and simplicity, which are important factors for handling of imprecise data (Kotti et al. 2013). This method develops multivalued, nonnumeric linguistic variables for modeling human reasoning in an imprecise environment. Nevertheless, it is noted that both ANN and FL control sometimes exist some shortages. For instance, ANN may have limitations in performing heuristic reasoning of the domain problem; on the other hand, the FL control may be difficult to design and adjust automatically. Moreover, the use of artificial neural networks is, however, challenged by the difficulty of network design and parameterization. Many factors affect the performance of ANN that include network topology, training algorithm and parameters setting, and network architecture. Likewise, the outcome of fuzzy classification highly depends on the predefined fuzzy rules (Dwarakish and Nithyapriya 2016). So ANFIS is designed as a fuzzy neural network model, it can use the both advantages. ANFIS consists of both ANN and FL including linguistic express of membership functions and if-then rules of Takagi and Sugeno’s type (Mingzhi et al. 2009). SVM is another novel soft learning algorithm that has been recently realized for a wide range of applications in the field of soft computing, hydrology, and environmental studies. It emerged as a set of supervised generalized linear classifiers and often provide higher classification accuracies than multilayer perceptron ANN. It is essentially a kernel-based procedure and relatively new machine learning method that has been recently applied as one of the leading techniques for pattern classification and function approximation (Gocic et al. 2015; Huang et al. 2010; Pai et al. 2011; Singh et al. 2011). This chapter is aimed at bringing forward original and the recent trends and efforts in the application of some soft computing methods in environmental engineering. It is especially interested in describing the successful application and advances in soft computing-based modeling of real-world environmental processes. The sections of this chapter summarize various applications of (1) artificial neural networks (ANN), (2) fuzzy logic (FL) control systems, (3) adaptive neurofuzzy inference systems (ANFIS), and (4) support vector machines (SVM) for modeling of various environmental problems based on water and wastewater treatment and air quality/pollution control/forecasting.

Description of Soft Computing Methods In this section, the basis of the widely used AI-based techniques, such as ANN, FL, ANFIS, and SVM, are briefly summarized and important mathematical aspects of these methods are highlighted. Moreover, computational issues, advantages, and particular theoretical principles are described, and some methodological techniques are discussed to make a comparative assessment of the present AI-based prediction models.

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Artificial Neural Networks (ANN) To better control a specific environmental process, a robust mathematical tool for predicting the process performance must be developed based on past observations of certain key parameters. Modeling a multivariate system is highly difficult due to the complexity of the environmental processes exhibiting nonlinear behavior that are difficult to describe by linear mathematical models (Hamed et al. 2004). Although deterministic models (also called white-box models) may provide insight into the mechanism, they require hard work before being applied to a specific environmental process. As an alternative to physical models, artificial neural networks (ANNs) are a valuable forecast tool in environmental sciences. They can be used effectively due to their learning capabilities and their low computational costs (Wieland et al. 2002). Because of their reliable, robust, and salient characteristics in capturing the nonlinear relationships between variables (multi-input/output) in multivariate systems, numerous applications of ANN-based models have been successfully utilized in the field of environmental engineering in the past decade (Yetilmezsoy and Demirel 2008). The ANN-based models are meant to interact with objects in the real world in the same way that the biological nervous system does. The calibration of ANN-based models is easier than the white-box models as fewer parameters are used in the model development process. For this reason, artificial intelligence techniques using ANN have recently become immensely popular and attractive mathematical tools for both modeling and controlling of several complex environmental processes. When the measured variables begin showing difference with the response of ANN, the model can be retrained using the newer data used for cross-checking. These facts and the quality of the results they provide make the ANN-based models more attractive than conventional models (Agirre-Basurko et al. 2006).

INPUT LAYER

HIDDEN LAYER OUTPUT LAYER

WH 1,1 X 1t

1

WH 1,HN

f (.) h

WO 1

WH 2,1 X 2t

o

WO HN

WH m,1 X mt

f (.)

WH 2,HN HN

f (.) WH m,HN b1

h

bo

b HN bias

bias

Fig. 1 Simple schematic of an ANN model (Adapted from Hamed et al. 2004)

Yo

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A simple diagram of an ANN model is depicted in Fig. 1. As seen in Fig. 1, each neuron is connected to several of its neighbors, with varying coefficients or weights representing the relative influence of the different neuron inputs to other neurons. The weighted sum of the inputs are transferred to the hidden neurons, where it is transformed using an activation function such as a tangent sigmoid activation function. In turn, the outputs of the hidden neurons act as inputs to the output neuron where they undergo another transformation. The output of a feed-forward ANN with one hidden layer and one output neural network is given as follows (Hamed et al. 2004; Antwi et al. 2017): Yo ¼ f o

" HN X j¼1

WOj  f h

m X

! WH ij  Xit þ bj

# þ bo

(1)

i¼1

where WHij is the weight of the link between the ith input and the jth hidden neuron, m is the number of input neurons, WOj is the weight of the link between the jth hidden neuron and the output neuron, fh is the hidden neuron activation function, fo is the output neuron activation function, bj is the bias of the jth hidden neurons, bo is the bias of the output neuron, Xit is the input variable, and HN is the number of hidden neurons. Hamed et al. (2004) reported that the tangent sigmoid (tansig) activation functions for the input and hidden neurons are needed to introduce nonlinearity into the network in order to make nets more powerful than plain perceptrons. Moreover, the authors reported that a linear activation function, such as linear transfer function (purelin), could be selected for the output neuron since it is appropriate for continuous valued targets. The logarithmic sigmoid function logsig(x) produces outputs between 0 and 1 as the node’s net input goes from negative to positive infinity. Alternatively, the tansig (x) as transfer function can be used. Sigmoid outputs nodes are often employed for pattern recognition problems, while linear or purelin(x) transfer function is applied for function fitting problems shows the purelin transfer function (Ghaedi and Vafaei 2017). The mathematical definitions of some widely used differentiable activation or transfer functions are given as follows (Yetilmezsoy and Sapci-Zengin 2009; Ghaedi and Vafaei 2017): Function graphs

Mathematical definitions y ¼ logsigðxÞ ¼ f ðxÞ ¼ ð1þe1 x Þ

+1 0

(2)

n

–1 +1 0

y ¼ tansigðxÞ ¼ f ðxÞ ¼ ð1þe22x Þ  1 ¼

ð1e2x Þ

(3)

ð1þe2x Þ

n

–1

(continued)

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Function graphs

Mathematical definitions

1.0

y ¼ radbasðxÞ ¼ f ðxÞ ¼ ex

2

2007

(4)

0.5 0.0

n –0.833

+0.833

y = purelin(x) = f(x) = x

+1

(5)

n 0 –1

Among the many types of ANNs, backpropagation (BP) networks have recently been considered as one of the simplest and most widely used network models (Cai et al. 2009). The learning process of a BP network consists of two main iterative steps: forward computing of data stream and backward propagation of error signals. During forward computing, original data are transmitted from the input layer to the output layer through the hidden processing layer, with the neurons of each layer only affecting the neurons of the succeeding layer. One of the main advantages of BP networks over other types of networks is that if the desired output cannot be obtained from the output layer, the error is propagated backwards through the network against the direction of forward computing (Cai et al. 2009; Liu and Meng 2009). According to the error signal of BP, the network changes the network connection of all layers to determine the best weight set and realize the correct network output (Liu and Meng 2009). Therefore, with these two steps performing iteratively, the error between network output and desired output can be minimized using the delta rule (Cai et al. 2009). The network training is a process by which the connection weights and biases of the ANN are adapted through a continuous process of simulation by the embedded network’s environment. The training function applies the inputs to the new network, calculates the outputs, compares them to the associated targets, and calculates a mean square error. If the error goal is met, or if the maximum number of epochs is reached, the training is stopped and the training function returns the new network and a training record. Otherwise, the training goes through another epoch. During the adaptation phase, the training algorithm receives part of the data (inputs and outputs) and automatically develops the ANN model. After development, the model could generate the appropriate responses for simulations with varying levels of data input. When the learning is complete, the neural network is used for prediction. The primary goal of training is to minimize an error function by searching for a set of connection strengths and biases that causes the ANN to produce outputs equal or close to the targets. In other words, the training aims at estimating the parameters (WHij, WOj, bj, and bo) by minimizing an error function (Yetilmezsoy et al. 2011a). As data set was trained, the input pattern given to the input layers of the network would compute the output in the output layer. The BP learning rule defined a method to adjust the weights of the networks (Antwi et al. 2017). A BP algorithm as one of the strongest learning algorithms is a gradient descent algorithm that can be employed to learn these multilayer feed-forward networks with differentiable transfer functions.

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The learning method is based on a gradient search, with a criterion of errors between the values of network output and desired output (Ghaedi and Vafaei 2017): E¼

N X

ðOn  Od Þ2

(6)

i¼1

where E is the total sum squared error of all data in the training set, in which On is the network output for the nth data and Od is the desired output. In the training process, the weights of all the connecting nodes are modified until the required error level is obtained or the maximum number of iteration is reached. In order to minimize the total error of the network trained by BP algorithm, the weights are adjusted according to the following equation (Ghaedi and Vafaei 2017; Pendashteh et al. 2011): Δwnki ðm þ 1Þ ¼ η 

@E þ μ  Δwnki ðmÞ @Δwnki

(7)

where Δwkin(m) is the correction of the weight at the mth learning step, η is the training rate (a small parameter to alter the correction each time), and μ is the momentum factor (decrease an oscillation and helps quick convergence). Network learning adjusts using suitable values of these parameters. Because gradient decent usually slows down near minima, so the Levenberg–Marquardt algorithm (LMA) method can be used to obtain faster convergence. LMA is a blend of simple gradient descent and the Gauss–Newton method. The algorithm for parameter updating is presented by the following equation (Pendashteh et al. 2011):  1 Δw ¼  J T J þ μI J T e

(8)

where e = [e1 e2 . . . eP]T is the error vector. μ is a positive constant, I is the identity matrix, and J is the Jacobian matrix given by (Pendashteh et al. 2011): 2

@e1 =@w1 6 @e2 =@w1 6 J¼6 6 4 @eP =@w1

@e1 =@w2 @e2 =@w2 @eP =@w2



3 @e1 =@wN @e2 =@wN 7 7 7 7 5 @eP =@wN

(9)

In general, ANNs are sensitive to the number of neurons in their hidden layers. Too few neurons may lead to underfitting. Conversely, too many neurons may contribute to overfitting, wherein all training points fit well, although the fitting curve may take wild oscillations between the points. In this case, the error on the training set is driven to a very small value, however, when new data are presented to the network, the error becomes enlarged. Although the network has memorized the training examples, it has not learned to generalize to new situations. This can be prevented either by training

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with Bayesian regulation, a modification of the Levenberg–Marquardt algorithm (LMA), or by using early stopping with any of the other training routines. In turn, this requires that the user pass a validation set to the training algorithm, in addition to the standard training set (Akkoyunlu et al. 2010). However, in practice, it is difficult to know which training algorithm will perform fastest for a given problem. It will depend on many factors, including the complexity of the problem and the number of data points in the training set (Yetilmezsoy and Saral 2007). In general, on networks that contain up to a few hundred weights, the LMA will have the fastest convergence. It has found to be the fastest method for training moderate sized feedforward ANN, where the training rate is 10 to 100 times faster than the usual gradient descent BP method (Al-Daoud 2009). However, when the number of network weights is large, the requirement for computation and memory becomes significant. Since in LMA, inversion of square matrix JTJ + μI is involved; thus, a large memory space is required to store the Jacobian matrix and the Hessian matrix (JTJ) along with inversion of approximated Hessian matrix in each iteration (Pendashteh et al. 2011). The Quasi–Newton methods are often the next fastest algorithms on networks of moderate size, while the Broyden–Fletcher–Goldfarb–Shanno (BFGS) Quasi–Newton BP algorithm is generally faster than the conjugate gradient algorithms. Of the conjugate gradient algorithms, the Powell-Beale procedure requires the most storage, but usually has the fastest convergence. Meanwhile, the Polak–Ribiére has performance similar to the Powell–Beale, the storage requirements for which (4 vectors) are slightly larger than for the Fletcher–Reeves (3 vectors). The Fletcher–Reeves generally converges in fewer iterations than the Resilient backpropagation algorithm (Rprop). Although more computation is required in each iteration, the Rprop and the scaled conjugate gradient algorithm do not require a line search and have small storage requirements. They are reasonably fast and are very useful for large problems. The variable learning rate algorithm is usually much slower than the other methods and has approximately the same storage requirements as Rprop; however, it can still be useful for some problems. The one-step secant algorithm requires less storage and computation per epoch than does the BFGS algorithm; however, it requires slightly more storage and computation per epoch than do the conjugate gradient algorithms. This algorithm can be considered a compromise between the Quasi–Newton algorithms and the conjugate gradient algorithms. In the batch gradient methods, the weights and biases are updated in the direction of the negative gradient of the performance function. The scaled conjugate gradient (SCG) algorithm uses a step size scaling mechanism and avoids a time-consuming line-search per learning iteration, which takes the algorithm faster than other second order conjugate gradient algorithms, Quasi–Newton algorithms, and heuristics algorithms. Therefore, this method shows superlinear convergence on most problems (Zakaria et al. 2010). The loss on the optimality of the estimates/predictions produced by some other training algorithms may be attributed to the combinatorial nature and nonlinear structure of the considered problem. Therefore, the complexity analysis of the present problem can be validated by the results of several training algorithms used in the benchmark comparison. Based on the above-mentioned facts, it can be noted that the performance of the various algorithms can be affected by the accuracy required of the

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approximation, which is dependent on the mean square error, versus that of several representative algorithms. When the problem formulation has a combinatorial nature, the definition of each process parameter results in a complex interaction of variables used in the calculations. A number of benchmark comparisons of the various training algorithms are needed in order to choose the best-suited algorithm for obtaining a good performance on the laborious interactive and nonlinear problems. In general, the LMA will have the fastest convergence on combinatorial function approximation (or nonlinear regression) problems (Akkoyunlu et al. 2010). Since ANN-based models contain no preconceptions regarding what the model shape will be, they are ideal for cases with low system knowledge. They are useful for functional prediction and system modeling where the physical processes are not understood or are highly complex. Consequently, it is believed that ANN-based techniques, which have recently been applied to various environmental problems, may provide a good alternative to statistical and theoretical techniques, as well as to iterative problems, because of their speed and capability of learning, robustness, nonlinear characteristics, nonparametric regression capabilities, generalization properties, and ease of working with regards to high-dimensional data.

Fuzzy Logic (FL) The fuzzy logic system based on linguistic expressions includes uncertainty rather than numerical probabilistic, statistical, or perturbation approaches. Fuzzy set theory (Zadeh 1965) was introduced to provide a definition for uncertainties caused by imprecision and vagueness present in real-world applications (Ozcan et al. 2009; Nasiri and Huang 2008). Rihani et al. (2009) reported that fuzzy logic has recently become a useful tool for modelling highly complex systems whose behaviors are not well understood. For instance, considering the complex qualitative relationships among the variables in a water-in-oil emulsion system, the fuzzy logic methodology has the advantage of the relatively simple mathematical calculations in linguistic terms instead of complicated equations used in the conventional methods. Since a fuzzy logic-based model does not need to handle tedious empirical formulations and complex mathematical expressions, this technique provides a transparent and a systematic analysis for the interpretation of dynamic behavior of a water-in-oil emulsion-based problem by a set of logical connectives (Yetilmezsoy et al. 2012). The key idea in fuzzy logic, in fact, is the allowance of partial belongings of any object to different subsets of a universal set instead of belonging to a single set completely. This is an artificial intelligence method utilizes fuzzy sets and linguistic terms to describe the complex qualitative relationships between model components (Ozcan et al. 2009; Nasiri and Huang 2008; Rihani et al. 2009). There are basically five parts of the fuzzy inference process: In the first step (fuzzification), crisp numerical inputs and outputs are divided into different fuzzy categories associated with linguistic terms (i.e., low, high, big, small, too-cold, cold, warm, hot, too-hot, young, old, etc.), where the output is

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always a fuzzified degree of a specific membership function within the range from 0 to 1 (Jantzen 1999; Altunkaynak et al. 2005; Sozen et al. 2004). Instead of a definition for the developed fuzzy set categories such as moderately low, low, moderate, moderately high, high, etc., the membership functions can be defined as A, B, C, D, E, etc., to simplify processing of the rules (Yetilmezsoy et al. 2012). Since multiple measured crisp inputs first have to be mapped into the specific fuzzy membership functions, Sozen et al. (2004) reported that the fuzzification process requires good understanding of all the variables. Before the rules can be evaluated, the inputs must be fuzzified according to each of these linguistic sets. In second step, after the inputs are fuzzified, the fuzzy operator (AND or OR) in different pieces of the antecedent is performed in the fuzzy inference system (FIS) for each fuzzy rule. It is noted that the fuzzy rule base contains some rules that include all possible fuzzy relations between inputs and output variables (or actions and conclusions). In fuzzy set theory, there are no mathematical equations and model parameters, and therefore, all the uncertainties, nonlinear relationships, and model complications are included in the descriptive fuzzy inference procedure in the form of if-then (if premise then consequent) logical statements, called fuzzy rules (Rihani et al. 2009; Akkurt et al. 2004; Acaroglu et al. 2008). If the antecedent of a given rule has more than one part, the fuzzy operator is applied to obtain one number that represents the result of the antecedent for that rule. This number is then applied to the output function. The input to the fuzzy operator is two or more membership values from fuzzified input variables. The output is a single truth value. Two kinds of built-in AND methods (min (minimum) and prod (product): prod(a,b) = ab), and two kinds of built-in OR methods (max (maximum) and probor (the probabilistic OR method: probor (a,b) = a + b  ab)) can be used in the fuzzy logic toolbox (Altunkaynak et al. 2005; Sozen et al. 2004; Akkurt et al. 2004). In the third step, an implication process from the antecedent to the consequent is performed in the FIS. This procedure is defined as the shaping of the consequent (a fuzzy set) based on the antecedent (a single number). The input for the implication process is a single number given by the antecedent, and the output is a fuzzy set (Kusan et al. 2010). Before applying the implication method, the rule’s weight must be determined. Every rule has a weight (a number between 0 and 1), which is applied to the number given by the antecedent. After proper weighting has been assigned to each rule, the implication method is implemented. A consequent is a fuzzy set represented by a membership function, which weights appropriately the linguistic characteristics that are attributed to it. The consequent is reshaped using a function associated with the antecedent (a single number). The input for the implication process is a single number given by the antecedent, and the output is a fuzzy set. Implication is implemented for each rule. For this process, two built-in methods are basically supported by the fuzzy logic toolbox, and they are the same functions that are used by the AND operator: min (minimum), which truncates the output fuzzy set, and prod (product), which scales the output fuzzy set (Altunkaynak et al. 2005; Sozen et al. 2004; Akkurt et al. 2004).

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In the fourth step, aggregation process is performed to fuzzy sets to obtain a single fuzzy set that represents the outputs of each fuzzy rule. Because decisions are based on the testing of all of the rules in a FIS, the rules must be combined in some manner in order to make a decision. Therefore, aggregation is the process by which the fuzzy sets that represent the outputs of each rule are combined into a single fuzzy set. The input of the aggregation process is the list of fuzzy sets that represent the outputs of each rule. Aggregation only occurs once for each output variable, just prior to the fifth and final step, defuzzification. The input of the aggregation process is the list of truncated output functions returned by the implication process for each rule. The output of the aggregation process is a fuzzy set. There are a number of aggregation methods (i.e., max (maximum), sum (simply the sum of each rule’s output set), probor, etc.) supported by the FIS (Altunkaynak et al. 2005; Sozen et al. 2004; Akkurt et al. 2004). The nature of the information retrieval dictates that the determination of the ranking should be done based on all of the rules. In this case, the sum aggregation method appears to be a much better fit (Rubens 2006). Finally, the defuzzifier produces the crisp values corresponding to the final fuzzy outputs as a conclusion (Jantzen 1999). The input for the defuzzification process is a fuzzy set (the aggregate output fuzzy set) and the output is a single number. There are many defuzzification methods such as center of gravity (COG or; centroid), bisector of area (BOA), mean of maxima (MOM), leftmost maximum (LM), rightmost maximum (RM), etc. (Nasiri and Huang 2008). In the defuzzification step, linguistic results obtained from the fuzzy inference are translated into a crisp numerical output (real value) by using the rule base provided (Kusan et al. 2010; Biyikoglu et al. 2005). In the literature, several defuzzification methods, such as center of gravity (COG or centroid), bisector of area, mean of maxima, leftmost maximum, rightmost maximum, have been reported (Jantzen 1999). It is apparent from several fuzzy logic-based studies (Turkdogan-Aydinol and Yetilmezsoy 2010; Yetilmezsoy et al. 2012; Altunkaynak et al. 2005; Akkurt et al. 2004; Rubens 2006; Sadiq et al. 2004), centroid method is most widely used defuzzification technique, since it satisfies the underlying properties of the system and exhibits the best performance. It is determined as follows (Turkdogan-Aydinol and Yetilmezsoy 2010; Yetilmezsoy et al. 2012; Sozen et al. 2004; Akkurt et al. 2004): n P

μðyi Þyi ðyi Þd ¼ i¼1 n P μ ðyi Þ

(10)

i¼1

where (yi)d is the defuzzified output, yi is the output value (or the centroidal distance from the origin) in the ith subset, and μ(yi) is the membership value of the output value in the ith subset. For the continuous case, the summations in Eq. (5) are replaced by integrals, as given by Sadiq et al. (2004). On the basis of above-

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mentioned fuzzy steps, a detailed schematic of a sample MISO (multiple inputs and single output) fuzzy system is depicted in Fig. 2. The situations of uncertainties in fuzzy-logic are defined via giving appropriate membership functions to the elements of the set that represent the situation. The value of the variation between 0 and 1 (the highest level) for each element is called membership degree and its value in subset is called membership function (Topcu and Saridemir 2008). In fuzzy models, the shape of membership functions of fuzzy sets can be triangular, trapezoidal, bell-shaped, sigmoidal, or another appropriate form, depending on the nature of the system being studied (Acaroglu et al. 2008;

START INTRODUCE INPUTS

CH4

CO

WS

WD

RH

SOLAR

O3

CRISP NUMERICAL INPUTS

FO3

FUZZY INPUTS

FUZZIFICATION

FCH4

FCO

FWS

FWD

FRH

FSOLAR

RULE 1: IF .... THEN RESULT 1

DATA BASE

RULE 2: IF .... THEN RESULT 2

DECISION MAKING LOGIC

RULE N: IF .... THEN RESULT N

FUZZY INFERENCE SYSTEM (FIS) (Mamdani)

RULE BASE

AND METHOD F - DUST AGGREGATION METHOD

SUM OPERATOR

END

DEFUZZIFICATION

DUST Concentr.

FUZZY OUTPUT

IMPLICATION METHOD

CENTROD METHOD PROD OPERATOR

CRISP NUMERICAL OUTPUT

Fig. 2 A detailed schematic of a sample MISO fuzzy system (Adapted from Yetilmezsoy and Abdul-Wahab 2012)

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K. Yetilmezsoy

Metternicht and Gonzalez 2005). Among them, triangular and trapezoidal shaped membership functions are predominant in current applications of fuzzy set theory, due to their simplicity in both design and implementation based on little information (Yetilmezsoy et al. 2012; Rihani et al. 2009). A schematic overview of the trapezoidal-based membership function is given in Fig. 3. The trapezoidal curve is the membership function of a vector, x, and depends on four scalar parameters, a, b, c, d, as follows (Turkdogan-Aydinol and Yetilmezsoy 2010; Yetilmezsoy et al. 2012; Altunkaynak et al. 2005; Sozen et al. 2004; Adriaenssens et al. 2006): 9 8 0, x  a > > > > x  a > > > > > > , a < x < b > > > > b  a > > > > = < μðxÞ ¼ μðx; a, b, c, d Þ ¼ 1, b  x  c > > > > > > > > d  x > > > > > > , c < x < d > > > > d  c ; : 0, x  d

(11)

In the applications of the fuzzy system in both control and forecasting, there are two types of fuzzy inference systems, namely, Mamdani-type (Mamdani and Assilian 1975) and Takagi-Sugeno-type (Takagi and Sugeno 1985) fuzzy systems (Rihani et al. 2009; Ozger and Sen 2007; Sadrzadeh et al. 2009). Sadrzadeh et al. (2009) reported that each if-then rule produces a fuzzy set for the output variable in the Mamdani approach, and hence defuzzification step is indispensable to obtain crisp values of the output variable. Because of allowing a simplified representation and interpretation of the fuzzy rules, Mamdani’s fuzzy inference method is the most commonly applied fuzzy methodology (Turkdogan-Aydinol and Yetilmezsoy 2010; Yetilmezsoy et al. 2012; Akkurt et al. 2004; Acaroglu et al. 2008; Adriaenssens et al. 2006; Traore et al. 2005).

Fig. 3 A schematic overview of the trapezoidal-based membership function (Adapted from Yetilmezsoy et al. 2011a)

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Adaptive Neuro-Fuzzy Inference System (ANFIS) The ANFIS consists of two parts, antecedent and conclusion, which are connected to each other by fuzzy rules based on the network form. Since the consequent parameters are calculated forward, while the premise parameters are calculated backward, operation of the ANFIS looks like feed-forward back propagated (FFBP) ANN (Atmaca et al. 2001). Zero or first-order Sugeno inference systems or Tsukamoto inference system can be used in the fuzzy section. The output variables ( fi) are then obtained by performing several fuzzy rules to fuzzy sets of input variables (Yetilmezsoy et al. 2011a, 2015; Atmaca et al. 2001; Cakmakci et al. 2010): Rule 1 :

If x is A1 and y is B1 ,

then f 1 ¼ p1 x þ q1 y þ r 1

(12)

Rule 2 :

If x is A2 and y is B2 ,

then f 2 ¼ p2 x þ q2 y þ r 2

(13)

where p1, p2, q1, q2, r1, and r2 are linear parameters, and A1, A2, B1, and B2 are the nonlinear parameters. The ANFIS architecture (equivalent of a two input first-order Sugeno FIS model) including the input (x and y) of nodes (A1, A2, B1, and B2), membership functions ðμAi ðxÞ or μBj ðyÞÞ, membership grades (or outputs of layers) of the fuzzy sets (Q1,i, Q2,i, Q3,i, Q4,i, Q5,i), weight functions of the next layers (w1 and w2), normalized firing strengths (w1 and w2 ), and the consequent parameters ( p1, q1, r1, p2, q2, r2) is illustrated in Fig. 4. As seen in Fig. 4, the equivalent ANFIS architecture consists of five layers: Fuzzy layer, product layer (π), normalized layer (N), defuzzy layer, and total output layer (Yetilmezsoy et al. 2011a, b, 2015; Cakmakci et al. 2010). As seen in Fig. 4, Layer 1 is the fuzzy layer, in which x and y are the input of nodes A1, A2, B1, and B2, respectively. A1, A2, B1, and B2 are the linguistic labels used in the fuzzy theory for dividing the membership functions. Parameters in this layer are referred to as premise parameters. Every node i in Layer 1 is an adaptive node with a specific function. Nodes in Layer 1 implement fuzzy membership functions, mapping input variables to corresponding fuzzy membership values. The membership relationship between the output and input functions of this layer can be expressed as (Yetilmezsoy et al. 2011a, b): Q1i ¼ μAi ðxÞ, for i ¼ 1, 2 or;

(14)

Q1i ¼ μBi ðyÞ, for i ¼ 1, 2

(15)

where x or y is the input to node i, and Ai or Bi is the linguistic label (such as small, large, etc.) associated with this node function, Q1i denotes the output functions, and μAi(x) or μBi( y) usually denotes the bell-shaped membership functions with a maximum equal to 1 and a minimum equal to 0, such as (Yetilmezsoy et al. 2011a, b; Jang 1993; Esmaeelzadeh and Dariane 2014):

Fig. 4 A five-layer ANFIS architecture (equivalent of a two input first-order Sugeno FIS model consisting of two inputs and rules) (Adapted from Yetilmezsoy et al. 2015)

2016 K. Yetilmezsoy

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Applications of Soft Computing Methods in Environmental Engineering

μAi ðxÞ ¼ 1þ



1

xci ai

2 bi or;

"

# x  ci 2 μAi ðxÞ ¼ exp  ai

2017

(16)

(17)

where (ai, bi, and ci) is the parameter set. As the values of these parameters change, the bell-shaped functions vary accordingly, thus exhibiting various forms of membership functions on linguistic label, Ai. In fact, any continuous and piecewise differentiable functions, such as commonly used trapezoidal and triangular-shaped membership functions, are also be used as node functions in this layer (Jang 1993). Layer 2 is the product layer that consists of two fixed circle nodes labelled π, which multiply the incoming signals and provides the outputs of the product. The output w1 and w2 are the weight functions of the next layer. The output of this layer is the product of the input signal, which is defined as follows (Yetilmezsoy et al. 2011a, b; Jang 1993; Esmaeelzadeh and Dariane 2014): Q2i ¼ wi ¼ μAi ðxÞ μBi ðyÞ, for i ¼ 1, 2

(18)

where Q2i denotes the output of Layer 2. Each node output represents the firing strength of a rule. The third layer is the normalized layer, whose nodes are labelled N. The ith node calculates the ratio of the ith rules firing strength to the sum of all rule’s firing strengths. Its function is to normalize the weight function in the following process (Yetilmezsoy et al. 2011a, b; Jang 1993; Esmaeelzadeh and Dariane 2014): Q3i ¼ wi ¼

wi , for i ¼ 1, 2 w1 þ w2

(19)

where Q3i denotes the output of Layer 3. The outputs of this layer are called normalized firing strengths. The fourth layer is the defuzzy layer, whose nodes are adaptive. Every node i in this layer is an adaptive node with a specific function. The output equation is wi ðpi x þ qi y þ r i Þ , where pi, qi, and ri denote the linear parameters or so-called consequent parameters of the node. The defuzzy relationship between the input and output of this layer can be defined as follows (Yetilmezsoy et al. 2011a, b; Jang 1993; Esmaeelzadeh and Dariane 2014): Q4i ¼ wi f i ¼ wi ðpi x þ qi y þ r i Þ, for i ¼ 1, 2

(20)

where Q4i denotes the output of Layer 4. The fifth layer is the total output layer, whose node is labelled Σ. The output of this layer is the total of the input signals, which represents the vehicle shift decision

2018

K. Yetilmezsoy

result. The results can be written as (Yetilmezsoy et al. 2011a, b; Jang 1993; Esmaeelzadeh and Dariane 2014): Q5i ¼ overall output ¼

X i

P wi f wi f i ¼ Pi i i wi

(21)

where Q5i denotes the output of Layer 5. Although ANN and fuzzy logic models are the basic areas of artificial intelligence concept, the ANFIS combines these two methods and uses the advantages of both methods. Since the ANFIS is an adaptive network which permits the usage of ANN topology together with fuzzy logic, it includes the characteristics of both methods and also eliminates some disadvantages of their lonely used case. Therefore, this technique it is capable of handling complex and nonlinear problems. Even if the targets are not given, the ANFIS may reach the optimum result rapidly. In addition, there is no vagueness in ANFIS as opposed to ANNs (Atmaca et al. 2001; Jang et al. 1997). Moreover, the learning duration of ANFIS is very short compared to ANN-based models. It implies that ANFIS may reach to the target faster than ANN. Therefore, when a more sophisticated system with a high-dimensional data is implemented, the use of ANFIS instead of ANN would be more appropriate to overcome faster the complexity of the problem (Atmaca et al. 2001). In the ANFIS structure, the implication of the errors is different from that of the ANN case. In order to find the optimal result, the epoch size is not limited. In training of high-dimensional data, the ANFIS can give results with the minimum total error compared to ANN and fuzzy logic methods. Moreover, fuzzy logic method seems to be the worst in contrast to others at a first look, since the rule size is limited and the number of membership functions of fuzzy sets were chosen according to the intuitions of the expert. However, if different types of membership functions and their combinations had been tested and more membership variables and more rules had been used to enhance the prediction performance of the proposed diagnosis system, better results would have been available (Turkdogan-Aydinol and Yetilmezsoy 2010; Atmaca et al. 2001).

Support Vector Machines (SVM) The SVM is a linear machine of one output y(x), working in the high dimensional feature space formed by the nonlinear mapping of the N-dimensional input vector x into a K-dimensional feature space (K > N ) through the use of the nonlinear function φ(x). The number of hidden units (K ) is equal to the number of so-called support vectors that are the learning data points, closest to the separating hyperplane. The learning task is transformed to the minimization of the error function, while keeping the weights of the network at minimum. The error function is defined through the so-called e-insensitive loss function Le(d, y(x)) (Vapnik 1998; Osowski and Garanty 2007; Yeganeh et al. 2012):

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Applications of Soft Computing Methods in Environmental Engineering

 Le ðd, yðxÞÞ ¼

jd  yðxÞj  e for jd  yðxÞj  e, 0 for jd  yðxÞj < e,

2019

(22)

where e is the assumed accuracy, d is the destination, x the input vector, and y(x) the actual output of the network under excitation of x and the actual output signal of the SVM network is defined by yðxÞ ¼

K X

wj φj ðxÞ þ b ¼ wT φðxÞ þ b,

(23)

j¼1

where w = [w1, . . ., wK]T is the weight vector, b the bias, and φ(x) = [φ1(x), . . ., φK(x)]T the basis function vector. The solution of the so defined optimization problem is solved by the introduction of the Lagrangian function and the Lagrange multipliers αi , α0i (i = 1, 2, . . ., p) responsible for the functional constraints defined by (1). The minimization of the Lagrangian function has been transformed to the so-called dual problem (Vapnik 1998; Platt 1998; Osowski and Garanty 2007; Yeganeh et al. 2012): ( ) p p p X p   X X     1X    0 0 0 0 max d i αi  αi  e αi  αi  αi  αi αj  αj K xi , xj 2 i¼1 j¼1 i¼1 i¼1 (24) at the constraints p X 

 αi  α0i ¼ 0,

0  αi  C,

0  α0i  C,

(25)

i¼1

where K(xi, xj) = φΤ(xi)φ(xj) is an inner-product kernel defined in accordance with the Mercer’s theorem (Vapnik 1998) for the learning data set x. After solving the dual problem, all weights are expressed through the Nsv nonzero Lagrange multipliers αi , α0i and the same number of learning vectors xi associated with them. The network output signal y(x) can be then expressed in the form (Vapnik 1998; Osowski and Garanty 2007; Yeganeh et al. 2012): y ð xÞ ¼

N SV X 

 αi  α0i K ðx, xi Þ þ b

(26)

i¼1

The most known kernel functions used in practice are radial (Gaussian), polynomial, spline, or even sigmoidal functions (Vapnik 1998; Schölkopf and Smola 2002). The most important is the choice of coefficients e and C. Constant e determines the margin within which the error is neglected. The smaller its value the higher accuracy of learning is required, and more support vectors will be found by the algorithm. The regularization constant C is the weight, determining the balance between the

2020

K. Yetilmezsoy

complexity of the network, characterized by the weight vector w and the error of approximation, measured by the slack variables and the value of e (Osowski and Garanty 2007; Yeganeh et al. 2012). For the normalized input signals, the value of e is usually adjusted in the range (103–102), and C is much bigger than 1 (Osowski and Garanty 2007).

Implementation of Soft Computing Methods in Environmental Engineering In this section, successful applications of soft computing-based prediction models (ANN, FL, ANFIS, SVM) in the field of environmental engineering are examined in terms of water/wastewater treatment and air pollution related problems, and the important findings obtained in these studies are summarized.

ANN-Based Applications for Water and Wastewater Treatment Yetilmezsoy et al. (2013) developed two three-layer ANN models to predict biogas and methane production rates in a pilot-scale mesophilic up-flow anaerobic sludge blanket (UASB) reactor treating molasses wastewater. A tangent sigmoid transfer function (tansig) at the hidden layer and a linear transfer function (purelin) at the output layer were conducted for the proposed ANN models. After backpropagation training combined with principal component analysis (PCA), the scaled conjugate gradient algorithm (trainscg) was found as the best of the other training algorithms. Computational results demonstrated that compared to the conventional multiple regression-based methodology, the proposed ANN-based models produced smaller deviations and exhibited superior predictive accuracy with satisfactory determination coefficients of about 0.935 and 0.924, respectively, for the forecasts of biogas and methane production rates. In a recent study, Podder and Majumder (2016) proposed a three-layer feedforward back propagation (BP) ANN (4:5:1) with Levenberg–Marquardt (LM) training algorithm for predicting the phycoremediation efficiency of both As(III) and As(V) ions from wastewater using Botryococcus braunii. The study concluded that the proposed ANN architecture exhibited good agreements with the actual experimental and predicted values of both As(III) and As(V) and could describe the behavior of the complex reaction system with very high determination coefficient (R2 = 0.99977 and 0.9998 for As(III) and As(V), respectively) under different conditions. More recently, Antwi et al. (2017) developed three-layered feedforward backpropagation (BP) ANN and multiple nonlinear regression (MnLR) models were to estimate biogas and methane yield in an upflow anaerobic sludge blanket (UASB) reactor treating potato starch processing wastewater. In the study, QuasiNewton method and conjugate gradient backpropagation (BP) algorithms were found as the best among other training algorithms. The authors have reported that

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compared with the MnLR model, BP-ANN model demonstrated significant performance, suggesting possible control of the anaerobic digestion process with the BP-ANN model. In another recent work, Ghaedi and Vafaei (2017) reviewed important research studies of ANN on dyes adsorption from aqueous solution. The study concluded that ANN approaches could be successfully applied for the modeling and forecasting of dye adsorption process with acceptable accuracy compared to conventional linear models such as multiple linear regression (MLR) and PLS. In particular, the hybrid networks with optimization approaches were found to be more efficient to the performance of dye adsorption. Furthermore, Qaderi and Babanezhad (2017) attempted to employ a feed-forward ANN-based model with four hidden layer and nine independent variables (i.e., concentrations of ions K, Na, Mg, Ca, Sr, Ba, CO3, HCO3, NO3, Cl, and SO4) to predict the costs of water treatment through reverse osmosis process for supplying drinking water from the available water resources. The results concerning the ANN indicated that the proposed predictive model performed desirably for estimating the costs of treating the groundwater in the region with the accuracy of approximately 98%, where the root mean square error (RMSE) percentage was 2.02% indicating an acceptable error level for the ANN model. Finally, Hu et al. (2017) developed a three-layer backpropagation BP-ANN model to predict the chemical oxygen demand (COD) removal performance of an expanded granular sludge bed (EGSB) reactor. Activation function of hidden layer and output layer were “tansig” and “purelin” individually. Several comparisons were conducted to obtain an optimal network structure. Dividerand function was chosen to divide the operating data into training group, testing group, and validation group. The Levenberg–Marquardt algorithm (trainlm) was found as the best of the tested training algorithms. The result indicated that the proposed ANN model exhibited high forecast accuracy (R2 = 0.8156) for the forecast of COD removal performance by EGSB system. Apart from the above-mentioned studies, several other successful ANN modeling studies (Oliveira-Esquerre et al. 2002; Molga et al. 2006; Sahinkaya et al. 2007; Daneshvar et al. 2006; Rangasamy et al. 2007; Raduly et al. 2007; Ozkaya et al. 2007, 2008; Yetilmezsoy and Demirel 2008; Yetilmezsoy and Sapci-Zengin 2009; Yetilmezsoy 2012; Sahinkaya 2009; Pendashteh et al. 2011) have been conducted previously in various parts of the field of wastewater engineering (Fig. 5).

ANN-Based Applications for Air Quality/Pollution Control/ Forecasting In the past years, it has become apparent that ANN-based prediction models have been effectively conducted on a substantial number of research activities in the field of air pollution engineering. In these investigations, several authors have developed different types of ANN models, and the results have been compared with the forecasts obtained using multiple regression models. For instance, Nunnari et al. (2004)

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K. Yetilmezsoy

Fig. 5 Various topological architectures of ANN models proposed for water and wastewater treatment (Adapted from (a) Yetilmezsoy 2010; (b) Podder and Majumder 2016; (c) Yetilmezsoy et al. 2013; (d) Qaderi and Babanezhad 2017; and (e) Hu et al. 2017)

modeled SO2 concentration at a point by intercomparing several stochastic techniques such as ANN, fuzzy logic, and generalized additive techniques. Because the ANN models worked better in the prediction of critical episodes, they recommended the ANN approach for the implementation of a warning system for air quality control. Yetilmezsoy (2006) proposed an ANN model and a new empirical model to determine optimum body diameter (OBD) of air cyclones for 505 different artificial scenarios given in a wide range of five operating variables, namely, gas flow rate, particle density, temperature, and two design parameters, namely, Ka and Kb, selected in the cyclone design. The study concluded that maximum diameter deviations from the well-known Kalen and Zenz’s model were recorded as 1.3 cm and 0.0022 cm for the empirical model and ANN outputs, respectively. Although both approaches

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produced promising results, the ANN model exhibited speed and practicality, as well as a more robust and superior performance in the prediction of OBD values. Agirre-Basurko et al. (2006) developed two multilayer perceptron (MLP)-based models and one multiple linear regression-based model to forecast ozone (O3) and nitrogen dioxide (NO2) levels in Bilbao, Spain. In their study, traffic variables were used as predictor variables in the developed models. Results indicated the MLP-based models showed remarkably better performance than the multiple linear regression model in predicting pollutant concentrations. In another study (Yetilmezsoy and Saral 2007), an ANN-based approach and nonlinear regression analysis were performed for the determination of single droplet collection efficiency (SDCE) of countercurrent spray towers. The authors reported that predicted results obtained from the nonlinear regression analysis and the ANN model were in agreement with the theoretical data, and that all predictions proved to be satisfactory with a correlation coefficient of approximately 0.921 and 0.99, respectively. The study concluded that the development of a new mathematical model and the creation of an ANN-based model for the prediction of SDCE of countercurrent spray towers eliminated complex interactions of variables and difficult iterative calculations typically performed in the theoretical approach. Finally, there have also been other studies (Wieland et al. 2002; Wotawa and Wotawa 2001; Abdul-Wahab and Al-Alawi 2002; Iliadis et al. 2007; Al-Alawi et al. 2008; Ozdemir et al. 2008) on the prediction of tropospheric and surface O3 concentrations reporting the advantages and adaptability properties of ANN-based models. Moreover, the use of ANN allows the prediction of daily and/or hourly particulate matter (PM2.5 and PM10) emissions (Chaloulakou et al. 2003; Chelani 2005; Grivas and Chaloulakou 2006; Kurt et al. 2008; Feng et al. 2015; Vakili et al. 2015; Bai et al. 2016; Biancofiore et al. 2017; Park et al. 2018) in many urban and residential areas. ANN-based models have also been used in the prediction of urban and ground-level SO2 concentrations, demonstrating successful results when considering the complex and nonlinear structure of the atmosphere (Akkoyunlu et al. 2010; Saral and Erturk 2003; Sofuoglu et al. 2006; Bai et al. 2016). Furthermore, ANN-based models have given reliable forecasts of carbon monoxide (CO) and nitrogen dioxide (NO2) concentrations in other studies (Kurt et al. 2008; Elangasinghe et al. 2014; Bai et al. 2016) (Fig. 6).

FL-Based Applications for Water and Wastewater Treatment Murnleitner et al. (2002) modelled and controlled two-stage anaerobic wastewater pretreatment using a Mamdani-type FL expert system. Hydrogen concentration together with methane concentration, gas production rate, pH, and the filling level of the acidification buffer tank were used as input variables for the FL system. With the use of the proposed FL system, very strong fluctuations in the concentration of the substrate and the volumetric loading rate could be successfully handled, and heavy overload could be avoided by taking proper control actions automatically.

2024

K. Yetilmezsoy

Fig. 6 Different architectures of MLP type ANN models proposed for air quality/pollution control/ forecasting (Adapted from (a) Yetilmezsoy and Saral 2007; (b) Kurt et al. 2008; (c) Elangasinghe et al. 2014; and (d) Feng et al. 2015)

In another anaerobic study, Turkdogan-Aydınol and Yetilmezsoy (2010) developed a FL-based model to predict biogas and methane production rates in a pilotscale 90-L mesophilic up-flow anaerobic sludge blanket (UASB) reactor treating molasses wastewater. In the study, trapezoidal membership functions with eight levels were conducted for the fuzzy subsets, and a Mamdani-type fuzzy inference system was used to implement a total of 134 rules in the if-then format. The authors concluded that compared to nonlinear regression models, the proposed FL-based model produced smaller deviations and exhibited a superior predictive performance on forecasting of both biogas and methane production rates with satisfactory determination coefficients over 0.98. Yetilmezsoy (2012) proposed a multiple inputs and multiple outputs (MIMO) FL-based model was proposed to estimate color and chemical oxygen demand (COD) removal efficiencies in the posttreatment of anaerobically pretreated poultry manure wastewater (PMW) effluent using Fenton’s oxidation process. The author used trapezoidal membership functions with eight levels that were conducted for the fuzzy subsets, and a Mamdani-type fuzzy inference system to implement a total of 70 rules in the if-then format. The product (prod) and the center of gravity (centroid) methods were applied as the inference operator and defuzzification methods, respectively. The results of the study demonstrated that a highly dynamic process, such as

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Fenton’s oxidation of anaerobically pretreated PMW effluent, could be successfully (R2 = 0.99 for both color and COD removals) and cost-effectively (CPU usage = 3–4% when simulating the model) modeled using FL methodology, compared to the classical regression-based method (R2 = 0.772 and 0.861 for color and COD removals, respectively, and CPU usage = 5–6%). Furthermore, there have also been other studies on modeling water-in-oil emulsion formation (Yetilmezsoy et al. 2012), biological oxygen demand (BOD) removal prediction in free-water surface constructed wetlands (Kotti et al. 2013), and modeling of an integrated process for predictions of COD, total organic carbon (TOC), color, and ammonia nitrogen (NH3–N) removal efficiencies in the treatment of landfill leachates (young, middle-aged, and stabilized) (Sari et al. 2013) reporting robustness and cost-effectiveness of FL-based modeling tools.

FL-Based Applications for Air Quality/Pollution Control/Forecasting Yetilmezsoy and Abdul-Wahab (2012) proposed a prognostic approach that is based on a FL model to estimate suspended dust concentrations (PM10) in a specific residential area in Kuwait with high traffic and industrial influences. The authors employed trapezoidal membership functions with 10 and 15 levels employed for the fuzzy subsets of each model variable. A Mamdani-type fuzzy inference system (FIS) was developed to introduce a total of 146 rules in the if-then format. The product (prod) and the center of gravity (centroid) methods were performed as the inference operator and defuzzification methods, respectively, for the proposed FIS. The study concluded that the proposed FL model produced very small deviations from the actual results, and showed better predictive performance than an multiple regressionbased exponential model with regard to forecasting PM10 levels, with a very high determination coefficient of over 0.99. In a recent study, Olvera-García et al. (2016) described a new evaluation model using weighted fuzzy inference systems combined with an Analytic Hierarchy Process (AHP), providing a new air quality index (AQI) for Mexico City and its Metropolitan area. The authors evaluated six key pollutants (ozone (O3), sulfur dioxide (SO2), nitrogen dioxide (NO2), carbon monoxide (CO), particulate matter smaller than 10 and 2.5 μm (PM10 and PM2.5)) as environmental parameters according to toxicological levels, and assessed different air quality situations using a fuzzy reasoning process. They employed five score stages, such as excellent, good, regular, bad, and dangerous, in order to define a set of 174 inference rules in the if-then format for the proposed FIS. The results showed that a good performance of the proposed AQI against those in literature depending on the assignment of weights according to an importance level for each environmental parameter using a priority analysis based on the AHP procedure. Additionally, some other FL-based studies on classification of air quality in Tehran, Iran (Sowlat et al. 2011), assessment and prediction of air quality in Mexico City and its Metropolitan area (Carbajal-Hernández et al. 2012), and modeling the

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indoor air quality (IAQ) of the underground trains in Athens, Greece (Assimakopoulos et al. 2013), can be found in the literature.

ANFIS-Based Applications for Water and Wastewater Treatment In addition to ANN modeling studies, several ANFIS-based models have been proposed to evaluate and optimize various water and wastewater treatment processes. For instance, autoregressive integrated moving average (ARIMA) and Takagi-Sugeno (TS) fuzzy methods were used by Altunkaynak et al. (2005) for predicting future monthly water consumption values from three antecedent water consumption amounts, considered as independent variables. The TS fuzzy predicted results better than the ARIMA. Civelekoglu et al. (2007) employed ANFIS-based models for the prediction of carbon and nitrogen removal in the aerobic biological treatment stage of a full-scale WWTP treating process wastewaters from the sugar production industry. In the study, a total of six independent ANFIS models were developed with or without PCA using the correlations among the influent and effluent data from the plant. With the use of PCA, results showed that the ANFIS modeling approach could be an effective advanced technique for performance prediction and control of treatment processes. An ANFIS-based model was used by Firat and Gungor (2007) to estimate the flow of River Great Menderes, located west of Turkey. As a result, they discovered that ANFIS could be successfully applied for river flow estimation, providing high accuracy and reliability. Firat et al. (2009) compared two types of FIS for predicting municipal water consumption time series. Their results demonstrated that the ANFIS model is superior to Mamdani fuzzy inference systems (MFIS). Cakmakci (2007) used an ANFIS-based technique for modeling of anaerobic digestion system of primary sludge of the Kayseri WWTP, Turkey. In the study, effluent volatile solid (VS) and methane yield were predicted by the ANFIS model using the routinely measured parameters in the anaerobic digester. The study concluded that due to highly nonlinear structure of the ANFIS model, a highly complex system such as anaerobic digestion process could be easily modeled. Filter head loss was also estimated by Cakmakci et al. (2008) using this ANFIS model. In their study, rule base sets were generated with subtractive clustering and grid partition. They determined that using a grid partition for modeling was superior to that of subtractive clustering. The correlation coefficients were greater than 0.99 in both tap and deionized water. Furthermore, filter iron removal rate was also modeled by Cakmakci et al. (2010). The best results for tap and deionized water were obtained with grid partition and subtractive clustering. The index of agreement (IA) values for tap water and deionized water were calculated as 0.996 and 0.971, and R2 values were determined as 0.99 and 0.89, respectively. The study concluded that neurofuzzy modeling could be successfully used to predict effluent iron concentration in sand filtration. In another study, for a real-scale anaerobic WWTP operating under unsteady state conditions, Perendeci et al. (2008) proposed a conceptual ANFIS-based using

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available on-line and off-line operational input variables to estimate the effluent COD. The study concluded that the developed ANFIS model with phase vector and history extension successfully represented the behavior of the considered treatment system. A principal component analysis-adaptive neuro-fuzzy inference systems (PCA-ANFISs) method was used by Goodarzi et al. (2009) for the analysis of ternary mixtures of Al(III), Co(II), and Ni(II) over the range of 0.05–0.90, 0.05–4.05, and 0.05–0.95 g/mL, respectively. As a result, the method accurately and simultaneously determined the content of metal ions in several synthetic mixtures. Finally, there have also been other computational studies (Tay and Zhang 2000; Wu and Lo 2008; Mingzhi et al. 2009; Pai et al. 2009; Erdirencelebi and Yalpir 2011; Mullai et al. 2011; Yetilmezsoy et al. 2011b; Wan et al. 2011; Pai et al. 2011; Mandal et al. 2015; Yetilmezsoy et al. 2015; Rahimzadeh et al. 2016) in the literature for modeling of various environmental problems based on water and wastewater treatment using ANFIS methodology.

ANFIS-Based Applications for Air Quality/Pollution Control/ Forecasting Several adaptive neuro-fuzzy techniques emerging from the fusion of ANN and FIS have successfully found application in various areas of air pollution control. For instance, Yildirim and Bayramoglu (2006) used an adaptive neuro-fuzzy logic method to estimate the impact of meteorological factors on SO2 and total suspended particular matter (TSP) pollution levels over the city of Zonguldak, Turkey. The study concluded that the proposed ANFIS model satisfactorily forecasts the trends in SO2 and TSP concentration levels, with performance levels between 75–90% and 69–80%, respectively. An artificial intelligence-based modeling approach was conducted in another study by Noori et al. (2010) to predict daily carbon monoxide (CO) concentration in the atmosphere of Tehran, Iran, by means of developed ANN and ANFIS models. In the study, forward selection (FS) and gamma test (GT) methods were implemented for selecting input variables and developing hybrid models with ANN and ANFIS. The authors concluded that FS-ANN and FS-ANFIS models were the best models, considering R2, mean absolute error, and developed discrepancy ratio statistics, for predicting pollution episodes. Apart from the foregoing studies, several researchers (Shahraiyni et al. 2015; Ausati and Amanollahi 2016; Mishra and Goyal 2016; Prasad et al. 2016; Taylan 2017; Xie et al. 2017) have been successfully used ANFIS-based models in air quality/pollution control/forecasting.

SVM-Based Applications for Water and Wastewater Treatment Singh et al. (2011) used support vector classification (SVC) and support vector regression (SVR) models for (1) classification of the sampling sites with a view to

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identify similar ones in the monitoring network for reducing their number for the future water quality monitoring; (2) classification of the sampling months into the groups of seasons for reducing the annual sampling frequency; and (3) to predict the biochemical oxygen demand (BOD) of the river water using simple measurable water quality variables. They studied with the data set comprised of 1500 water samples representing 10 different sites monitored for 15 years. The study concluded that The SVC model achieved a data reduction of 92.5% for redesigning the future monitoring program, and the SVR model provided a tool for the prediction of the water BOD using set of a few measurable variables. Garcia Nieto et al. (2013) proposed a hybrid approach based on support vector regression (SVR) in combination with genetic algorithms (GA), namely, genetic algorithm support vector regression (GA-SVR) model, in forecasting the cyanotoxins presence in the Trasona reservoir, Northern Spain. The authors reported that a correlation coefficient equal to 0.98 was obtained when the hybrid GA-SVR technique was applied to the experimental data set, and the predicted results for the model demonstrated to be consistent with the history of observed actual cyanobacteria blooms from 2006 to 2011. In another study, Liu et al. (2013) described a hybrid approach, known as real-value genetic algorithm support vector regression (RGA-SVR), to forecast aquaculture water quality in a high-density river crab culture situation. The authors concluded that the RGA-SVR forecasting method could help avoid economic losses caused by water quality problems to a certain extent. On the other hand, they reported that different types and rates of crossover and mutation should be set for different problems, since the operation of the genetic algorithm was difficult in the training process of the RGA-SVR model. Furthermore, there have also been other recent studies on prediction of effluent concentration in a wastewater treatment plant in Ulsan Metropolitan city, Korea (Guo et al. 2015), prediction of Cd(II) removal by biosorption in Iasi city, Romania (Hlihor et al. 2015), numerical modeling for algal blooms of freshwater in in Macau Main Storage Reservoir located at south of China (Lou et al. 2017), prediction of five-day biochemical oxygen demand (BOD5) parameter in the Sefidrood River basin, Iran (Noori et al. 2015), lake management to prevent eutrophication in in Chaohu Lake located in southeast China (Xu et al. 2015), predicting the sorption capacity of lead (II) ions in India (Parveen et al. 2016), and eutrophication (enrichment of a water body with nutrients) classification in Dez reservoir located in Iran (Bashiri et al. 2017) reporting the advantages and generalization ability of the SVM method based over multilayer perceptron (MLP) or neuro-fuzzy network.

SVM-Based Applications for Air Quality/Pollution Control/ Forecasting It has been reported that air quality is essential to people’s health and the environment, and accurate forecasting of the concentration of air pollutants is

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crucial to the effective monitoring of air quality (Lin et al. 2011). From this point of view, the accurate models for air pollutant prediction are needed because such models would allow forecasting and diagnosing potential compliance or noncompliance in both short- and long-term aspects (Lu and Wang 2005). In recent years, based on the emission and meteorological data collected from air-monitoring stations in different parts of the world, SVM paradigm has become popular and gained importance in forecasting problems related to air quality (Yeganeh et al. 2012). Lu and Wang (2005) examined the feasibility of applying SVM to predict air pollutant levels in advancing time series based on the monitored air pollutant database in Hong Kong downtown area. The experimental comparisons between the SVM model and the classical radial basis function (RBF) network demonstrated that the SVM was superior to the conventional RBF network in predicting air quality parameters with different time series and of better generalization performance than the RBF model. The study concluded that SVM model provided a promising alternative and advantage in time series forecast and offered several advantages (i.e., it contains fewer free parameters (or small number of learning data), and eliminates the typical drawbacks, such as over-fitting training and local minima, of conventional neural network) over the conventional feed-forward RBF neural networks. Osowski and Garanty (2007) used SVM and wavelet decomposition for daily air pollution forecasting based on the observed data of NO2, CO, SO2, and dust in the northern region of Poland. The authors decomposed the measured time series data into wavelet representation and predicted the wavelet coefficients in order to obtain the acceptable accuracy of predictions. The study concluded that application of SVM instead of classical MLP had enabled to obtain much better accuracy of forecast of the wavelet coefficients and the whole pollutant concentration at all stations. Lin et al. (2011) proposed a support vector regression with logarithm preprocessing procedure and immune algorithms (SVRLIA) model to forecast concentrations of air pollutants, namely, particulate matter (PM10), nitrogen oxide, (NOx), and nitrogen dioxide (NO2), in Taiwan. Experimental results of the study indicated that the proposed SVRLIA model provided more accurate forecasting results than the other models such as general regression neural networks (GRNN), seasonal autoregressive integrated moving average model (SARIMA), and backpropagation neural networks (BPNN). Yeganeh et al. (2012) conducted studies on an innovative method of daily air pollution prediction using combination of SVM as predictor and Partial Least Square (PLS) as a data selection tool in the forecasting of CO concentrations. The authors aimed to examine the feasibility of applying SVM and hybrid PLS-SVM models to predict air pollutant levels in short- and long-term periods based on the measured air pollutant database in Tehran. The study concluded that the proposed hybrid PLS-SVM model required lower computational time than SVM model and had better performance (more accurate and faster prediction ability) to predict air pollution in different time intervals.

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In Rio de Janeiro City (Brazil), Luna et al. (2014) analyzed the behavior of the variables (nitrogen dioxide (NO2), nitrogen monoxide (NO), nitrogen oxides (NOx), carbon monoxide (CO), ozone (O3), scalar wind speed, global solar radiation, temperature, and moisture content in the air), using the method of PCA for exploratory data analysis, and proposed forecasts of O3 levels from primary pollutants and meteorological factors, using nonlinear regression methods like ANN and SVM, from primary pollutants and meteorological factors. The study concluded that the models’ predictions and the actual observations were consistent, and PCA-ANNSVM demonstrated their robustness as useful tools for modeling and analysis of O3 concentrations in tropospheric levels. In recent study, Moazami et al. (2016) proposed a modeling approach to analyze the uncertainty of support vector regression (SVR) and FS-SVR models for the prediction of the next day CO concentration in Tehran metropolitan. They compared the results of the present study and another research on uncertainty determination of ANFIS and ANN. The results showed that the SVR had less uncertainty in CO prediction than the ANN and ANFIS models. On the other hand, they reported that the running time for uncertainty determination of SVR and FS-SVR models were more than one day, and high computational time was one of the most limitations of the implemented methodology. For this reason, the authors suggested using the faster optimization techniques for tuning the SVR parameters and applying the stop training algorithm instead of cross validation technique in order to reduce the running time for uncertainty determination of SVR model.

Illustrative Soft Computing Examples for Environmental Engineers In this section, some special illustrative soft computing examples on ANN and FL ® modeling and the respective MATLAB -based solutions are presented for environmental engineers.

Example 1 A three-layer feed-forward back propagated (FFBP) artificial neural network (ANN) model is proposed to predict the daily biogas production from a laboratory-scale anaerobic sludge bed reactor (ASBR) (Fig. 7). The input variables of the proposed ANN model are selected as follows: Total chemical oxygen demand (TCOD = X1 = S0: kg/m3), daily operating temperature (X2 = T:  C), and pH of the feeding slurry (X3 = pH). In the model structure, logarithmic sigmoid function (logsig) is used for both hidden layer and the output layer as an activation and transfer function, respectively. The learning rate is selected as η = 0.90. The model variables (X1, X2, X3, and Y) will be normalized for the scale factors of a = 0.60 and b = 0.20 by

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Fig. 7 A three-layer feedforward back propagated (FFBP) artificial neural network (ANN) proposed to predict the daily biogas production from a laboratoryscale anaerobic sludge bed reactor (ASBR)

X1 = S0

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1 4

X2 = T

2

6 5

X3 = pH

Y = Qg

OUTPUT LAYER = 1

HIDDEN LAYER = 2

3

INPUT LAYER = 3

Table 1 Ranges of model variables Model variables Total chemical oxygen demand Daily operating temperature pH of the feeding slurry Daily biogas production

Symbols and units TCOD = X1 = S0 (kg/m3) X2 = T ( C) X3 = pH Y = Qg (L/day)

Ranges 11.2–19.6 26–37 6.2–8.5 8.4–14.6

Table 2 Initial values of weights (w) and bias (y) terms for the first iteration Weights (wij and wjk) w14 = + 0.20 w24 = + 0.40 w34 = 0.50

w15 = 0.30 w25 = + 0.10 w35 = + 0.20

w46 = 0.30 w56 = 0.20

Bias terms (θ) θ4 = 0.40 θ5 = + 0.20 θ6 = + 0.10

  min using the min-max rule based on the following formulation: Oi ¼ a XXmaxi X þ b . Xmin The ranges of model variables are given in Table 1. For the first iteration, the initial values of weights (wij and wjk) and bias terms (θj and θk) are given in Table 2. ® Based on the above-noted facts, write a MATLAB script to determine the following questions: (a) At the steady-state conditions, an experimental data is given as follows: The daily biogas production is measured as Qg = 13.2 L/day for the values of S0 = 14.8 kg TCOD/m3, T = 32  C, and pH = 7.1. According to this data, evaluate the performance of the proposed ANN model in prediction of the daily biogas production by denormalizing the predicted value. (b) Update the initial values of weights and bias terms by performing a backpropagation (BP) process. After the BP operation, perform a new feed-forward process to observe the recovery in prediction of the daily biogas production (Table 3).

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Solution of Example 1 MATLAB® script %Artificial Neural Network Model %Programmed by Assoc.Prof.Dr. Kaan Yetilmezsoy clear, clc %Scale factors a = 0.60; b = 0.20; %Learning rate L = 0.90; %Ranges of variables X1min = 11.2; X1max = 19.6; X2min = 26; X2max = 37; X3min = 6.2; X3max = 8.5; Ymin = 8.4; Ymax = 14.6; %Experimental data X1 = 14.8; X2 = 32;

X3 = 7.1;

Y = 13.2

%Normalization of input variables O1 = a*((X1-X1min)/(X1max-X1min))+b; O2 = a*((X2-X2min)/(X2max-X2min))+b; O3 = a*((X3-X3min)/(X3max-X3min))+b; Yn = a*((Y-Ymin)/(Ymax-Ymin))+b %Weight values w14 = 0.20; w24 = 0.40; w25 = 0.10; w35 = 0.20; %Bias values t4 = -0.40; t5 = 0.20;

w34 = -0.50; w46 = -0.30;

w15 = -0.30; w56 = -0.20;

t6 = 0.10;

%ITERATION #1 %Computation of hidden neurons 4 and 5 I4 = (O1*w14 + O2*w24 + O3*w34) + t4; I5 = (O1*w15 + O2*w25 + O3*w35) + t5; %Computation of activation functions O4 = 1/(1+exp(-I4)); O5 = 1/(1+exp(-I5)); %Computation of hidden neuron 6 I6 = (O4*w46 + O5*w56) + t6; %Computation of transfer function O6 = 1/(1+exp(-I6)) %Denormalization of predicted value Yp1 = (((O6-b)/a)*(Ymax-Ymin))+Ymin %Error at node 6 E6 = O6*(1-O6)*(Yn-O6); E5 = O5*(1-O5)*E6*w56; E4 = O4*(1-O4)*E6*w46; %Update w14_1 = w34_1 = w25_1 = w46_1 =

of weights w14 + L*E4*O1; w34 + L*E4*O3; w25 + L*E5*O2; w46 + L*E6*O4;

w24_1 w15_1 w35_1 w56_1

= = = =

w24 w15 w35 w56

+ + + +

L*E4*O2; L*E5*O1; L*E5*O3; L*E6*O5;

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%Update of bias terms t4_1 = t4 + L*E4; t5_1 = t5 + L*E5;

t6_1 = t6 + L*E6;

%ITERATION #2 %Computation of hidden neurons 4 and 5 I4_1 = (O1*w14_1 + O2*w24_1 + O3*w34_1) + t4_1; I5_1 = (O1*w15_1 + O2*w25_1 + O3*w35_1) + t5_1; %Computation of activation functions O4_1 = 1/(1+exp(-I4_1)); O5_1 = 1/(1+exp(-I5_1)); %Computation of hidden neuron 6 I6_1 = (O4_1*w46_1 + O5_1*w56_1) + t6_1; %Computation of transfer function O6_1 = 1/(1+exp(-I6_1)) %Denormalization of predicted value Yp2 = (((O6_1-b)/a)*(Ymax-Ymin))+Ymin %Error E6_1 = E5_1 = E4_1 =

at node 6 O6_1*(1-O6_1)*(Yn-O6_1); O5_1*(1-O5_1)*E6_1*w56_1; O4_1*(1-O4_1)*E6_1*w46_1;

%Update w14_2 = w34_2 = w25_2 = w46_2 =

of weights w14_1 + L*E4_1*O1; w24_2 = w34_1 + L*E4_1*O3; w15_2 = w25_1 + L*E5_1*O2; w35_2 = w46_1 + L*E6_1*O4_1; w56_2

w24_1 + w15_1 + w35_1 + = w56_1

L*E4_1*O2; L*E5_1*O1; L*E5_1*O3; + L*E6_1*O5_1;

%Update of bias terms t4_2 = t4_1 + L*E4_1; t5_2 = t5_1 + L*E5_1; t6_2 = t6_1 + L*E6_1; %ITERATION #3 (with no updates for weights and bias terms) %Computation of hidden neurons 4 and 5 I4_2 = (O1*w14_2 + O2*w24_2 + O3*w34_2) + t4_2; I5_2 = (O1*w15_2 + O2*w25_2 + O3*w35_2) + t5_2; %Computation of activation functions O4_2 = 1/(1+exp(-I4_2)); O5_2 = 1/(1+exp(-I5_2)); %Computation of hidden neuron 6 I6_2 = (O4_2*w46_2 + O5_2*w56_2) + t6_2; %Computation of transfer function O6_2 = 1/(1+exp(-I6_2)) %Denormalization of predicted value Yp3 = (((O6_2-b)/a)*(Ymax-Ymin))+Ymin

Answer Y = 13.2000

Yn = 0.6645

O6 = 0.4659

Yp1 = 11.1475

O6_1 = 0.4824

Yp2 = 11.3186

O6_2 = 0.4977

Yp3 = 11.4759

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Table 3 Mathematical expressions for the proposed ANN model Explanation Net value transferred from input layer to hidden layer (or net value transferred from hidden layer to output layer) Output of logsig activation or transfer function Error value at the output layer (only for logsig function) Error value at the hidden layer (only for logsig function) Updated weight terms (wij and wjk) Updated bias terms (θj and θk)

Mathematical expression Ij =  wijOi + θj Ik =  wjkOj + θk Oj = 1/[1 + exp (Ij)] Ok = 1/[1 + exp (Ik)] δk = Ok(1  Ok) (Tk  Ok) δj = Oj(1  Oj)  δkwjk wnij ¼ wij þ ηδj Oi wnjk ¼ wjk þ ηδk Oj θnj ¼ θj þ ηδj θnk ¼ θk þ ηδk

Table 4 Ranges of model variables Model variables Organic loading rate Daily operating temperature Daily biogas production

Symbols and units X1 = OLR (kg COD/m3/day) X2 = T ( C) Y = Qg (L/day)

Ranges 3–12 10–40 20–100

Example 2 A fuzzy logic (FL) model is introduced to estimate the daily biogas production obtained from an experimental study. The properties of the proposed FL model are summarized below. (a) The proposed FL model is a MISO (Multiple Input Single Output) type model which is consisted of 2 inputs (X1 = OLR = Organic loading rate (kg COD/m3/ day), X2 = T = Temperature ( C)) and 1 output (Y = Qg = Daily biogas production (L/day)). (b) Trapezoidal membership functions (trapmf) with two and three levels are used, respectively, for input and output variables, and the functions are categorized as LOW, MOD, and HIGH for processing of the fuzzy rules. (c) The ranges of model variables are given in Table 4. The ranks of the membership functions for input and output variables considered in the fuzzy sets are presented in Table 5.

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Table 5 Ranks of the trapezoidal membership functions selected for the model variables Model variables X1 = OLR = organic loading rate (kg COD/m3/day) X2 = T = temperature ( C) Y = Qg = daily biogas production (L/day)

Low [4 2 4 10]

Mod –

High [4 10 14 20]

[15 5 15 35] [10 0 40 50]

– [40 50 80 90]

[15 35 45 65] [80 90 110 120]

Table 6 Rule sets for the proposed FL-based model X1 = OLR = organic loading rate (kg COD/m3/day) Low High High

Table 7 Steady-state data obtained from experimental studies

# 1 2 3

X2 = T = temperature ( C) Low High Low

Y = Qg = daily biogas production (L/day) Low High Mod

T ( C) 16 32 17

OLR (kg COD/m3/day) 5 9 8

Qg (L/day) 41 87 60

(d) The rule base is developed by taking into account the experimental results and the suggestions of the experts (Table 6). The weight factors are taken as equal (1) for each fuzzy rule. (e) The fuzzy inference system (FIS) is proposed as the Mamdani’s type, and the prod (product) method is used by the AND operator for each fuzzy rule in the FIS. The other methods implemented for implication, aggregation, and defuzzification processes are prod, max, and centroid (COG), respectively. (f) The steady-state data obtained from experimental studies are given in Table 7. ®

According to the foregoing points, write a MATLAB script to estimate the outputs of the FL-based model for each experimental data given in Table 7 and to calculate the value of the determination coefficient (R2) associated with these predictions.

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Solution of Example 2 MATLAB® script (development of the fuzzy inference system, FIS) [System] Name='fuzzy1' Type='mamdani' NumInputs=2 NumOutputs=1 NumRules=3 AndMethod='prod' OrMethod='max' ImpMethod='prod' AggMethod='max' DefuzzMethod='centroid' [Input1] Name='OLR' Range=[3 12] NumMFs=2 MF1='LOW':'trapmf',[-4 2 4 10] MF2='HIGH':'trapmf',[4 10 14 20] [Input2] Name='Temperature' Range=[10 40] NumMFs=2 MF1='LOW':'trapmf',[-15 5 15 35] MF2='HIGH':'trapmf',[15 35 45 65] [Output1] Name='Qg' Range=[20 100] NumMFs=3 MF1='LOW':'trapmf',[-10 0 40 50] MF2='MOD':'trapmf',[40 50 80 90] MF3='HIGH':'trapmf',[80 90 110 120] [Rules] 1 1, 1 (1) : 1 2 2, 3 (1) : 1 2 1, 2 (1) : 1

The numbers in the parentheses represent weights that can be applied to each rule if desired. You can specify the weights by typing in a desired number between zero and one. If you do not specify them, the weights are assumed to be unity (1). Save the file as fuzzy1.fis by selecting File > Save (or use CTRL + S).

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>> a = readfis('fuzzy1') a = name: type: andMethod: orMethod: defuzzMethod: impMethod: aggMethod: input: output: rule:

'fuzzy1' 'mamdani' 'prod' 'max' 'centroid' 'prod' 'max' [1x2 struct] [1x1 struct] [1x3 struct]

2037

>> plotfis(a)

OLR (2)

fuzzy1 (mamdani) 3 rules Qg (3)

Temperature (2)

System fuzzy1: 2 inputs, 1 outputs, 3 rules

>> showrule(a) ans = 1. If (OLR is LOW) and (Temperature is LOW) then (Qg is LOW) (1) 2. If (OLR is HIGH) and (Temperature is HIGH) then (Qg is HIGH) (1) 3. If (OLR is HIGH) and (Temperature is LOW) then (Qg is MOD) (1) >> showrule(a,1) ans = 1. If (OLR is LOW) and (Temperature is LOW) then (Qg is LOW) (1) >> showrule(a,2) ans = 2. If (OLR is HIGH) and (Temperature is HIGH) then (Qg is HIGH) (1) >> showrule(a,3) ans = 3. If (OLR is HIGH) and (Temperature is LOW) then (Qg is MOD) (1)

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>> plotmf(a,'input',1)

>> plotmf(a,'input',2) LOW

HIGH

1

0.8

0.8 Degree of members hip

Degree of members hip

LOW

1

0.6

0.4

HIGH

0.6

0.4

0.2

0.2

0

0 3

4

5

6

7

8

9

10

11

10

12

15

20

OLR

25 Temperature

>> plotmf(a,'output',1) LOW

MOD

HIGH

1

Degree of membership

0.8

0.6

0.4

0.2

0 20

30

40

50

60 Qg

70

80

90

100

30

35

40

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Show overall results (measured and predicted) in the form of a table. Calculation of determination coefficient (R2) MATLAB® script clear, clc a = readfis('fuzzy1'); data = [5 16; 9 32; 8 17]; predictions = evalfis(data, a); measured = [41; 87; 60]; %Create a table using “measured” and “predictions” vectors. %Since these vectors are in vertical format, use transpose of them. %Add a semicolon operator between them, and insert this table %in fprintf function. table = [measured';predictions']; %Location of the title words can be adjusted %as desired by pressing the SPACE key on the keyboard. %i.e., 4 Spaces for "Observed" word %and 3 Spaces for "Predicted" word. disp(' Observed Predicted') fprintf('%10.2f %11.2f \n',table) O = measured; P = predictions; Om = mean(O); Pm = mean(P); n = numel(O); %Determination of linear regression coefficient (b) %and constant term (a) %Microsof Excel uses this formulation b = (n*sum(O.*P)-sum(O)*sum(P))/((n*sum(O.^2))-((sum(O))^2)); a = (sum(P)-b*sum(O))/n; %Determination coefficient (Rsqr) Rsqr = ((b^2)*sum((O-Om).^2))/(sum((P-Pm).^2)) %Determination coefficient (Rsqr): %Same result by using the compact form Rsqr1 = ((sum((O-Om).*(P-Pm)))^2)/((sum((O-Om).^2))*(sum((P-Pm).^2)));

Answer Observed 41.00 87.00 60.00

Predicted 40.30 82.02 58.44

Rsqr = 0.9994

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Conclusion In this chapter, important applications of the soft computing-based prediction models, such as ANN, FL, ANFIS, and SVM, are specifically explored for the real-life problems of environmental engineering field. It is apparent from the literature that soft computing methods can be successfully implemented as complementary technologies in various applications of water/wastewater treatment and air quality/pollution control/forecasting. Modeling of environmental processes is very difficult, since they include biological, chemical, and physical phenomena, together. At this point, soft computing techniques serve as a modern paradigm for computing and simulating complex natural processes with basic principles of the prediction modeling using environmental data sets obtained from various real applications. Additionally, the applicability of these models is very simple, posing no need to identify nonlinear relationships between multiple variables and define the complex reactions in the environmental problems. It is worth mentioning that many investigators have compared the performance of soft computing-based techniques with conventional methods. Based on the literature review, it can be concluded that soft computing-based models have provided better results compared to traditional linear/nonlinear regression methods due to their ability to precisely discriminate the arbitrary nonlinear functional relationship between input and output data sets. Furthermore, the literature findings clearly corroborate that the soft computing methodology could describe the behavior of the complex reaction system with the range of experimental conditions adopted. Simulation based on these models can estimate the behavior of the system under different conditions. To conclude, a simulation on the basis of the soft computing model can deliver further contribution in developing a better understanding of the dynamic behavior of the environmental processes where still some phenomena cannot be clarified in all details. The encouraging results obtained from the application of the described soft computing-based approaches in modeling of water and air pollution-related problems indicate that these techniques are worth for further research and extension to other similar real-life problems from the environmental engineering field. Considering the predictive capability and robustness, of the soft computing-based methodology, these prognostic models may be integrated into full-scale water/wastewater treatment plants and mobile pollution air monitoring stations as advanced control, early warning, and decision support systems using different on-line and off-line control strategies in a cost-effective manner by means of energy and environment.

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Modelization of Trihalomethanes Formation in Drinking Water Distribution Systems in France Otmane Boudouch, C. Galey, C. Rosin, and A. Zeghnoun

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Materials and Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Study Sites . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Variation Range of the Studied Parameters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Modelization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Results . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Simplified Model . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Complete Model . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

In France, trihoalomethanes (THM) are regulated and regularly monitored at the water treatment plant and more recently in the drinking water system. THM concentrations at tap water depend on many factors like chlorine level, organic precursor’s concentrations, water temperature, residence time in the network, and O. Boudouch (*) Transdisciplinary Team of Analytical Sciences for Sustainable Development, University Sultan MoulaySlimane, BeniMella, Morocco Environmental and Agro-Industries Processes Team, University Sultan Moulay Slimane, Beni Mellal, Morocco e-mail: [email protected] C. Galey · A. Zeghnoun Agence nationale de santé publique 12 rue du Val d’Osne 94415, Saint Maurice Cedex, France C. Rosin Agence nationale de sécurité sanitaire de l’alimentation, de l’environnement et du travail (Anses) Direction de l’Evaluation des Risques, Unité évaluation des risques liés à l’eau, Maisons-Alfort Cedex, France © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_155

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presence of rechlorination stations. To predict concentrations in the water distribution system using data collected from treated water at the plant (i.e., the entrance of the distribution system), a first mathematical model was developed in 2009, from three sites supplied by surface water. Predicted concentrations produced with this model for five new sites didn’t match with observed concentrations. New efforts were then made in order to adapt this mathematical model to cover more types of water. Two formulations have been developed: a first model based on a minimum of variables and those easily available (from the French national SISEEaux database collecting all data from drinking water regulations) and a second model that includes more information about the reactivity of the organic matter with chlorine. The choice of variables and the general shape of the models were made by dividing the database into two random editions of the couples of data (75% of the data to build the models/25% to validate them). The validation of both models (simplified and complete model) was satisfactory, explaining respectively 87% and 88% of the variance, with a good capacity of generalization. The models developed herein can be used to assess THM concentrations at different points between the treated water at the plant and the consumer’s tap in a large range of French water systems supplied by surface waters.

Keywords

Drinking water · Trihalomethanes · Chlorine · Chlorination byproducts · Mathematical model · Water distribution system

Introduction Chlorination of drinking water is widely used around the world to prevent and the infectious risk conveyed by tap water. In France, its use dates from more than one century in several large cities. Since 2003, the French authorities have recommended to extend its use to all water systems regardless of the size of the population served. In 2007, more than 99% of produced drinking water were disinfected with chlorine (Davezac et al. 2008). Because of its oxidizing properties, the chlorine reacts with water organic matter to form chlorination byproducts (SPC). Nearly 600 SPC are identified to date (Richardson et al. 2007). Trihalomethanes (THM) and haloacetic Acids (HAAs) account for between 20% and 30% of the total mass of the SPC produced generally (Weisel et al. 1999). Drinking water chlorination in France is mandatory under the national legislation, while regular inspections of recreational waters are also conducted regularly (Galey et al. 2015). Water sampling is carried out at the outlet of the treatment stations having a chlorination step, and in network if the chlorine concentration in the distribution system exceeds 0.5 mg/L. The formation of SPC depends on the nature of the raw water, the treatments used to remove the organic matter and the disinfection strategy (injection points, applied doses, contact time).

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The presence of SPC poses a public health problem due to associated health risks and the large size of exposed population. Epidemiological studies indicate anassociation between exposure to SPC, generally assessed by THM measurements as part of regulatory controls, and the occurrence of bladder cancer (Villanueva et al. 2007). An association between THM exposure and colorectal cancer is also doubtful (Rahman et al. 2010; Azhar et al. 2015). Suspected effects on reproduction and development, even if they are widely studied, are still controversial (Grellier et al. 2010; Lewis et al. 2011; Hwang and Jaakkola 2012; Levallois et al. 2012). Exposure estimation is generally the weak point of epidemiological studies. THM formation evolves in the water distribution network. Several studies have showed an increase in THM concentrations by a factor of 2–6 between the treatment plant exit and periphery of the drinking water distribution system (Mouly et al. 2010). A first regression model was constructed based on three production and distribution sites of drinking water in 2009 (Mouly et al. 2009, 2010) in order to predict THM concentrations in water systems from measured output data of treatment plants, with the aim of better estimating the population exposure. Data from five other production and distribution sites were used for external validation purposes. The comparison of “2009” model predictions to the data measured on these five sites did not, however, allow to establish the validity of the model beyond the three sites considered for its construction. The aim of the study is therefore to propose two variants of a new regression modelbased on the analysis of all the data (data from the three sites used for the establishment of the “2009” model and the five sites used for its external validation), in order to have a new model with a wider range of application. A “complete model” using all the variables provided by the operators of the different sites was constructed, as well as a “simplified model” retaining a minimal subset of variables, reduced to those that are indispensable, or easily accessible and routinely produced.

Materials and Methods Study Sites Eight sites were used for model construction and validation. All these sites are fed by surface or retaining water, and comprise a complete treatment process with a filtration step on activated carbon or two-layer filtration, and an ozonization step. There is no prechlorination step in the treatment process. Final disinfection by chlorine is carried out at the exit of the treatment plant before the distribution of the water in the network. The data come from various sampling campaigns of analyzes carried out in different seasons. During each campaign, a sample was systematically carried out at the outlet of the treatment plant, downstream of the chlorination step at the treatment plant, and one

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to several samples were taken in different points of the distribution network, before or after a possible re-chlorination step. As a result, the complete data used are distributed as follows for the different sites (Table 1): Depending on the study site, several sampling points were chosen along the drinking water system. At each study site, sampling points included one point before the chlorination step, one point at the treated water at the plant (i.e., at the entrance to the drinking water network: reference point 0) and several points along the drinking water network with different residence times (Fig. 1).

Variation Range of the Studied Parameters Table 2 presents the description of water quality variables and operating variables, which may influence the formation of THM. The incorporation of these variables in the “simplified” and “complete” models is given in the table. The concentration is expressed in molar concentration (μmol.L1) because the distribution of individual THM (chloroform, dichlorobromomethane, chlorodibromomethane, bromoform) is different depending on site and because the molar mass is different for each THM. The use of all data in the same model requires translation of the concentration into molar concentration.

Table 1 Synthesis of sampling campaigns realized on the different study sites Site Site 1 Site 2 Site 3 Site 4 Site 5 Site 6 Site 7 Site 8 Total

Campaigns number 3 3 3 4 7 7 7 2

Total number of THM values in network 48 62 55 16 48 14 16 3 262

Fig. 1 Diagram of the sampling points chosen for the study

Hydraulic residence time (min–max, in hours) (11–27) (26160) (30–210) (64–160) (5–280) (19–57) (5–53) (15–53)

chlorination

Raw Water

Plant

rechlorination

Pplant P0 = reference P1 - Pi

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Table 2 List of variables tested during model construction Explanatory variable Description Water quality variables (sanitary control parameters) THM0 THM concentration at the treated water at the plant (P0) Free residual chlorine at the treated water at the plant (P0) Temp0 Water temperature at the plant (P0) TOC0 Total organic carbon at the treated water at the plant (P0) pH0 pH at treated water at the plant Operating variables Cl2inj Dose of chlorine injected into the chlorination tank CTtp Contact time in the chlorination tank at the treatment plant Water residence times between a given point in the RTi system (Pi) and the treated water at the plant (P0) RCPi Presence of one rechlorination point upstream of point i (Pi) Cl20

Br0 Absuv0

Bromide ion concentration at treated water at the plant UV absorbance at 254 nm, at the treated water at the plant

Unit

Min

Max

μg.L1 μmol. L1 mg.L1

1.3 0.01

68 0.5

0.05

1.3

7 1.1

23 4

7.2

8.5

mg.L1 Hours

1.2 0.5

6 6.9

Hours

4.5

280



C mg.L1

RCPi = 1 if rechlorination RCPi = 0 otherwise mg.L1 0.003 0.97 m1

0.003

0.08

Modelization The method used to adjust the two models is based on the random division of datain two subsamples. The first, called the training sample, is made up of 75% of the available data and it’s used to build the model. The second, called test or validation sample, consists of the remaining 25% of the data and it’s used to measure the generalization capacity of the model by comparing its predictions to the observed values. Explanatory variable is introduced as polynomial functions of 1–3 degrees in order to take into account the possible nonlinearity of the relationship between the levels of THM present in the network and the explanatory variables. Different regression models were then tested with the variables by introducing possible interactions. These models were assessed by considering: 1. R2: the coefficient of determination which determines the contribution of the tested variables in the explanation of variability of the response; RMSE: the residual mean standard error which corresponds to the error making on prediction 2. Assessment of the fit quality of the model by analyzing the graphic distribution of residues

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3. Prediction capacity of the model on data not used for its construction (validation sample), evaluated on the basis of: (a) RMSE: root mean square error (b) Relative error N25: which represents the percentage of predictions with a relative error less than 25% (c) Relative error related to uncertainty N5unc: which represents the percentage of predictions with a relative error less than 5% when uncertainty on explanatory variables is taking into account Higher values of N25% and N5unc mean that the model has a great prediction and generalization capacity. The stability of the two models selected was verified by cross-validation on eight subsamples made randomly from the starting data sample. The work was done with software R (V2.14.2).

Results Simplified Model The search for a simplified model aims to have a predictive tool, using a minimal subset of easily accessible explanatory variables (present in the SISE-EAUX French database). After exploring the relationship between THMi (THM concentration in the distribution network) and the available explanatory variables, the form of the simplified model is a polynomial form, of 1–3 degrees according to the variables, with a term of interaction between network rechlorination and water temperature (Table 3). The fitting quality and predictive performance of this model are as follows: Construction on the training sample (N = 197) R2 = 87.15% RMSE = 0.0484 Validation on the test sample (N = 65) RMSE = 0.0625 N25 = 67.7%

p < 2.2e  16 N5unc = 81.5%

The simplified model adjusts well the observed data. Indeed, the histogram and the Q-Q plot of the residues show that the distribution of the residues is close to a normal distribution. Moreover, the residual values do not exhibit any particular tendency (Fig. 2). Good predictive performances were also observed for the vast majority of the predictions of the validation sample. Predicted THM values were close to the observed ones (Fig. 3 – N25 close to 70% and N5unc greater than 80%). The four observed atypical concentrations between 0.23 and 0.37 μmole.L1, for which the simplified model predicts a value around 0.1 μmole.L–1, belong to the same site (site 7). They were all measured in the spring during the same campaign. The four sampling points are different, but have a double chlorination in the network and a residence time RT probably underestimated.

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Table 3 Variables of the simplified model obtained using the training sample: coefficients with their standard error and their degree of significance Simplified model variables Coefficient Standard deviation Constant 145.00 40.10 THM0 1.25 0.12 THM0  THM0 1.24 0.27 Cl20 0.08 0.02 RTi 0.0012 0.0004 RTi  RTi 0.000009 0.000003 RTi  RTi  RTi 0.00000003 0.00000001 pH0 55.00 15.40 pH0  pH0 6.97 1.96 pH0  pH0  pH0 0.29 0.08 TOC0 0.11 0.03 TOC0  TOC0 0.02 0.01 RCPi [0 if no, 1 if yes] 0.33 0.08 Taking into account the interaction If RCPi = 0 (without rechlorination in the network before the sampling point) Temp0 0.01 0.01 Temp0  Temp0 0.0005 0.0003 If RCPi = 1 (rechlorination in the network before the sampling point) Temp0 0.05 0.01 Temp0  Temp0 0.0018 0.0003

Pr(>|t|) 0.0004 < 0.0001 < 0.0001 < 0.0001 0.0025 0.0055 0.0011 0.0004 0.0005 0.0005 0.0004 0.0007 < 0.0001

0.2870 0.0758 < 0.0001 < 0.0001

The form of the relationships observed between levels of THMi present in the network and each explanatory variable of the simplified model allows to assess the coherence of the relations with the mechanisms involved (Fig. 4). A growing relationship is observed between the formation of THMi in the network and THM0 (THM concentration at the plant outlet), Cl20 (residual chlorine leaving the plant), RTi (residence time of water at the sample point i), and Temp0 (water temperature) when no rechlorination is used in the network. These results are in line with expectations. The bell shape of the relationship with temperature in the presence of network rechlorination is more difficult to apprehend. The relationship observed for the higher TOC0 (organic carbon of the distributed water greater than 3.5 mg.L1) or high pH (pH > 8.3) have no explanation. The campaigns associated with these conditions are limited in number and concern only a few sites.

Complete Model After exploring the relationship between THMi and the available explanatory variables, theform of the “complete model” was a polynomial form, of 1–3 degrees

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Fig. 2 Adjustment quality of the simplified model (training sample): histogram and Q-Q residue plot, residues as a function of predicted values and comparison between predicted and observed values

according to the variable, with a term of interaction between network rechlorination and water temperature (Table 4). The complete model uses the UV absorbance (at 254 nm) of water, as well as the variable R which define the chlorine consumption rate at the plant.   Cl2inj  Cl20 R¼ CTtp The fitting quality and predictive performance of this model are as follows (Figs. 5 and 6): Construction on the training sample (N = 197) R2 = 88.45% RMSE = 0.0467 Validation on the test sample (N = 65) RMSE = 0.0563 N25 = 67.7%

p < 2.2e-16 N5unc = 86.1%

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0,75 0,7 0,65

Predicted THM conc. (mmol/L)

0,6 0,55 0,5 0,45 0,4 0,35 0,3 0,25 0,2

THMi exp (μmol/L)

0.15

THMi pred (μmol/L)

0.1 0,05 0

0

0,05

0,1

0,15

0,2

0,25

0,3

0,35

0,4

0,45

0,5

0,55

0,6

0,65

0,7

0,75

Experimental THM conc. (mmol/L)

0.3 0.25 0.2

0.4 0.35 0.3 0.25 0.2

0.0 0.1 0.2 0.3 0.4

0.0

THM0 (μg/L)

0.3 0.25

8.0

pH0

1.0

0 50 100 150 200 250

8.5

0.28 0.26 0.24 0.22 0.2 0.18 0.16

RT (hours) THM (μg/L)

0.35

7.5

0.5

Cl20 (mg/L) THM (μg/L)

THM (μg/L)

THM (μg/L)

0.45 0.4 0.35 0.3 0.25 0.2 0.15

THM (μg/L)

THM (μg/L)

Fig. 3 Validation of the simplified model on the validation sample: predicted concentrations vs observed concentrations

1.0 1.5 2.0 2.5 3.0 3.5 4.0

TOC0 (mg/L)

0.38 0.36 0.34 0.32 0.3 0.28 0.26 0.24 0.22 10

15

20

Temp0 (°C)

Fig. 4 Relationships between predicted THM concentrations in the network and each explanatory variable used in the simplified model (black curves), with the confidence interval (red curves)

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Table 4 Variables of the complete model, obtained using the training sample: coefficients with their standard error and their degree of significance Complete model variables Coefficient Standard deviation Constant 139.00 40.10 THM0 1.30 0.13 THM0  THM0 1.55 0.31 Cl20 0.06 0.02 RT 0.00 0.00 RTi  RTi 0.00 0.00 RTi  RTi  RTi 0.00 0.00 pH0 53.20 15.40 pH0  pH0 6.79 1.96 pH0  pH0  pH0 0.29 0.08 TOC0 0.14 0.04 TOC0  TOC0 0.03 0.01 Rcpi [0 if non, 1 if yes] 0.26 0.07 R 0.10 0.03 RR 0.03 0.01 RRR 0.00 0.00 Absuv0 4.23 2.43 Absuv0  Absuv0 107.00 59.20 Absuv0  Absuv0  Absuv0 825.00 432.00 Taking into account the interaction If RCPi = 0 (without rechlorination in the network before the sampling point) Temp0 0.02 0.01 Temp0  Temp0 0.00 0.00 If RCPi = 1 (with rechlorination in the network before the sampling point) Temp0 0.05 0.01 Temp0  Temp0 0.00 0.00

Pr(>|t|) 0.00 < 0.0001 < 0.0001 0.00 0.01 0.02 0.01 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.03 0.17 0.08 0.07 0.06

0.04 0.01 < 0.0001 < 0.0001

The forms of relations between THM concentrations present in the network and the explanatory variables used in the complete model are similar to those observed for the simplified model (not shown).

Conclusion The model built in 2009 (Mouly et al. 2009) using data from three production and water distribution sites have not been validated on the new data collected from other sites. The quality of the water produced by the three initial sites was fairly similar, with THM concentration ranging from 10 to nearly 90 μg/L. A new modeling was then undertaken, using data from eight sites: the three sites used for the construction of the 2009 model, and the five new sites. All these sites are

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Modelization of Trihalomethanes Formation in Drinking Water. . .

-1

-0

-1

2

3

0.0

0.2

0.3

0.4

linear predictor

Histogram of residuals

Response vs. Fitted Values

Response

60 40 20

Frequency

0.1

Theoretical Quantiles

0 -0.2

-0.15-0.10-0.05 0.00 0.05 0.10 0.15

residuals -2

80

-3

-0.1

0.0 Residuals

0.1

0.2

0.5

0.6

0.5

0.6

0.0 0.1 0.2 0.3 0.4 0.5 0.6

Sample Quantiles

Resids vs. linear pred.

-0.15-0.10-0.05 0.00 0.05 0.10 0.15

Normal Q-Q Plot

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0.0

0.1

0.2

0.3

0.4

Fitted Values

Fig. 5 Adjustment quality of the complete model (training sample): histogram and Q-Q residue plot, residues as a function of predicted values and comparison between predicted and observed values

fed by surface water and include a complete water treatment process with ozonation and filtration steps. Two models were then built. The first is called “simplified.” It was built based on variables usually available from the sanitary control French basis and other indispensable variables as hydraulic residence time of water in the distribution network. The second model is called “complete.” It is constructed from all the available variables. Compared to the “Simplified” model, it includes variables that better characterize the reactivity of organic matter to chlorine as UV absorbance and the rate of chlorine consumption in the plant. The performances of these two models are very similar, with a slight improvement when moving from the simplified model to the complete model (increase of R2 from 87.15% to 88.45% and N5unc increase from 81.5% to 86.1%). The field of application of these models seems to cover surface water and French conditions water treatments, for a wide range of THM concentration levels at the outlet of the treatment plant (between 1.3 and 68 μg/L). The overall validity of the “simplified” and “complete” models leads us to propose their use to estimate THM content in a distribution network.

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0,75 0,7 0,65

Predicted THM conc. (mmol/L)

0,6 0,55 0,5 0,45 0,4 0,35 0,3 0,25 0,2 0.15

THMi exp (μmol/L)

0.1

THMi pred (μmol/L)

0,05 0 0

0,05

0,1

0,15

0,2

0,25 0,3 0,35 0,4 0,45 0,5 0,55 Experimental THM conc. (mmol/L)

0,6

0,65

0,7

0,75

Fig. 6 Validation of the complete model (validation sample): predicted concentrations vs observed concentrations

Many difficulties were met during this work in collecting entry data especially for hydraulic residence time data. Several sites initially proposed to contribute to the modeling work were not selected due to lack of exact data on relevant variables. The use of these two models to predict a THM level at a point of a water ® distribution network is possible and easy to do under Excel , providing data availability of explanatory variables. These models can be used to determine levels of THM concentrations at different points of the same network, and help identify the most critical areas, close to the regulatory standard for example. The two models were not validated on waters and treatment processes other than thoseused for their construction. It would be interesting to have other datasets of new sites, in particular with underground water, in order to verify their ability to be generalized.

References Azhar S, Sumayya S, Majid M, Mirza MH, Haider AK (2015) Multipathways human health risk assessment of trihalomethane exposure through drinking water. Ecotoxicol Environ Saf 116:129–136 Davezac H, Grandguillot G, Robin A, Saout C (2008) L’eau potable en France 2005–2006, p 63

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Galey C, Zeghnoun A, Boudouch O, Beaudeau P, Rosin C (2015) Modélisation de la formation des trihalométhanes dans les réseaux de distribution d’eau destinés à la consommation humaine en France. Tech Sci Méthodes 6(1):20–31 Grellier J, Bennett J, Patelarou E, Smith RB, Toledano MB, Rushton L et al (2010) Exposure to disinfection by-products, fetal growth, and prematurity: a systematic review and meta-analysis. Epidemiology 21(3):300–313 Hwang BF, Jaakkola JJK (2012) Risk of stillbirth in the relation to water disinfection by-products: a population-based case-control study in Taiwan. PLoS One 7(3):e33949 Levallois P, Gingras S, Marcoux S, Legay C, Catto C, Rodriguez M et al (2012) Maternal exposure to drinking-water chlorination by-products and small-for-gestational-age neonates. Epidemiology 23(2):267–276 Lewis C, Hoggatt KJ, Ritz B (2011) The impact of different causal models on estimated effects of disinfection by-products on preterm birth. Environ Res 111(3):371–376 Mouly D, Joulin E, Rosin C, Beaudeau P, Zeghnoun A, Olszewski OA et al (2009) Les sousproduits de chloration dans l’eau destinée à la consommation humaine en France. Campagnes d’analyses dans quatre systèmes de distribution d’eau et modélisation de l’évolution des trihalométhanes. Institut de veille sanitaire, Saint-Maurice, p 73 Mouly D, Joulin E, Rosin C, Beaudeau P, Zeghnoun A, Olszewski OA et al (2010) Variations in trihalomethane levels in three French water distribution systems and the development of a predictive model. Water Res 44(18):5168–5179 Rahman MB, Driscoll T, Cowie C, Armstrong BK (2010) Disinfection by-products in drinking water and colorectal cancer: a meta-analysis. Int J Epidemiol 39:733–745 Richardson SD, Plewa MJ, Wagner ED, Schoeny R, Demarini DM (2007) Occurrence, genotoxicity, and carcinogenicity of regulated and emerging disinfection by-products in drinking water: a review and roadmap for research. Mutat Res 636(1-3):178–242 Villanueva CM, Cantor KP, Grimalt JO, Malats N, Silverman D, Tardon A et al (2007) Bladder cancer and exposure to water disinfection by-products through ingestion, bathing, showering, and swimming in pools. Am J Epidemiol 165(2):148–156 Weisel CP, Kim H, Haltmeier P, Klotz JB (1999) Exposure estimates to disinfection by-products of chlorinated drinking water. Environ Health Perspect 107(2):103–110

Part X Environmental Nanotechnology: Management of Nano-waste (Nanomaterials)

Nano-wastes and the Environment: Potential Challenges and Opportunities of Nano-waste Management Paradigm for Greener Nanotechnologies

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Sherif A. Younis, Esraa M. El-Fawal, and Philippe Serp

Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Nanomaterials: Definition and Current Markets . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Nano-engineering Techniques . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Nano-waste: Definition of Terms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Toxicity Concept of Nano-waste to the Environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Modeling NM-Risk Profile Relevant to Their End-of-Life (EOL) Cycle and Fate Behavior into the Environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Classification of Nano-waste Risk Hazards . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Analytical Tools for Nano-waste Detection and Monitoring . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Challenge in Nano-waste Managing Protocols . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Ongoing Treatment Opportunities of Nano-waste Products . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Potential Opportunities for the Recovery and Reuse of Nano-waste Technologies . . . . . . . Nano-specific Regulations and Legislation Versus Voluntary Environmental Programs . . . . . Nanotechnology Vision for the Next Decade: Green “Benign” Innovation for Sustainable Development . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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S. A. Younis (*) Analysis and Evaluation Department, Egyptian Petroleum Research Institute, Nasr City, Cairo, Egypt Laboratoire de Chimie de Coordination UPR CNRS 8241, composante ENSIACET, Université de Toulouse, UPS-INP-LCC, Toulouse Cedex 4, France e-mail: [email protected]; [email protected] E. M. El-Fawal Analysis and Evaluation Department, Egyptian Petroleum Research Institute, Nasr City, Cairo, Egypt e-mail: [email protected] P. Serp Laboratoire de Chimie de Coordination UPR CNRS 8241, composante ENSIACET, Université de Toulouse, UPS-INP-LCC, Toulouse Cedex 4, France e-mail: [email protected] © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_53

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Summary, Recommendation, and Future Remark . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2125 Cross-References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2126 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2127

Abstract

Although nanoscience has been positioned as the source of the next revolution of novel technology benefits, progress moving nanoscience from the laboratory to develop planet-friendly nanotechnologies was slow over the last decade. This is because of knowledge gaps and a number of major uncertainties in regard to understanding the new nano-hazard behavior and fat in the environment as well as the lack of specific policies to identify, monitor, and manage associated nano-risk due to the unknown environmental transformation processes that affect nano-toxicity exposure mechanisms. Recognition of these challenges will require a new eco-friendly knowledge and guidelines for the legalization of risk-associated nanomaterials, for effective design greener “benign” nano-innovation-based technologies wherever possible in the immediate future. In the present chapter, one has to consider the combination of different life cycle assessment (LCA) concepts with the evolving knowledge from nano-waste risk assessment and their end-of-life cycle thinking to design nano-products at zero-waste level. The data provide a nano-waste basis and legalization opportunities for abatement of nano-pollutions by the industry and regulators. In this context, this chapter can be useful to policymakers in developing the framework of perspective and post-active for safe nanoproduction to assess the environmental sustainability performance for future development into the largest industrial and economic sectors. Keywords

Nanomaterial · Characterization · Nano-waste · Environment risk assessment · Management paradigm · Legalization and regulation protocols · Treatment and recycle processes · Green innovation

Introduction Nanotechnologies and nanoscience are an emerging field with a vast potency to bring benefits for many areas of research and for many applications. What is really new about nanoscience is that the advances in material science and characterization tools allow for monitoring materials at the nanometer (nm = 109 m) scale, which provided unprecedented opportunities to understand interaction at atomic scale. Although study on constructed nano-materials (NMs) can be traced back for centuries, the current fever of nanotechnology is at least partly driven by the continued shrinking of semiconductor devices dimensions by a factor of 2 every 18 months since 1950 as illustrated in Fig. 1 by the well-known Moore’s law predicted in 1965 (Pokropivny et al. 2007).

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1st integrated circuit

50 nm

1950

1960

1M transistors per chip

Moore’s Law Trend Line 1 μm

1st transistor

Transistor Size

1 cm

1–5 nm 1970

1980

1990

2000

2010

2020

Fig. 1 “Moore’s law” plot of transistor size versus year as predicted since 1965

The novelty and revolutionary character of nanotechnologies has resulted in dramatic growth of nanomaterial (NM) production, as demonstrated by NANOWERK catalogues that reported fast development of fabricated NMs from 2600 types in 2015 (Nanowerk 2015) to more than 4000 types of NMs on the markets in 2017 (Nanowerk 2017). Comparatively, marketplace based on Nanotechnology Products Database (NPD) at 2017 estimated that there are now more than 6970 nano-products based on the developed NMs produced by 1378 companies, from 52 countries worldwide (Engelmann et al. 2017; Khan et al. 2017). In parallel, in 2010, it was estimated that the worldwide production rate of manufactured NMs varied from 268 to 318 thousand metric tons, and it has since been increasing at a rate of about 25% per year (Mrowiec 2016). Market survey forecasts claimed that although nanotechnology is estimated to be a multi-trillion dollar industry in the next decade, mainly two challenges may raise that have slowed the move forward of nanoscience from the research to industrialization: (1) the technical, i.e., poor characterization of the new nano-hazards and their safety, and (2) the regulatory, i.e., lack of adequate legalization policies to manage the new nano-risks (Musee 2011; Schulte et al. 2013; Johnson 2016). These issues originated from a well-known fact: the unregulated fast production of novel NM-based products (such as nanoelectronics, molecular assemblies, tissue engineering, biomedicine, and nanocomposites), without having a scientific certainty about risk assessment or specific safety regulatory framework. So, the developed nanotechnologies could pose new forms of challenges to the current waste “nano-waste or nano-pollution” management paradigm either by rendering them inadequate or inappropriate (Musee 2011). This raises a fundamental technological question: “is the new nano-products likely to cause a disruptive approach in terms of technologies required for managing and treating waste streams containing NMs?” In this case, there is an urgent need for new scientific research plans dealing with the toxicology effects of existed and newly

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developed forms of NMs or nanotechnologies. Therefore, it is important to take a life cycle and risk assessment approach to evaluate the potential environmental as well as human health benefits or drawbacks (both occupational and end-use “nano-wastes”) at each stage of a nano-enabled product (Fig. 2), in order to manage, produce, and deploy novel and safe nanotechnologies into the industrial sector, which become key R&D priorities in Europe and North America (Bystrzejewska-Piotrowska et al. 2009; Part et al. 2015b; Bhatia 2016; Di Sia 2017; Engelmann et al. 2017; Iavicoli et al. 2017; Purohit et al. 2017). For assessing the sustainable production of novel nanotechnologies, novel discoveries dealing with nano-waste management paradigms can provide an unusual opportunity to design the inherent interconnection between the composition, structure, properties, technology, and applications of nano materials, in order to take technical decisions. This should help, for example, to classify novel NMs as possible “safe-inventions” for future “greener” nanotechnology in order to implement a secure development of nanoscience. In this chapter, we will use the latest science, engineering, and risk assessment and management policy knowledge to provide an overview of opportunities for the development effective nano-hazardous and nano-waste management paradigms,

Impacts (economic, society, environment, human health)

Novel-nanotechnology Technical and Technology Decision

Life-Cycle Analysis Artificial Commercialization

Technology Production

Novel Nano-products

Characteristic Properties

Unique Properties

nc Sc ie ry ist em Ch

Chemical Constituents

Materials Science

e nc ie Sc al ic ys Ph

e

Application and Consumption

Physical Characterization

kt Ris men s s e ss

A

& cs ysi istry h P em Ch

NMs Manufacture

Fig. 2 Fundamental trade of advanced materials science to implement secure development

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which could lead to the development of the “green nanoscience” concept for greener nanotechnologies in future applications and markets. The chapter is written in a way to provide illustrations on key questions such as: What is the difference in disposal macro- and nanoscale wastes? How should the final nano-products residues be disposed and managed? What are the legalization protocols currently used to regulate nano-waste management programs? What is being done currently to address the associated risk hazards accompanying development, application, and disposal of nano-products? How should benign and safe nano-products be constructed for greener nanotechnologies?

Nanomaterials: Definition and Current Markets Since the steels of the nineteenth century to date, there has been a very rapid development in the implementation of new nanomaterials with the ability to control and improve their structure (Gajewicz et al. 2012; Adams and Barbante 2013). Definitions of nanoscience and/or nanotechnology refer to the manipulation of individual atoms and molecules to manufacture nanoscale dimension materials below the submicroscopic level (approximately from 1 up to 100 nanometers: according to the National Nanotechnology Initiative (NNI) in the USA). The International Organization for Standardization (ISO) classified NMs into three main groups: nanoparticles (all three dimensions (3D) between 1 and 100 nm), nanoplates (two dimensions (2D) between 1 and 100 nm), and nanofibers (one dimension (1D) between 1 and 100 nm) (Vittori Antisari et al. 2013; Bhatia 2016). Carbonbased NMs are material composed mostly of carbon and commonly taking the form of a hollow spheres and ellipsoids shapes (referred as fullerenes C60) or cylindrical tube shapes called nanotubes (0.5–3 nm diameters and 20–1000 nm in length) or irregular shapes (like carbon black), which is by far the most important carbon nanomaterial produced industrially for many years. The metal-based NMs are, in general, fabricated mainly from group II to VI and III to V elements in the periodic table, metal oxides (size 50%). Analysis of a block of flats constructed by assembling prefabricated elements and demolished through controlled blasting. Despite 6 impact categories were considered, the results were mainly referred to energy and global warming potential. Although the use phase was found to be the most important stage, significant benefits can be obtained paying attention to other phases. For instance, a proper strategy for managing wastes

Summary

Block of flats

Content

Process

Approach

1 m2 of net floor area over a year

FU

40 years

Period

Materials, construction, use, maintenance and end-oflife

Phases

Gross energy requirement, global warming potential, ozone depletion, acidification, eutrophication, and photochemical ozone creation potential

Impacts

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(Monahan and Powell 2011)

England

might lead to a net environmental gain of 29% and 18% in terms of energy and greenhouse emissions, respectively. Comparison of a Modern Method of Construction (MMC) combining an offsite modular timber frame with a larch cladding with two alternative scenarios consisting of a brick cladding and a masonry cavity wall. The results suggested that a reduction of 34% in embodied carbon might be obtained by replacing traditional construction techniques with modern practices, including up to 24% savings through the substitution of brick cladding by larch cladding. MMC timber frame larch cladding, MMC timber frame brick cladding and conventional masonry cavity wall Process

1 m2 of usable floor

– Materials and construction

Primary energy and CO2

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Country USA

USA

Reference (Horvath and Hendrickson 1998)

(Zapata and Gambatese 2005)

Summary Evaluation of two different bridge girders (steel and steel reinforce concrete) using the Economic InputOutput Life- Cycle Assessment (EIO-LCA) approach. After pointing out the uncertainties of the used data, the concrete bridge was concluded to be the best option if both materials and construction phase are taking into account. However, these results might change if the recycling and reuse rates of steel are considered. Appraisal of a Continuously Reinforced Concrete Pavement (CRCP) and an asphalt pavement. CCRCP Asphalt pavement and CRCP

Content Steel bridge girders and steel- reinforced concrete bridge girders

Process

Approach IO

1 km, 2 lane highway

FU Whole bridge

Phases Materials, construction and maintenance (painting the steel bridge)

Materials and construction

Period 80 years

–

Energy

Impacts Chemical emissions, hazardous waste generation and conventional air pollutant emissions

Table 3 Overview of the main research articles related to the Life-Cycle Assessment (LCA) and Life-Cycle Inventory (LCI) of infrastructures

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(Keoleian et al. 2005)

USA

was found to consume approximately 21% more energy than asphalt pavements during the materials and construction phases, with the production of cement and the mixing and drying of aggregates being the main cause of energy consumption. A significant reduction in these figures can be achieved by either replacing cement by fly ash or changing the storage of aggregates. Assessment of two new bridge deck technologies consisting of: (1) conventional steel expansion joints and (2) engineered cementitious composites (ECC) link slabs. The ECC system demonstrated to have a better Bridge containing steel expansion joints and bridge using ECC link slabs

Process

Whole overpass (0.1 mi length, 4 lanes width and 9 in. height)

60 years

Materials, construction, maintenance and end-oflife

Life-Cycle Assessment of Construction Materials: Analysis of. . . (continued)

Energy and material resource consumption, air and water pollutant emissions, and solid waste generation

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Country

England

Reference

(Huang et al. 2009)

Table 3 (continued)

environmental behavior due to an extension in the deck lifetime and subsequent reduced maintenance. Development of a new LCA model for pavements, which enables considering recycled materials. The proposed approach was applied to the case study of the LHR Terminal-5 access road, proving that the production of asphalt mix accounted for 62% of the total energy consumed, which suggested that using cold asphalt mixtures might be beneficial to reduce these impacts.

Summary

Conventional asphalt pavement and asphalt pavement containing waste glass, incinerator bottom ash and RAP

Content

Process

Approach

30,000 m2

FU

–

Period

Materials and construction

Phases

Energy and CO2

Impacts

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reports. To further enlighten the usefulness and applicability of this tool, these case studies involved the LCA of raw materials, as well as their environmental impacts when performing as components in the two fundamental facilities providing services to the society: infrastructure and buildings.

Production of Clinker with Fossil and Municipal Solid WasteDerived Fuel Concrete is the second material with highest consumption rate worldwide, only after water (Sedgwick 1991). Its main component is cement, whose production involves synthesizing a mixture of clay and limestone in a kiln to produce clinker. This process requires an important amount of energy to ensure that the temperature in the kiln is about 2000  C. The consumption of cement in the world is forecasted to reach 3.4 billion of tons by 2020, including the corresponding increase in energy usage, raw materials, and generation of pollutants (UNEP 2011). For these reasons, cement usage implies huge environmental impacts, which highlights the relevance of managing it properly. On a different note, population growth and changes in consumption habits are causing an increase in the generation of Municipal Solid Waste (MSW), which requires integral management systems to consider treatment and disposal alternatives capable of ensuring an adequate response in terms of economic feasibility, environmental efficiency, and social acceptance. In this sense, the cement industry might have a key role in MSW management (MSWM), since it enables co-processing high calorific value fractions, i.e., Inorganic Fraction of MSW (IFMSW), using them as Refuse Derived Fuel (RDF). In fact, there are several studies recommending the use of IFMSW as an alternative fuel in cement kilns (Genon and Brizio 2008; Mokrzycki et al. 2003; Strazza et al. 2011), due to their better environmental performance as a result of its suitability for dealing with high temperatures. However, this condition must be evaluated through the holistic, objective, and systematic consideration of environmental impacts in the particular context of Mexico (Güereca et al. 2015).

Methodology An investigation was conducted to develop a comparative LCA of two alternative fuels for clinker production: petroleum coke and a fuel mixture consisting of petroleum coke and IFMSW. Hence, the following scenarios were defined: REFERENCE, which considers 100% petroleum coke as fuel, and IFMSW, which assumes a combination of 80% petroleum coke and 20% IFMSW. Both scenarios were characterized using real data acquired from the cement plant of CEMEX in Tepeaca and the MSWM services of Mexico City, which currently provides IFMSW to Tepeaca. The Functional Unit (FU) for the LCA was defined as 1 ton of clinker, taking into account every consumption related to the raw materials, water, energy, atmospheric emissions, discharges to water, and waste generation associated with each of the unit processes in the life cycle of clinker (quarry, grinding, homogenization, and kiln), fuels and transportation of materials, as shown in Table 4. The composition of the MSW used was 32% plastics, 50% paper and board, 10% textiles, and 8% timber. The

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Table 4 Unit processes considered in the Life-Cycle Assessment (LCA) Process Raw-meal Kiln Refinery Tra-coke Electricity Fuel Refractory RDF Tra-RDF Landfill

Description Quarry and crushing for the production of raw meal Kiln in the cement plant in Tepeaca Production of petroleum coke at a refinery Transport of petroleum coke by train (national and imported) and ship (imported) Production and use of electricity in the cement plant in Tepeaca Production, transport, and use of fuel oil in relation to the cement plant in Tepeaca Production, transport, and use of refractory material in kiln Collection, transfer, selection, and compaction of waste to produce RDF Transport of RDF from the sorting plant to the cement plant in Tepeaca Collection, transfer, selection, and transport of waste from the sorting plant to landfill

environmental impact categories considered in the assessment were: (1) Abiotic Depletion (MJ of fossil fuels), (2) Acidification (kg SO2 eq), (3) Eutrophication (kg PO4 eq), (4) Global Warming Potential (kg CO2 eq), (5) Ozone Layer Depletion (kg CFC-11 eq), (6) Photochemical Oxidation (kg C2H4), and (7) Terrestrial Toxicity (kg 1.4-DB eq).

Results and Discussion Figure 3 illustrates the environmental impacts derived from both scenarios expressed as a percentage, which were determined from the results shown in Table 5. The results are very revealing in general terms, since the IFMSW scenario exhibits a better environmental performance with respect to all the impact categories analyzed, which is mainly due to the decreased coke consumption and the subsequent mitigation of the impacts stemming from its manufacturing. Another relevant overall inference is related to the decrease in the amount of MSW that is sent to landfill thanks to its use as a co-fuel. By virtue of the importance of Climate Change and human toxicity, the results obtained for these two categories are discussed more in detail in next paragraphs. Regarding global warming, the greenhouse gas (GHG) emissions are found to reach 425 kg of Carbon Dioxide Equivalent (CO2 eq) per ton of manufactured clinker in the REFERENCE scenario, with the kiln being responsible for 77% of the emissions generated, followed by the coke manufacturing process at the refinery (11%). The IFMSW scenarios amounts to 407 kg of CO2 eq per ton, which is consistent with the results achieved in previous similar studies (Genon and Brizio 2008). As for human toxicity, the REFERENCE scenario generated 132 kg 1.4Dichlorobenzene Equivalent (DBe), while the IFMSW alternative yielded a value of 72 kg 1.4-DBe. This difference is caused by the transfer of heavy metals to the clinker matrix in the co-processing of MSW, avoiding their release to the atmosphere (Genon and Brizio 2008). Although other studies have reported that mercury (Gendebien et al. 2003) and chlorine (Genon and Brizio 2008) can produce a slightly increase in their emissions when using MSW as fuel, this consideration was not included in this research.

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Fig. 3 Contribution of each unit process to the environmental impacts produced in the scenarios and categories under analysis

Table 5 Environmental impacts for the fuel scenarios and categories considered ID 1 2 3 4 5 6 7

Impact category Abiotic depletion (fossil fuels) Acidification Eutrophication Global warming (GWP100a) Ozone layer depletion (ODP) Photochemical oxidation Terrestrial ecotoxicity

Unit MJ kg SO2 eq kg PO4 eq kg CO2 eq kg CFC-11 eq kg C2H4 eq kg 1.4-DB eq

REFERENCE 1.101 2.390 0.399 966.036 5.37E-05 0.097 0.090

IFMSW 0.937 1.676 0.193 931.576 4.38E-05 0.088 0.055

Dioxins were other main concern in the co-processing due to their influence in the generation of impacts related to toxicity. In this respect, no correlation was found between the formation of dioxins and the use of MSW as fuel, which was consistent with other specific studies carried out to assess the generation of dioxins as a result of using MSW a RDF in cement kilns (SINTEF 2006).

Conclusion In summary, the results obtained in this research demonstrated that substituting 20% of petroleum coke by MSW with high calorific value involves a decrease in 18 kg of CO2 eq and 60 kg 1.4-DBe per ton of manufactured clinker. Consequently, this replacement is an environmentally favorable alternative, mainly due to the following reasons:

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• A large amount of MSW is not only not sent to landfill, but reused as RDF, with the corresponding mitigation of environmental impacts. • The emissions produced by some compounds with harmful environmental effects decrease, due to the temperatures and time reached in cement kilns. • The amount of petroleum coke used is reduced, resulting in a decrease in the environmental impacts associated with the refining process. • The depletion of fossil fuels is attenuated, because of the substation of petroleum coke by MSW. Special attention must be paid to the quality of the MSW incorporated into the co-generation process, in order to control the emission of heavy metals. The quantification of emissions under a life-cycle approach supports the process of decision-making from a holistic and informed perspective, which in this case enables reusing the fraction of MSW with high calorific value as an alternative fuel in cement kilns.

Comparative Assessment of Asphalt and Concrete Pavements The construction, operation, and maintenance of roads involve large amounts of materials and energy, which results in important economic and environmental impacts throughout their life cycle. There is an increasing interest in Mexico in analyzing such impacts and adopt measures to minimize them as much as possible, as well as in improving the capacity and performance of pavements. Pavements are mainly classified into flexible and rigid, with asphalt and concrete being the most representative materials for each of them, respectively. Mexico has 95,000 km of paved roads, of which only 5% are made of concrete. Both types of pavements are different in several aspects, e.g., the thickness of their base and subbase layers, as well as the requirements in terms of rehabilitation and maintenance associated with each of them. The service life of asphalt pavement is about 20 years, while that of concrete pavement might reach up to 50 years (McLawhorn 2004). Deciding which type of pavement to use is a crucial aspect when developing or improving road infrastructures. However, this decision must not only be made according to a traditional approach strongly based on investment and operation costs, but should also consider the minimization of the environmental impacts caused by these infrastructures. LCA was used to compare the aforementioned two main types of pavements in an objective and systematic manner, in order to support an environmentally respectful decision-making process. In this context, the aim of this study was the evaluation of the environmental impacts of asphalt and concrete pavements under a life-cycle approach, taking into consideration the conditions of the Mexico-Queretaro highway as a real case study. The assessment was carried out according to the ISO 14044:2006 standard (ISO 2006b) and considered 18 impact categories, which were modeled using the World ReciPe v1.07 method (Güereca Hernández et al. 2014).

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Methodology Both types of pavements were designed for a life service of 20 years, in line with the recommendations established by the AASHTO Guide for design of Pavement Structures (AASHTO 2006) and the methodology Dispav 5 (version 3.0), developed by the Engineering Institute of the UNAM (Corro et al. 2014). The FU to assess these two alternatives, namely, hydraulic cement concrete pavement and asphalt concrete pavement, was 1 km of wearing course in a road linear section. The width of such road section was 21 m, which represented six lanes divided into two driving directions. The processes analyzed in the case of the concrete pavement included the following steps: extraction of raw materials for the production of cement, manufacturing of Compound Portland Cement CPC40 (quarry, crushing, kiln, grinding and homogenization), transportation of materials, construction of the pavement (manufacturing of the mixture and installation of the wearing course), maintenance activities in the seventh and fifteenth years, and disassembly and final disposal in the year 20. Regarding the asphalt pavement alternative, its characterization took into account the next processes: extraction of raw materials for the production of asphalt from an asphalt refinery, transportation of materials, construction of the pavement (manufacturing and installation of the asphalt mixture), maintenance, and final disposal. The maintenance plan for this type of pavement included two sealcoats every 3 years, milling in the year 9, sealcoating in the year 12, milling and replenishment in the year 15, and sealcoating in the year 18. The impact categories selected for comparing both types of pavements were the following: (1) Global Warming Potential (kg CO2 eq), (2) Ozone Depletion (kg CFC-11 eq), (3) Terrestrial Acidification (kg SO2 eq), (4) Freshwater Eutrophication (kg P eq), (5) Marine Eutrophication (kg N eq), (6) Human Toxicity (kg 1.4DB eq), (7) Photochemical Oxidant Formation (kg NMVOC), (8) Particulate Matter Formation (kg PM10 eq), (9) Terrestrial Ecotoxicity (kg 1.4-DB eq), (10) Freshwater Ecotoxicity (kg 1.4-DB eq), (11) Marine Ecotoxicity (kg 1.4-DB eq), (12) Ionizing Radiation (kg U235 eq), (13) Agricultural Land Occupation (m2a), (15) Natural Land Transformation (m2), (16) Water Depletion (m3), (17) Metal Depletion (kg Fe eq), and (18) Fossil Depletion (Kg Oil eq). Results and Discussion Figure 4 represents the environmental impacts produced by the two scenarios under analysis, expressed as a percentage. These results indicated that the concrete-based alternative outperformed the asphalt pavement in 17 of the 18 impact categories under consideration. This was mainly due to the impacts generated in the refinery to produce asphalt, as well as because of the more intensive maintenance strategy required for ensuring a proper behavior of the asphalt pavement solution. Concrete pavement only resulted in greater impacts than asphalt pavement in the category of metal reduction, as a consequence of using steel in construction stage of this alternative. In terms of Climate Change, the asphalt alternative resulted in greater impacts mainly due to GHG emissions generated during the manufacturing of raw materials

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Fig. 4 Environmental impacts per category produced by the two alternatives under study: asphalt and pavement concrete

and the need for disposing the pavement in hazardous waste management facilities. In particular, the impact of the asphalt pavement in relation to this category was 12,467 tons of CO2 eq per km of road, of which 55% were caused by the production of asphalt. In this sense, it is worth highlighting that this value includes the manufacturing of asphalt used both in the construction and maintenance phases. The Climate Change impact determined in this study for asphalt pavement was greater than those reported by Vidal et al. (2013) and Butt et al. (2014), but kept within the range provided by Noshadravan et al. (2013). One of the main reasons behind these differences was related to the design characteristics of the pavement, which required 12 cm layer in the Mexico-Queretano road, due to their high Annual Average Daily Traffic (AADT) of 28,098 vehicles. Instead, the case study presented by Vidal et al. (2013) considered an AADT of 1,000 vehicles, which corresponded to an 8 cm thick asphalt wearing course. As for the concrete pavement, its whole Climate Change impact amounted to 5374 tons of CO2 eq per km of road, a value which was in the range of 440 to 6,670 tons of CO2 eq reported by Loijos et al. (2013) for 12 types of roads paved with hydraulic cement concrete in the United States. Furthermore, the impact of the manufacturing process of Compound Portland Cement CPC40 produced 1,920 tons of CO2 eq per km, which was consistent with the results achieved by Loijos et al. (2013) too. Another process with relevant influence on both alternatives was the transportation of materials, which was due to long distance considered (1,000 km round trip) for the final disposal of the pavement at the end of its service life. This value was estimated based on the scarce number of facilities devoted to this purpose in Mexico.

Conclusion The life-cycle environmental comparison between the two main types of pavement used worldwide, made of concrete and asphalt, respectively, revealed a clear

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supremacy of the former, which presented lower impacts than the latter in 17 of the 18 impact categories considered. The main impacts produced by the concrete pavement solution were associated with energy consumptions and the generation of emissions to the atmosphere due to the manufacturing process of cement, as well as because of the steel requirement for construction and the amount of fuel consumed to transport the pavement to landfill once it reached its end of life. Similarly, the environmental impacts produced throughout the life cycle of the asphalt pavement alternative were mainly determined by the refinery process to produce asphalt, as well as by fuel consumption required for the final disposal of the pavement after its disassembling.

Selection of Materials for the CASA UNAM System The CASA UNAM project was conceived to produce a positive impact on the urban area of Mexico City and its Metropolitan area (ZMVM), populated by 20.1 million inhabitants, whose particular context requires flexible solutions in the form of smart building systems. Thus, the core principles behind the philosophy of the CASA UNAM project, aimed at developing an energy-efficient and sustainable house prototype, are as follows: • Reducing the cost of land shortening the distance between people and places of interest. • Generating a growing capacity to enable overcoming adverse situations and avoiding persistent damage. • Converting implementation sites into healthy and diverse environments representing genuine expressions of life. • Setting out reciprocal and interdependent relationships. Consequently, the CASA UNAM system was designed to be implemented in residual zones in the city, such as empty lots, interstitial spaces in infrastructures, in-between party walls and existing buildings, as an extension in rooftops or terraces (Team Mexico UNAM 2014).

Methodology The selection of materials for the CASA UNAM system was carried out with the reduction of weight in mind as a main goal, since this factor was expected to be determinant for the reduction of environmental impacts. A decrease in weight can result in an increase in the efficiency of transportation in terms of fuel consumption, as well as an underlying reduction of the amount of raw materials and energy required for building the house. A conventional building envelope might contain up to 20 different materials, which hinders reaching a remarkable recyclability ratio at the end of life of the house. The transformation of the traditional envelope into a graduated interface,

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according to the bioclimatic strategy adopted for the CASA UNAM system, was aimed at leading to a recyclability rate of 85%, which resembles common figures for the car industry. Overall, the CASA UNAM system is based on a dry construction process that maximizes the efficiency of the assembly and disassembly tasks, thanks to the high percentage of prefabricated elements used. The selection of materials, which was undertaken with the support of specialized literature and the Leadership in Energy and Environmental Design (LEED) international standards, included the following elements: gypsum, steel, aluminum, fiberglass, glass, textiles, plywood, wood board, rock wool, plastic, and wood. The environmental sustainability of these materials was evaluated through a LCA, which compared the CASA UNAM system with a conventional Mexican house based on the use of the ReCiPe 2008 method and the consideration of the following impact categories: (1) Global Warming Potential (kg CO2 eq), (2) Ozone Depletion (kg CFC-11 eq), (3) Terrestrial Acidification (kg SO2 eq), (4) Freshwater Eutrophication (kg P eq), (5) Marine Eutrophication (kg N eq), (6) Human Toxicity (kg 1.4DB eq), (7) Photochemical Oxidant Formation (kg NMVOC), (8) Particulate Matter Formation (kg PM10 eq), (9) Terrestrial Ecotoxicity (kg 1.4-DB eq), (10) Freshwater Ecotoxicity (kg 1.4-DB eq), (11) Marine Ecotoxicity (kg 1.4-DB eq), (12) Ionizing Radiation (kg U235 eq), (13) Agricultural Land Occupation (m2a), (15) Natural Land Transformation (m2), (16) Water Depletion (m3), (17) Metal Depletion (kg Fe eq), and (18) Fossil Depletion (Kg Oil eq).

Results and Discussion Table 6 lists the materials used for the construction of the CASA UNAM system, arranged according to the assemblies and subassemblies in which they were included. In comparison with conventional Mexican houses, the CASA UNAM system highlights by the replacement of materials with high environmental impacts by recycled products and natural components. For instance, conventional house foundations in Mexico are made of concrete, while recycled co-polymer polypropylene piles were chosen for the CASA UNAM system. Another similar example may be found in use of gypsum, which produces lower environmental impacts and is more recyclable than cement in Mexico. Finally, the floor of Mexican houses, traditionally built with ceramic materials or cement, has been substituted by wood, which is less environmentally harmful and also provides bioclimatic properties. The life-cycle processing of the materials considered for the CASA UNAM system produced the environmental impacts depicted in Fig. 5, which represents them arranged according to the assemblies included in Table 6. The results were obtained for each impact category, considering a 100 year time horizon for Climate Change. The higher environmental impacts in Fig. 5a correspond in almost all categories to the subassemblies related to the superstructure and masonry of the system, mainly due to the emissions derived from the production of steel. As for the values illustrated in Fig. 5b, their interpretation suggests that the elements belonging to the assembly of partitioning, backing wall, ceiling, and interior joineries caused the highest impacts, as a result of the glass wool used for heat insulation.

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Fig. 5 Environmental impacts per category and assembly produced by the CASA UNAM system (a) Assemblies from 1 to 4 (b) Assemblies from 5 to 8

Most of the emissions generated were a consequence of the manufacture and transport of materials. The textile materials used in the tasks related to the seventh assembly (see Table 6) involved the largest emissions to the atmosphere, due to their transport by ship from Europe. The second and third most polluting materials were found to be in the steel and clay included in the third and eight assemblies listed in Table 6, respectively. In this case, the main emissions produced by these materials stemmed from their manufacturing processes.

Conclusion As an alternative to the conventional practices adopted to build houses in Mexico, a novel selection of materials was proposed to form the different assemblies required for the CASA UNAM system, an efficient and eco-friendly house prototype to provide flexible solutions in urban areas. The design of the CASA UNAM system enabled substituting some environmentally harmful materials causing large impacts during their manufacture, such as cement or steel, by more sustainable components involving recycled or naturebased solutions capable of both lowering the emissions released to the atmosphere and ensuring an adequate functional behavior.

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Table 6 Materials used in the CASA UNAM system Assembly 1. Roads, networks, and utilities

2. Foundations and subsoil

Subassembly 1.1 Sanitation pipe 1.2. Industrial rack 1.3. Total screws for roof and floor 2.1. Foundations 2.2. Piles

3. Superstructure – Masonry

3.1. Main structure

4. Roofing – Waterproofing – Frame

4.1. Metallic pillars 4.2. Heat insulation 4.3. Paper tape 4.4. Waterproof layer 4.5. Roof

5. Partitioning – Backing wall – Ceiling – Interior joineries

5.1. Interior wall 5.2. Heat insulation 5.3. Paper tape 5.4. Interior doors

5.5. Furniture

5.6. Floor

6. Exterior joineries and façade

6.1. Floor

6.2. Windows glazing 6.3. Unfoldable doors 6.4. Plaster wall 6.5. Wetland 7. Cladding – Screeds – Painting – Ornamental

8. Heating – Ventilation – Cooling – Sanitary hot water

7.1. Walls 7.2. Plant pots 7.3. Painting 7.4. Textile 8.1. Solar heating waver 8.2. Façade ventilation

Material Polyvinylchloride Aluminum Steel Plywood Recycled polypropylene Galvanized steel sheet Steel Glass wool Paper fibers Fiber glass Wood board Plywood Plaster Recycled paper Glass wool Paper fibers National pine plywood Brass National pine plywood Textile cotton Plywood Glass wool Cumaru wood National pine plywood Cumaru wood Isolated glass Polyvinylchloride Isolated glass Polyvinylchloride Gypsum with fiber glass Polyethylene terephthalate Gypsum paste Clay Natural components Polyester threads Solar collector glass tube Pottery clay

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The highest environmental impacts identified during the lifespan of the CASA UNAM system were found in the assemblies related to the superstructure and partitioning of the building, with the great majority of these emissions corresponding to the manufacture and transport of textile materials, steel, and clay.

Conclusions and Recommendations This chapter overviewed the concept of Life-Cycle Assessment (LCA) and examined its usefulness to improve the management of materials from an environmental point of view. After providing a theoretical framework about LCA and the main phases and terms upon which it is based, a review of the most relevant investigations conducted in the last 20 years concerning the use of this tool in the field of materials management was conducted to prove how clarifying its implementation might be for making long-term eco-friendly decisions. A series of case studies were presented and discussed in detail to further shed light on the practical utility of LCA for managing the selection of raw materials for both buildings and infrastructures. The results yielded by these applications demonstrated that recycling wastes and using nature-based solutions can lead to considerably reduce the cradle-to-grave environmental damage caused by different materials. Special attention should be paid to the life cycle of these materials, since the emissions derived from their production might be compensated with less intensive maintenance practices if having enhanced aging properties. Another decisive factor in the results yielded by the LCAs of materials was found to be the transportation of materials throughout the different phases considered, e.g., from production to use or from exploitation to disposal. Therefore, the logistics related to resource location and supply, especially when materials are intended to form larger elements like buildings and infrastructures, must be considered carefully to minimize fuel consumption and carbon dioxide (CO2) emissions. In consequence, LCA provides a comprehensive framework to appraise the cradle-to-grave environmental impacts of products, supporting the adoption of measures related to the management of raw materials themselves, as well as that of two of the main materials-based drivers for modern daily life: buildings and infrastructures. In this sense, further research in the field of LCA should delve into both developing easy-to-use and interpret web-based interfaces for its application and seeking consensus in relation to the aggregation of impacts belonging to different categories, in order to facilitate global policy decision-making processes without requiring expertise in the field.

Cross-References ▶ Integrated Assessment of Environmental Factors: Risks to Human Health ▶ Performance Evaluation of Global Environmental Impact Assessment Methods Through a Comparative Analysis of Legislative and Regulatory Provisions

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Dispersion, Photochemical Transformation, and Bioaccumulation of Pollutants in the Vicinity of Highway

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Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Characteristics of On-Road Pollutants Emission . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Pollutants Dispersion Near Highway . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Pollutants Concentrations Within Highway . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Bioaccumulation of Pollutants in Roadside Vegetation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract

Road traffic is one of the most significant sources of the atmosphere pollution in urban areas, which leads to serious ecological, economical, and social problems. Among the most toxic components of the road traffic emission are carbon, sulfur, and nitrogen oxides, hydrocarbons, airborne particulate matter, polycyclic aromatic hydrocarbons, heavy metals, and dioxins. Some other toxic components are formed within the zone of a highway influence during the day time due to photochemical processes. This chapter gives an analysis of various physical, chemical, and biochemical processes that occur in the vicinity of highway and lead to transformation of pollutants and their interaction with the environment. Keywords

Dispersion · Photochemical transformation · Concentrations · Biochemical processes

G. Gerasimov (*) Institute of Mechanics, Moscow State University, Moscow, Russia e-mail: [email protected] © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_98

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Introduction The increase in road traffic leads to significant pollution of the environment in the vicinity of highway despite the environmental protection measures applied (Gilbert et al. 2007). Some components of the combustion products of automotive fuels, even at low concentrations in the exhaust gases, have a negative effect on the biosphere and cause significant material damage (Olson and McDow 2009). The scientific research in this field is mainly concerned with the study of the dispersion of harmful components of the vehicle exhaust in the environment, taking into account the characteristics of the road traffic and meteorological conditions, as well as with the investigation of the accumulation and photochemical conversion of harmful substances. The pollutants dispersion in the vicinity of the highway is usually considered on the base the Gaussian theory of admixtures dissipation in the atmosphere (Nagendra and Khare 2002; Seinfeld and Pandis 2016). The highway in this theory is interpreted as a line source, in which thermal and mechanical turbulences induced by moving vehicles contribute to mixing of the vehicle exhaust (Sahlodin et al. 2007). The integration of pollutants emissions along the trajectory of the vehicles motion gives mean pollutants concentrations in the vicinity of highway. The photochemical transformation of pollutants in the atmosphere numbers hundreds of chemical reactions and includes photochemical processes, nitrous and sulfur compounds conversion, and processes with participation of organic compounds (Szopa et al. 2005). The photochemical rate constants are dependent on sun zenith angle, cloudiness, and air temperature (Dechaux et al. 1990). The oxidation of saturated and unsaturated hydrocarbons that are presented in vehicle exhaust occurs at their interaction with hydroxyl radicals that are formed in photochemical processes. This leads to formation of organic peroxyradicals followed by their conversion to aldehydes and organic nitrates in reaction with nitrogen oxide. The end products of this photochemical process are peroxyacetyl nitrates that are the basic components of the photochemical smog (Leone and Seinfeld 1985). The bioaccumulation of hydrophobic chemicals in aquatic ecosystems in the vicinity of the highway leads to their inclusion in the air/soil – surrounding vegetation – animal – human food chain (Gobas et al. 2015). This process can be considered on the example of the most toxic representatives of pollutants such as polychlorinated dibenzo-p-dioxins and furans (PCDD/Fs) and benzo[a]pyrene (BaP). The corresponding equilibrium and kinetic models include mass exchange between air and vegetation, carcinogenic risk correlations, as well as air toxic admixtures degradation (Czub and McLachlan 2004). In this chapter, the current progress in the investigation of physical, chemical, and biochemical processes in the vicinity of highway that lead to transformation of pollutants and their interaction with the environment is reviewed. The review includes the examination of available experimental data and corresponding modeling studies. Since the number of publications on this subject of research is very large, the review fixes its attention on the most typical works.

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Characteristics of On-Road Pollutants Emission The most toxic components of the combustion products of motor fuels are nitrogen oxides (NOx), carbon monoxide (CO), sulfur dioxide (SO2), volatile organic compounds (VOCs), particulate matter (PM), polycyclic aromatic hydrocarbons (PAHs), PCDD/Fs, and heavy metals. During the daytime, other toxic substances may form under the influence of photochemical processes in the area of the highway. These pollutants include acids, peroxyacetyl nitrates (PANs), and ozone, which are the main components of the photochemical smog in cities. The formation of NOx in the combustion chamber occurs during the oxidation of air nitrogen. The main parameters that significantly influence on the amount of the formed NOx are the excess air factor, the maximum temperature in the combustion zone, and the residence time of gases in the high-temperature region. Nitrogen oxides can also form in the initial stage of combustion as a result of the interaction of HC radicals with N2 (so-called prompt NOx). The bulk of the total NOx produced in the combustion process is nitrogen monoxide: NO2 content in gasoline engines usually does not exceed 1%, while in diesel engines, where the excess air factor is always greater than 1, the amount of NO2 reaches 10% (Karavalakis et al. 2012). There are a large number of measurements of the emission factor (EF) ENOx from different engines. The emission factor ENOx = 0.1 g km1 for gasoline passenger cars was received with the help a chassis dynamometer (Wang et al. 2013). The measured emission factor of NOx that is emitted by light-duty vehicles in the Leopold II tunnel in the Brussels city center, Belgium, in January 2013 is equal to (0.544  0.199) g km1 (Ait-Helal et al. 2015). The total NOx emissions for the Euro6 fuel from light-duty diesel vehicles received using a portable emissions measurement system are ranged from 0.158 to 1.025 g km1 (Kwon et al. 2017). A grouping of measured NOx emission rates by hour of day in comparison with the hourly fraction of heavy-duty vehicles was made by Gordon et al. (2012). These data are given in Fig. 1. The median hourly emission rates demonstrate a linear increase with growth of heavy-duty diesel vehicle (HDDV) fraction. Fig. 1 The median NOx emission rates ENOx as a function of the heavy-duty diesel vehicle fraction (points) and the emission rates calculated with the lowest (dotted line) and highest (solid line) inventory emission rates (Gordon et al. 2012)

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The formation of CO occurs at the initial stage of hydrocarbon fuel combustion in the combustion chamber, followed by CO after-burning to CO2. A significant amount of CO in the combustion products is observed at the combustion of fuelenriched mixtures, as well as in the areas adjacent to the walls of the combustion chamber, where the temperature is insufficient for after-burning of CO. The content of SO2 in combustion products is determined by the sulfur content in the fuel, and if CO emission can be controlled by improving the quality of fuel combustion, then SO2 emissions can be reduced only by cleaning of the initial fuel. For example, the sulfur concentration in Euro3 and Euro4 fuels is 236 mg kg1 and 49 mg kg1, respectively (Karavalakis et al. 2012). The measured EF for CO that is emitted under highway conditions by diesel vehicle (Renault Kangoo van) equipped with a diesel oxidation catalyst (DOC) ranges within 0.14–0.19 g km1 (Rey et al. 2014). The measurements of the EF for SO2 in an urban tunnel (Pearl River Delta, China) averaged 21 mg km1 (Zhang et al. 2015). The VOCs are defined as a group of hydrocarbons (HCs) that evaporate at the room temperature. In a normally running engine and in the absence of misfiring, the unburned fuel HCs remain mainly in relatively cold areas adjacent to the walls of the combustion chamber. The VOCs emission is highly dependent on the speed of the car. The measured VOCs EFs for different gasoline passenger cars range from 0.10 to 0.25 g km1 at the low speed conditions and from 0.01 to 0.02 g km1 at the high speed conditions (Wang et al. 2013). The VOC compounds have been grouped into four categories: aromatics, alkanes, alkenes, and others. The percentage of the VOC categories varies depending on the speed conditions and averages 37.57, 29.13, 29.38, and 3.92%, respectively. The EF of total HCs for light-duty diesel vehicles is equal to 0.146 g km1 for the highway fuel economy cycle (Tsai et al. 2012). The formation of diesel exhaust PM that consists mainly of carbonaceous particles (soot) occurs as a result of the pyrolysis of HC molecules in highly enriched hightemperature zones of the combustion chamber, followed by their condensation into larger aromatic structures. Soot can be categorized into the following two types of particles: (a) fractal-like agglomerates of primary particles 15–30 nm in diameter, composed of carbon and traces of metallic ash, and coated with condensed heavier end organic compounds and sulfate; (b) nucleation particles composed of condensed hydrocarbons and sulfate (Maricq 2007). PM2.5, particulate matter with an aerodynamic diameter of less than 2.5 μm, is a major constituent of air pollution and is associated with respiratory and cardiovascular diseases as well as skin cell alterations (Frank et al. 2013). Conditions that promote the formation of soot are realized mainly in diesel engines. Average emission rates of 22.8 mg veh1 km1 and 187 mg L1 for PM2.5 were obtained for gasoline-powered vehicles in the tunnel study in Monterrey, Mexico (Mancilla and Mendoza 2012). The urban tunnel measurements (Pearl River Delta, China) with an average fleet composition of 61% light-duty gasoline vehicles, 12% heavy-duty diesel vehicles, and 27% liquefied petroleum gas vehicles give EFs of 82.7  28.3, 19.3  4.7, and 13.3  3.3 mg veh1 km1, respectively, for PM2.5, organic carbon (OC), and elemental carbon (EC) (Zhang et al. 2015). Polycyclic aromatic hydrocarbons and, in particular, their most toxic representative BaP are formed as a result of pyrolysis of heavy fractions of motor fuels

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and lubricant at relatively low temperatures (up to 1000 K) under conditions of deficiency in oxygen. Benzo[a]pyrene is well adsorbed in the pores of fine soot particles and is emitted into the environment practically only with them. The impact of diesel fuels (Euro3 and Euro4) on the PAHs emission from passenger cars was studied by Karavalakis et al. (2012). Total PAHs EF is ranged from 2 to 3.5 μg km1, depending on the fuel type and driving cycle (cold or hot). Total nitro-PAHs EF is ranged from 0.4 to 1.9 μg km1. The use of the Euro4 fuel gives a 30–40% reduction in the PAHs EF compared with Euro3 fuel. This fact is explained by the fact that the Euro4 fuel contains about 10 times less PAHs than the Euro3 one. The benzo[a]pyrene EF is equal to 0.047 μg km1 for the Euro3 fuel and 0.015 μg km1 for the Euro4 fuel (cold start). Corresponding values for the hot start driving cycle are equal to 0.005 and 0.001 μg km1, respectively. Approximately, the same EF values are obtained for diesel light-duty vehicles (Perrone et al. 2014). Polychlorinated dibenzo-p-dioxins and dibenzofurans are formed practically in all high-temperature processes associated with the combustion of organic fuels in the presence of various chlorine compounds (Gerasimov 2016). The problem of the environmental pollution by PCDD/F arose from the fact that noticeable amount of dioxins was found in the products of municipal solid waste incineration (Tuppurainen et al. 1998). As the results of experimental studies show, the formation of dioxins occurs mainly in the low-temperature area behind the combustion chamber as a result of heterogeneous catalytic reactions on the surface of condensed particles. The emission of dioxins for different engines and fuels varies widely. The United States’ mobile source inventory for on-road diesel engines offers the value of 946 pg I-TEQ L1 fuel consumed and 1.28 pg I-TEQ L1 fuel consumed for modern engines equipped with a catalyzed diesel particle filter and urea selective catalytic reduction (Laroo et al. 2012). According to the air sampling from tunnel in Taipei (Taiwan) from November 2001 to January 2002, the measured PCDD/Fs EFs are ranged from 5.83 to 59.2 pg I-TEQ km1 for gasoline vehicles and from 23.32 to 236.65 pg I-TEQ km1 for diesel vehicles (Chang et al. 2004). The results of the PCDD/F EFs measurements for light-duty diesel vehicles give 455 pg I-TEQ km1 (in the presence of the DOC) and 1330 pg I-TEQ km1 without catalyst (Rey et al. 2014). An important role in the pollution of the environment is played by heavy metals, anthropogenic emission of which causes great ecological damage to the biosphere. According to the USEPA, As, Be, Cd, Co, Cr, Hg, Ni, Mn, Pb, Sb, and Se are classified as the most toxic elements, whose compounds are subject to control (Sandelin and Backman 1999). The main sources of heavy metals in the vehicular traffic are: combustion of fuel, emission of wear products (tires, brakes, bearings, bodywork, and road surfaces), as well as secondary emission of the road dust. The emission of such elements as As, Hg, and Se is determined by fuel combustion, while Cd, Cr, Cu, Ni, Pb, and Zn emission is determined by lubricant oil combustion (Pulles et al. 2012). The average EFs (μg veh1 km1) of these elements, corrected for possible dust resuspension, are: 19, 22, 10, 32, 1, 29, 46, 32, and 137, respectively (Mancilla and Mendoza 2012).

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Pollutants Dispersion Near Highway Atmospheric dispersion models that describe the spread of air pollutants near highway are used for environmental impact assessment of vehicular exhaust emissions. These models have different levels of details and can be subdivided on box models, Gaussian models, Lagrangian/Eulerian models, and computational fluid dynamic (CFD) models (Holmes and Morawska 2006). Box models consider the vicinity of highway as a box into which vehicle exhaust is emitted and undergo chemical transformations. This type of models usually includes more detailed photochemical reaction schemes (Zhong et al. 2016). Gaussian models are based on the Gaussian dispersion formula that is frequently used within other types of models (Gibson et al. 2013). CFD models are based on the solution of the NavierStokes equations and give the most complete analysis of the pollutants dispersion in combination with the Eulerian or Lagrangian approach for the particulate phase description (Brusca et al. 2016). Gaussian type models are widely used in modeling of the pollutants dispersion from highway traffic (Venkatram et al. 2013; Shorshani et al. 2015). In these models, the highway is considered as uniform line source of admixture with the intensity QL per unit length, which can be directed along the Y axis of the Cartesian coordinate system (see Fig. 2). Since it is assumed that the source has an infinite length, the observation point P can be located on the X axis. The surface concentration c(X, θ) of the admixture at the observation point P as a function of the wind direction angle θ can be determined as follows. The linear source can be represented as an infinite sequence of point sources, the dispersion of the admixture from which is described by means of the Gaussian scattering theory (Seinfeld and Pandis 2016). In this case, the contribution to the concentration c(X, θ) from the point source considered as an infinitely small segment of length dY in the vicinity of the point Y can be represented in the form dc = QLΦ(X, Y )dY. The function Φ describes the scattering of the admixture across the direction of the wind as a function of the distance from the point source along the direction of the wind. The total surface admixture concentration at the point X is determined by the expression (Seinfeld and Pandis 2016): Yð max

cðX, θÞ ¼

QL ΦðX, Y ÞdY:

(1)

1

It should be noted that in the Gaussian scattering theory, the turbulent diffusion in the direction of the wind and against it is assumed to be zero, since its influence is insignificant in comparison with the admixture transfer due to the wind (Bange et al. 1991). Therefore, the integration in Eq. 1 is carried out from – 1 to Ymax = Xctgθ. The quantity Ymax is defined as the coordinate of the point of intersection with the Y axis of the straight line drawn through the observation point P perpendicular to the direction of the wind, as is shown in the Fig. 2.

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Fig. 2 Scheme of the pollutants dispersion from a point source located at the point Y; P is the point of observation

In the Gaussian scattering theory, the scattering function Φ in the coordinate system (x, y), where x is the direction along the wind propagation and y is the transverse direction, is written as (Bowne 1984):  1 Φðx, yÞ ¼ 2πUσ ð x Þσ ð x Þ y z h i expðλx=U Þ   2 2  exp y =2σy ðxÞ  ð1 þ PÞexp h2 =2σ2z ðxÞ :

(2)

Here U (m/s) is the average wind speed; σy and σz (m) are the transverse scattering coefficients in the horizontal and vertical directions; λ is the coefficient of the admixture decomposition; P is the degree of the admixture reflection from the earth’s surface; and h is the height of the line source. Under normal conditions of the highway, the coefficient of the admixture decomposition λ is equal to zero, the degree of the admixture reflection P is equal to unity, and the height of the line source h is small in comparison with the scattering coefficient σz and it can be set equal to zero. The scattering coefficients σy and σz depend on the state of the atmosphere and are determined in accordance with the classification scheme of Pasquill-Giffiod (Bowne 1984). According to this scheme, the state of the atmosphere varies from periods with a relatively small turbulence (stability class F) to periods with high turbulence (stability class A). For linear sources with a small distance from the source (not more than a few kilometers), these coefficients are linear functions of x: σy = Ayx and σz = Azx (Griffiths 1994). The Pasquill-Giffiod stability classes in the dependence on the horizontal wind direction fluctuations φ obtained over the 30-min observation period (Bowne 1984), and the corresponding parameters of the scattering coefficients (Griffiths 1994), are given in Table 1. Substitution of Eq. 2 into Eq. 1 with constant value of the source intensity QL and connection between coordinates (x, y) and (X, Y ) received from Fig. 2 leads after

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Table 1 The stability classes of the atmosphere as a function of horizontal wind direction fluctuations, and the parameters of the scattering coefficientsa Stability class A B C D E F a

φ, degree >22.5 17.5–22.5 12.5–17.5 7.5–12.5 3.8–7.5 1, and the y coordinate in this case coincides with the radius of the cylinder R, then the coordinate x, at which the diffusion coefficient Di must be calculated, may be estimated as: x R(21/2A)1. In the daytime at the stability class A (high level of the solar radiation and small wind speed), the diffusion coefficient Di can be estimated as 0.16 m2 c1 at R = 10 m and U = 0.1 mc1. The gas dynamic model described by Eq. 7 is a modification of the well-stirred reactor, in which the rate of the admixture feed into the reactor volume and the rate of its removal to the environment are known. The most important feature of the processes occurring in the reactor at a constant mass transfer rate is the formation of a stationary regime. For those admixture that do not participate in chemical reactions, the equality hRii = 0 must hold, and the average admixture concentration is easily obtained by equating the right side of Eq. 7 to zero. The currently available kinetic schemes of atmospheric chemical processes number hundreds of chemical reactions and include the conversion of nitrogen-containing and sulfur-containing components, photochemical processes, and processes involving organic compounds (Stockwell et al. 1997; Atkinson et al. 2004; Goliff et al. 2013). These schemes are intended for numerical modeling of the formation and conversion of air pollutants. A key role in the photochemical oxidation cycles in the atmosphere belongs to NOx and VOCs. The scheme of gas phase and heterogeneous reactions with participation of these air pollutants is given in Fig. 5. The description of the photochemical transformation of harmful admixtures within the highway is based on the simplified kinetic mechanism RADM (Regional Acid Deposition Model) (Stockwell 1986), which is the origin for the later RASM mechanism (Goliff et al. 2013). The RADM mechanism takes into account the main groups of gas-phase reactions and allows a fairly accurate description of the behavior of the chemically reacting system. The chain of chemical transformations in the RADM mechanism is initiated by the photochemical decomposition of NO2: O2

NO2 þ hν ! O3 þ NO

(8)

Further, the formation of OH and HO2 radicals takes place, which together with O3 are the main oxidants of NO and SO2 in the gas phase. The rate constant kNO2 of photochemical decomposition of NO2 depends on the solar zenith angle, cloudiness and temperature of the environment. For the summer cloudless sky, this value varies from 7.8  103 to 5.0  103 s1, when the zenith angle changes from 30 to 60 (Leone and Seinfeld 1985). The continuous cloudiness reduces the intensity of the ultraviolet radiation near the earth’s surface by a factor of 1.5–2 depending on the

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a Gas phase reactions of VOCs hv NO2

Products (HNO3)

NO2 hv

O O/ CH

O3 NO

Carbonyl Alcohol

RO• (OH•)

RO2• (HO2

•)

HO2•

ROOH (H2O2)

NO

2

RO2•

) 3

(R

RO

O2

ON

VOCs OH/NO3/O3/hv R• (H•)

Biogenic sources

O

2

Anthropogenic sources

b Heterogeneous reactions of VOCs VOCs

Trace gases sources

VOCs

Particle growth HMCs Hygroscopicity

e.g., C-S/N species

LMCs

aerosols

e.g., NH3, H2S CO2

peroxides Optical property

carbonyls organic acids

Nucleation ability

SOA formation

Reactivity

Health effects HMCs: high molecular weight compounds; LMCs: low molecular weight compounds

Fig. 5 Gas phase and heterogeneous reactions of NOx and VOCs (Shen et al. 2013)

day time Td. The averaged values of kNO2 in the daytime, obtained from the analysis of the results of measurements of NO, NO2, and O3 concentrations in the atmosphere, are proposed by Janssen et al. (1988). For summer, spring, and winter time of the year, these values are equal to 5.8  103, 4.2  103, and 2.5  103, respectively. The corresponding dependence of kNO2 on the day time Td can be obtained using a correlation (Gerasimov 2006): kNO2 = (kNO2)max sin1/2[π(Td – Tr)Tl1], that well approximates the data Janssen et al. (1988) for medium latitudes (53.5 of the Northern

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latitude). Here (kNO2)max is the maximum value of the rate constant during the day, Tl is the longitude of the day (hours), and Tr is the time of the sunrise. The kinetic scheme of gas-phase oxidation of nitrogen and sulfur oxides in the RADM mechanism is quite simple and can be represented by a chain of reactions leading to the formation of nitric and sulfuric acids vapors: O3

OH

OH

NO

NO ! NO2 ! HNO3 ! NO3 ! 2NO2 , O2 H2 O OH SO2 ! HSO3 ! SO3 ! H2 SO4 :

(9)

In the presence of droplet moisture (fog, low clouds), the bulk of SO2 is oxidized in liquid-phase chemical reactions. The determining role in the liquid-phase transformation of SO2 molecules is played by their chain-radical oxidation. The process is initiated by the interaction of HSO 3 with OH radicals, which enter the droplets from the gas phase (Gerasimov 2007). The role of liquid-phase processes in the oxidation of NO and HC is negligible (Hewitt 2001). Organic components in the RADM mechanism are combined into 16 groups of stable components and 8 groups of intermediate radicals. The groups of stable components include classes of alkanes, alkenes, aldehydes, ketones, and aromatic compounds. The oxidation of saturated and unsaturated HCs occurs mainly during their interaction with OH radicals. The resulting organic peroxyradicals (OPR) react further with NO, converting to aldehydes (ALD) and organic nitrates (ONIT): OH

NO

HC ! OPR ! ALD þ NO2 þ HO2 # NO ONIT

(10)

The rate constants of the individual stages depend on the initial hydrocarbon. The rate constant of the first stage for alkenes exceeds the corresponding value for alkanes by more than an order of magnitude. The interaction of aldehydes with OH radicals under atmospheric conditions induces a chain of reactions of peroxyacetyl nitrates (PAN) formation: OH

O2

NO2

ALD ! RCO ! RCðOÞO2 ! PAN

(11)

Here R is the HC radical. Peroxyacetyl nitrates are essentially a concentrated extract of smog and represent a strong oxidizing agent that is detrimental to plants and causes eye irritation. The OH radicals play a dominant initiating role in the chain of destruction reactions of PAHs and PCDD/Fs molecules. The process proceeds along the path of the OH radical addition to the aromatic molecule with an aromatic radical formation, followed by the aromatic radical destruction at its interaction with molecular oxygen and nitrogen oxide (Atkinson 2000). The rate of the process is determined by the rate of the first stage with the rate constant kOH that is equal, respectively, to 6.0  1013 and 4.6  1012 cm3 mol1 s1 for benzo[a]pyrene and 2,3,7,8-TCDD (Gerasimov 2015).

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Table 2 Comparison of the calculated concentrations of pollutants within highway with maximum permissible concentrations (MPCs) Admixture CO SO2 NO NO2 H2SO4 HNO3 PM2.5 Benzo[a]pyrene PCDD/Fs2 O3 PAN

Ei, mg km1 1.5  102 2.1  101 5.0  102 5.0  101 – – 1.0  102 2.0  105 1.0  107 – –

hcii, mg m3 5.2  101 7.6  102 2.0  100 7.6  102 1.9  105 1.7  104 4.1  101 7.7  106 4.2  1010 1.2  104 2.1  103

MPC1, mg m3 3.0  100 5.0  102 6.0  102 4.0  102 1.0  101 1.5  101 3.5  102 1.0  106 5.0  1010 3.0  102 1.0  101

1

Hygienic standard (2003) International Toxic Equivalency (I-TEQ)

2

Table 2 presents the calculated data of the average concentrations hcii of the main air pollutants within the highway 30 min after the beginning of chemical processes. The calculations were carried out at the traffic intensity N = 2 vehicles s1 and background concentrations of O3 and PAN equal to 30 and 2 ppb, respectively. The calculated concentrations are compared with the daily maximum permissible concentrations (MPCs) (Hygienic standard 2003). It can be seen that the hcii value for almost all air pollutants exceeds the daily MPC. Chemical processes lead to an increase in the concentration of NO2 by about four times compared with nonreactive gas. It is observed that the O3 concentration is less than its background concentration due to O3 consumption in the oxidation process. Chemical processes practically do not change the concentration of benzo[a]pyrene and PCDD/Fs due to the smallness of the corresponding rate constants.

Bioaccumulation of Pollutants in Roadside Vegetation The negative impact of trace organic contaminants in the atmospheric air on humans is primarily due to the bioaccumulation of these compounds by vegetation and their inclusion in the food chain. Being a powerful source of harmful substances, the highway leads to the pollution of agricultural lands (fields, meadows, pastures) located in the immediate vicinity of the highway. There are three main channels of air pollutants arrival to the plant leaves: (1) due to the mass exchange of vapors of organic contaminants with the plant material; (2) from the soil through the roots of plants; and (3) from aerosol particles precipitated on the plant leaves with adsorbed organic compounds on their surfaces. The first cannel is dominated, particularly for hydrophobic compounds (Bacci et al. 1990). Thus, the bioaccumulation of pollutants in the plant leaves can be treated as the process of establishing of phase equilibrium between the atmospheric air and the plant material.

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The plant/air partition coefficient KPA can be written as: KPA = (cP)eq/(cA)eq, where (cP)eq and (cA)eq are equilibrium concentrations of organic compound, respectively, in the plant (g m3 of the plant material) and in the air (g m3 of the air). The largest ability to accumulate harmful admixtures in plants belongs to fatlike substances (lipids and cutin) that are part of living cells and play an important role in biochemical processes. The most common model system for these substances is octanol (Rieder 1990). Therefore, the plant-air partition coefficient KPA can be represented in the form: KPA = vLKOA, where vL is the volume of the lipids and the cutin in fresh plant material, KOA is the octanol/air partition coefficient. A more accurate correlation KPA = mKOAn was proposed by Kömp and McLachlan (1997) on the basis of analysis of the experimental data, and parameters m and n for a number of plants were determined. In particular, the parameters of the above dependencies for one of the most common fodder and lawn grasses – ryegrass (Lolium multiflorum) – are equal: vL = 0.01, m = 2.75  104, n = 1.15. There are a large number of models with different levels of complexity that are designed to assess the effect of air pollution on the bioaccumulation of organic contaminants in plants (Fantke et al. 2016). In the simplest case, the process of accumulation of organic contaminants in the plant material is described by a general kinetic equation (McLachlan et al. 1995):   1 dcP =dt ¼ keff V 1 P cA  cP K PA , h i1 (12) 1 : keff ¼ D1 t þ ð k P aP V P Þ Here, keff is the effective mass transfer coefficient between the atmospheric air and vegetation (m s1); VP is the volume of the plant material per unit of the earth’s surface (m3 m2); Dt is the velocity of the admixture deposition due to turbulent diffusion in the vertical direction (m s1); kP is the coefficient of mass exchange through the cuticle, which is the layer of the cutin covering the leaf surface with a continuous film (m s1); and aP is the specific surface of leaves (m2 m3 of the plant material). To simplify further consideration, the direction of the wind is chosen to coincide with the coordinate direction X, when the wind direction angle θ is equal to zero (see Fig. 2). The surface concentration cA of the admixture is calculated from Eq. 3: h i1 cA ¼ 2QL ð2πÞ1=2 UAz x ð1 þ PÞ:

(13)

The degree of the admixture reflection from the earth’s surface P in the case, when the admixture absorption by the earth’s surface occurs only due to its bioaccumulation by vegetation, can be written in the form (Gerasimov 2002):  1 P ¼ 1  kef Dt ct A ðcA  cV =K VA Þ:

(14)

As can be seen from Eq. 14, in the presence of equilibrium between vegetation and air for a given admixture, the reflection coefficient P is equal to unity. At the initial time, when cV is equal to zero, the value of P depends on the nature of the mass

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transfer between air and vegetable material. If the limiting process is mass exchange through the cuticle of the leaves, the ratio Dt  kV aVVV is satisfied for the velocity of the admixture deposition and the reflection coefficient P is equal to unity. In the opposite case, the ratio P = 0 takes place, which means that a complete absorption of the admixture precipitated on the earth’s surface occurs in the vegetation leaves. The calculation of bioaccumulation of trace organic contaminants in roadside vegetation according to formulas (12, 13, and 14) was carried out by Gerasimov (2002) using PCDD/Fs as an example. Data for calculation of the kinetics of mass transfer between air and vegetation were selected in accordance with McLachlan et al. (1995): the volume of plant material per unit of ground surface VV is equal to 1.5  103 m3 m2; the specific surface area of leaves aV is equal to 7.2  103 m2 m3; the mass transfer coefficient through the cuticle of the leaves kV is equal to 8.0  1012KOA, where KOA for PCDD/Fs is equal to 5.0  1010 (Horstmann and McLachlan 1998); the volume vL of lipids and cutin in the fresh plant material is equal to 0.01. For the state of the atmosphere, a stability class C is chosen with the average wind speed U = 3 m s1. The velocity Dt of the admixture deposition due to turbulent diffusion in the vertical direction is calculated in accordance with the equation: Dt = UHx1, where the height H of the rise of a linear admixture source above the earth’s surface is equal to 1 m. The intensity of traffic N along the highway is equal to one vehicle per second, which corresponds to an average highway intensity of 86,400 vehicles per day. The average value of the PCDD/Fs emission E was chosen equal to 10 pg I-TEQ km1 at the optimal speed of vehicles 100 km h1. The dynamics of bioaccumulation of PCDD/Fs is shown in Figs. 6 and 7. Analysis of Fig. 6 shows that the absolute value of the surface concentration cA of the admixture increases with the progress of the process due to an increase in the reflection coefficient P. The characteristic equilibration time between the gas phase and the plant material at a distance x = 20 m from the highway is about 50 days. The results of calculation shown in Fig. 7 indicate that the concentration cV differs more and more from the equilibrium with increasing x, and at a distance x of more than 100 m from the highway, equilibrium with the gas phase is not established during the summer period of time. The presence of residential areas in the immediate vicinity of the highway leads to a constant direct exposure of harmful substances on human. Nevertheless, the entering of the admixture into the human body directly from the gas phase through the respiratory organs for some pollutants is only a small fraction of the total entering, in which the food chain “atmosphere – forage – animal – food products – human” plays a key role (McLachlan 1996). Therefore, it is interesting to consider the process of accumulation of pollutants in the human body on an example of the most toxic their representatives such as 2,3,7,8-TCDD and BaP. The degree of distribution of pollutants by phases for multiphase systems is determined by comparison of the fugacities of a given admixture in different phases. The admixture fugacity f is calculated using the relation f = c/ZM, where c is the admixture concentration in a given phase (g m3), Z is the fugacity phase capacity for a given admixture (mole m3 Pa1), and M is the molecular weight of the

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Fig. 6 Dynamics of PCDD/Fs concentration change in gas phase (cA) and plant material (cV) at distance x = 20 m from the highway

Fig. 7 Dynamics of equilibration between gas phase and plant material for PCDD/Fs at distance x = 10 (1), 20 (2), 50 (3), and 100 m (4) from the highway

admixture (g mole1). In the presence of phase equilibrium in the system, the fugacities of a given admixture in different phases coincide. The surface ecosystems are immersed in the air. Therefore, the air is a convenient test phase for consideration of the pollutants bioaccumulation in living organisms (McLachlan 1996). The fugacity air capacity ZA is determined by the formula: ZA = (RT)1, where R = 8.31 Pa m3 mole1 K1 is the universal gas

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constant, and T is the temperature. The fugacity capacity ZV of the plant material, in which the main accumulators of pollutants are fatlike substances modeled by octanol, is equal to vLKOAZA. The animal organisms, including the human body, are treated similarly with the replacement of vL by the volume fraction of fatty tissues vF. It should be noted that the temperature of the animals, as a rule, significantly differs from the ambient temperature. Therefore, the fugacity capacity of the living organisms at determination of the equilibrium concentrations is calculated at a temperature that is typical for a given organism. In particular, for PCDD the following relation occurs: Z(T ) = Z(298) exp.[11,300(1/T -1/298)] (McLachlan et al. 1995). According to the calculation data of the bioaccumulation of pollutants by vegetation, the surface concentration cA of 2,3,7,8-TCDD at a distance of 50 m from the highway 50 days after the beginning of the process is equal to 3.89  104 pg I-TEQ m3. The concentration cV is equal, respectively, to 3.33  104 pg I-TEQ m3, and the ratio cV/cAKVA is equal to 0.62, that is, the concentration of this compound in the plant material is only 62% of the equilibrium concentration. If it is assumed that a person with weigh 70 kg and a body density of 103 kg m3 drinks 1 l of milk daily, for the production of which 3 kg of plant material with a density of 900 kg m3 is required, then the consumed dose D of this compound in the absence of losses is equal to 1.6 pg I-TEQ kg1 day1. To assess the carcinogenic risk, the formula r = SD is usually used, where r is the probability of the disease (one additional case per r1 personы throughout life), and S is the carcinogenicity coefficient of the compound. The coefficient S for 2,3,7,8TCDD is equal to 105 kg day pg1 (Dowd 1988). Therefore, the consumed dose D obtained above gives the probability of disease r = 1.6  105. It should be noted for comparison that, the daily dose limit (norm) of the 2,3,7,8-TCDD according to USEPA (United States Environmental Protection Agency) is estimated in accordance with the value r = 106 (Dowd 1988). It is interesting to estimate the equilibrium concentration of 2,3,7,8-TCDD in the human body. The half-life of this compound is approximately 7 years (McLachlan 1996), which gives the value of the decay rate constant γ = 3.14  109 s1. The relation D = γcH takes place in equilibrium, whence a value of 5.9  106 pg I-TEQ m3 for the equilibrium concentration cH is obtained, which makes it possible to evaluate the fugacity fH of 2,3,7,8-TCDD in the human body: fH = 7.57  103 pPa at vF = 0.1, M = 322 g mole1, and T = 310 K. The fugacity of this compound in the air is equal to 2.99  103 pPa at T = 298 K and cA = 3.89  104 pg I-TEQ m3, which gives a value of 2.53 for the ratio fH/fA. According to the data obtained from the measured concentrations of cH and cA (McLachlan 1996), the ratio fH/fA is equal to 1.5. Thus, the concentration of 2,3,7,8-TCDD in the human body exceeds the equilibrium concentration (cH)0 = fAZHM, which indicates the predominant role of the food chain in the inflow of this compound into adipose tissues, which are the main accumulator of toxic substances in living organisms. The results of calculation of the BaP bioaccumulation in the roadside vegetation give the following concentrations at a distance of 50 m from the highway 50 days after the beginning of the process: cA = 1.4  105 μg m3, cV = 2.19  103 μg m3

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(cV/cAKVA = 0.099). The dose D of consumption of this compound by a human under the conditions discussed above is equal to 0.104 μg kg1 day1, which for a carcinogenicity factor S = 7.3  103 kg day μg1 (Peters et al. 1999) leads to a probability of disease r = 7.67  104. The half-life of BaP is 45 days (Peters et al. 1999), which gives the value of the decay rate constant γ = 1.76  108 s1 and the equilibrium concentration of BaP in the human body cH = 6.82  104 μg m3. The fugacities fA and fH in this case are equal to 1.37  104 and 9.76  106 μPa, respectively, and the ratio fH/fA is equal to 0.07. Thus, the role of the food chain in the accumulation of BaP in the human body is not so important, in the case of 2,3,7,8-TCDD.

Conclusions Road traffic is one of the most significant sources of the atmosphere pollution in urban areas. Therefore, the investigation of physical, chemical, and biochemical processes in the vicinity of highway is of great importance. The most toxic components of the on-road emission are NOx, CO, SO2, VOCs, PM, PAHs, PCDD/Fs, and heavy metals. Their dispersion is usually described by Gaussian type models with corresponding scattering coefficients that depend on the state of the atmosphere. The high air pollution within the highway has a negative impact on the health of people, who are somehow connected with road transport. The photochemical transformation of pollutants within the highway includes photochemical processes, nitrous and sulfur compounds conversion, and processes with participation of organic compounds. Measurements and calculations show that the concentrations of almost all air pollutants exceed the daily maximum permissible concentrations at typical traffic conditions. The negative impact of trace organic contaminants in the atmospheric air on humans is primarily due to the bioaccumulation of these compounds by vegetation and their inclusion in the food chain. There are a large number of models with different levels of complexity that are designed to assess the effect of air pollution on the bioaccumulation of organic contaminants in plants. The calculation of bioaccumulation of trace organic contaminants in roadside vegetation using PCDD/Fs as an example shows that the absolute value of the surface concentration of PCDD/Fs increases with the progress of the process due to an increase in the reflection coefficient P. The characteristic equilibration time between the gas phase and the plant material at a distance x = 20 m from the highway is about 50 days. The concentration of 2,3,7,8-TCDD in the human body exceeds the equilibrium concentration (atmospheric air/human body) at a distance of 50 m from the highway 50 days after the beginning of the process. This indicates the predominant role of the food chain in the inflow of this compound into adipose tissues, which are the main accumulator of toxic substances in living organisms. The role of the food chain in the accumulation of BaP in the human body is not so important.

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Part XII Environmental Risk Assessment

Metabolic Toxicity and Alteration of Cellular Bioenergetics by Hexavalent Chromium

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Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Basics of Cellular Bioenergetics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Solar Energy: The Source of Energy for Living Creature . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Characterization of Energy . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Transformation of Cellular Energy: Role of Metabolism . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chromium: A Transition Metal of Health Concern . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Detection . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . The Metal Chromium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Production, Use, and Human Exposure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Absorption, Distribution, and Metabolism . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . General Health Effects of Chromium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chromium as an Oxidative Stress-Producing Molecule . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chromium: A Potential Metabolic Disruptor . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Role of Hexavalent Chromium in Carbohydrate Metabolism . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Effect of Hexavalent Chromium on Protein Metabolism . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Effect of Cr(VI) on Serum Lipid Profile and Fat Metabolism . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Extensive industrialization, exhaustive mining, and whirlwind urbanization disseminate deep impact upon the world’s living being since dawn of the modern civilization. The consequences of all these environmental and anthropogenic

S. Pal (*) Nutritional Biochemistry and Toxicology Laboratory, Department of Human Physiology, Tripura University (A Central University), Suryamaninagar, West Tripura, India e-mail: [email protected] K. Shil Department of Human Physiology, Tripura University, Agartala, Tripura, India © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_58

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attributions in the atmosphere lead toward non-returnable noxious situation. Both intentional and accidental activities of human accelerate environmental degradation via release of pollutants and give a momentum in gradual bioaccumulation of toxicants in the food chain resulting in deleterious feedback to its creator in terms of hazardous and toxic health manifestations. Cr(VI) is an industrial, anthropogenic, and airborne environmental toxicant that acquires to have carcinogenic, genotoxic, and organ-specific irreversible complications. Due to its high diffusional proficiency, Cr(VI) penetrates into the cellular compartments through phosphate and sulfate anion exchange carriers. Cr(VI), being a potential oxidizing agent for organic compounds and strong free radical generator in the biological system, gets oxidized by ascorbic acid and glutathione and liberates different reduced forms of chromium. These in turn stimulate ample amount of free radical formation through Fenton reaction. Cellular macro- and micromolecules are very much delicate and sensitive to these free radicals like ROS, RNS, and singlet oxygen species which interrupt their normo-physiological purpose. Being a potent neurotoxic and antioxidant suppressive agent, Cr(VI) enhances lipid peroxidation and other neurological complications among the exposed organisms. Recently it has been also enlisted in the endocrine disruptor chemicals thus executing its hormonal and developmental toxicities. Through reduction of antioxidant level and flavoenzyme activity, Cr(VI) expresses sensitive mutational, structural, and regulatory intervention inside the cells. Cr(VI) promotes posttranslational modification principally by abnormal glycosylation which conveys conformational changes in active biomolecules and biological catalysts. Metabolism is an organized series of life-sustaining biochemical transformation of crucial biomolecules within the cells. Cr(VI) insensitively acts on metabolic pathways by altering the enzymological parameters, shifting the conformational architecture of enzymes as well as by exhaustion of substrate level in the cytoplasm. Cr(VI) toxicity compels significant depletion of glycolytic substrate and end products inducing hypoglycemic situation in the organism. Additionally, it also affects the glycogenolytic activity in the muscle and liver. The principal mitochondrial energy-generating pathways, viz., TCA cycle and oxidative phosphorylation, are found to be suppressed due to Cr(VI) toxicity in hepatic, muscular, renal, and cerebral tissue. The linking enzymes of cytoplasmic and mitochondrial metabolic pathways such as lactate dehydrogenase and pyruvate dehydrogenase are adversely affected by Cr (VI) toxicity. Different proteolytic enzymes and their activities as well as substrates of protein metabolism along with transaminase enzyme activity are notably altered in the organism owing to exposure of this heavy metal. Thus Cr(VI) toxicity prominently disturbs metabolic integration and subsequently alters cellular bioenergetics in the mammalian system. Keywords

Hexavalent chromium · Cellular bioenergetics · Carbohydrate metabolism · Protein content · Protease activity · Serum lipid profile · Fatty acid synthase

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Introduction Metal toxicity, a global concern, comes into picture from the very commencement of the so-called civilization which the human bore for their benefit and exposed for different purposes. From the dawn of industrialization in the eighteenth century to the present times, anthropogenic activities have triggered accumulation of huge amount of pollutants and industrial effluents to the environment (Steffen et al. 2015). The long-run effect of the atmospheric pollutants and toxicants leads to ecosystem imbalance, increases earth temperature, enhances evolutionary mutations in microorganisms, induces irreversible health complications, deteriorates atmospheric air particulate nature, and, last but not the least, causes severe and concerning global impact (Gunther and Hellmann 2017). Concomitant with development of science and technology, the use of metals has increased vigorously and simultaneously offers a dangerous and never-ending consequence of environmental plunders to the biosphere. Metal pollution has been amplified globally in a significant manner as an effect of industrialization since the late nineteenth and early twentieth century. The widespread application of the metallic ores and metalloids produces an alarming environmental situation with its efflux and effluents exposed to the organism through occupational, accidental, inhalational, and contact absorption (Yadav et al. 2017). Metals in their transitional form act as trace element in the animals’ and plants’ nutrition but in the valence state become toxic at a critical amount. Some heavy metals such as chromium, cobalt, mercury, cadmium, lead, etc. are important to note as these cause severe toxicity to the exposed organisms (Wang et al. 2012; Das and Pal 2017). Chromium (Cr) is a naturally occurring transition metal usually found in rocks, volcanic eruption, soil, as well as plant and animals. Chromium acquiring a molecular weight of 59.54 with 44 electrons in the electron shell can exist in several valence states. It can appear in the environment in organic and also as inorganic form; the organic form acts as trace element for the benefit of organisms, but the inorganic salt form is much toxic due to its powerful redox potential. Chromium exists in various oxidation states from divalent to hexavalent state (Doisy et al. 2013). Potassium dichromate (K2Cr2O7) and sodium chromate (Na2CrO4),two important inorganic salts of hexavalent chromium [Cr(VI)], have been widely used in numerous industrial processes such as metallurgy, chrome plating, chemical industry, textile manufacture, wood preservation, photography and photoengraving, refractory and stainless steel industries, and cooling systems (Park et al. 2004). The inappropriate use, disposal, and management of these products lead to heavy metal pollution that is a serious concern worldwide. Seven countries – South Africa, India, Kazakhstan, Zimbabwe, Finland, Brazil, and Turkey – which are leaders in terms of chromium pollution, produce large amount of industrial effluents containing Cr(VI) (Mishra and Malik 2012). India is the third and most important exporter of the chromium ore in the global scenario. It contributes 19% of world production, of which 99% is mined from Odisha where a huge number of opencast chromium mines meet the export demand throughout the year (Soudani et al. 2012; Mishra et al. 2016). Chromium compounds have various industrial applications including

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tanning, corrosion inhibition, plating, glassware cleaning, safety match manufacturing, metal finishing, and pigments where high concentration of chromium (40–50,000 ppm) has been reported in the effluents of these industries (Park et al. 2004). The most important feature of chromium’s toxicity is its persistence in the environment and its existence in several form and valence states. The hexavalent form, Cr(VI), leaches to the water bodies and penetrates to the groundwater level causing acute toxicity via contact, use, and consumption of natural water. Cr(VI) can penetrate through the cell membrane more easily than the other chromium compounds thus executing its potential toxic manifestation; the severity of chromium toxicity depends on its valence state. Chromate and dichromate are highly soluble in water and thus readily absorbed by the gastrointestinal tract due to its high diffusional capacity through all types of cell membrane (Clarkson 1993). Once inside the cell, Cr(VI) is metabolized to trivalent chromium, either enzymatically (via microsomal enzymes) or nonenzymatically (via ascorbate and GSH). The liver is the main organ responsible for the metabolism, detoxification, and secretory function of the organism; it regulates plenty of metabolic pathways of mammalian systems including metabolic interconversion of Cr(VI) to Cr(III). The primary transmit of this toxicant is inhalation (Bagchi et al. 2002); other probable forms include oral ingestion of contaminated water or direct dermal contact with products manufactured using chromium such as pressure-treated wood, paint, welding materials etc. (Ahmed et al. 2014). Of the several valence forms, Cr(0), Cr(III), and Cr(VI) predominate, yet only Cr(VI) is found to be carcinogenic as well as mutagenic and also abundant in natural water (Bagchi et al. 1997, 2000). The diffusible form of chromium as Cr(VI) accumulates in various tissues of the exposed organism and executes patterns of cellular toxicity, disrupts normal morphophysiological prototype of cellular integrity, and seriously disturbs biochemical as well as enzymological functions related to the metabolic process (Arslan et al. 1987). Cr(VI) causes several health complications ranging from chronic skin to malignant cancer (Naz et al. 2016). Cr(VI), being a potent oxidative stress modulator and apoptotic signal enhancer in the renal tissue, significantly executes acute renal damage and advanced tubular necrosis in mammalian systems of the experimental animal (Hegazy et al. 2016). Genotoxicity of Cr(VI) is well evident by formation of excessive free radicals and by producing DNA cross-linking, DNA protein cross-binding, chromosomal aberration, and genomic instability in the nuclear environment of the specific cells (Velma and Tchounwou 2013).

Basics of Cellular Bioenergetics Living organisms are extremely structured and composite creatures. The cell, tissue, and organ systems exhibit unique property to respond to the stimulus from internal and external origin and accordingly alter cellular as well as physiological attribution to cope up with the circumstances and also to maintain equilibrium in cellular context (Alberts 1998). Subsequently organism grows and reproduces, and in the

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course of evolution, diversification and speciation occur. All life-sustaining physicochemical responses necessitate sufficient and uninterrupted supply of energy. ATP, the high-energy phosphate molecule, is produced maximally through aerobic oxidation of fuel molecules such as glucose, protein, and fat. Insufficient ATP production due to external or internal factor or genetic shortcoming facilitates alteration in the cellular structure and nature of its functions, i.e., growth, development, repairment, and maintenance of metabolic homeostasis (White 1943). Extraction of energy and transformation into fuel for general functions as a uniform and composite functioning unit are exceptional attributes of all living organisms.

Solar Energy: The Source of Energy for Living Creature Plants and some photosynthetic bacteria can entrap solar energy and photochemically convert it into food energy for their metabolic purpose as well as to provide energy to the saplings. To accomplish this, carbon assimilation process helps to produce biologically essential macromolecules like glucose, amino acids, etc. in plants. Organisms surviving through holozoic nutrition are incapable to trap solar energy thus dependent on plants and animals to import chemical bond energy to accomplish energy-linked physiological functions (Raubenheimer et al. 2009). Eventually, an organism may get its energy from plant or animal, but the sole source of energy is the solar energy that is circulated from one trophic level to the other through the food chain of ecosystem. Living organisms synthesize energy from organic macromolecules through cellular respiration where oxygen operates a significant role with resultant production of CO2 and H2O. Within the cell, mitochondria are considered for the ultimate production house of burning fuel in terms of ATP (Alberts et al. 2013). That ATP is utilized for cellular function to its optimum level by a healthy and energetic tissue. However, unutilized biofuel is preserved in terms of glycogen or lipids for future use in energy-deprived situation (Fig. 1).

Characterization of Energy Energy can be categorized as the capability to perform work. In biology, work signifies one of the three properties such as chemical work, transport work, and mechanical work. Energy can also be characterized depending on its source, mode of action, and site of action. Different forms of energy are electrical, thermal, and photon energy. All these forms are distinct from each other though share a common characteristic to materialize into two forms, the kinetic energy and potential energy (Barak et al. 1997). Energy of motion is solely related to the kinetic energy, whereas potential energy is depicted as the stored energy. The classical example of kinetic energy in biological context is the transport of molecules across the biological membranes. The important attribution of all types of energy is reversible conversion of potential energy to kinetic energy.

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Fig. 1 Major energy source and energy metabolic pathways in organism

Transformation of Cellular Energy: Role of Metabolism All types of cellular activity require energy through specific metabolic process. Metabolism is a multifaceted network of extremely synchronized biochemical reactions in which the physiological functions proceed inside the cell to make a balance between its demand and supply. Every step of metabolic pathways is a diverse enzymatic reaction considering a specific enzyme, substrate, and definite cofactor. These reactions are of two categories: extraction of energy from calorigenic nutrients

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like proteins, carbohydrates, and fats and production of high phosphate molecule and reduced electron acceptor via synthesis or breakdown of the nutrient molecule. Energy from high-energy phosphate bonds of ATP or high-energy electrons of NADPH, FADH2, or NADH is transferred to the covalent bonds of the biomolecules (Alberts et al. 2013). Metabolism has its two wings, the anabolism and catabolism. Anabolic reaction leads to the formation of new molecules partially using the cellular energy with the help of enzymes. On the other hand, catabolism refers to breakdown of one molecule to produce another in response to cellular demand. Catabolic and anabolic reactions happen concurrently in the cells of the organism to maintain homeostasis between breakdown and synthetic phenomena. Catabolism is processed in two ways using aerobic and anaerobic respiration. Aerobic respiration is economical over the anaerobic reaction for the cells. In aerobic respiration glucose generally pursues two metabolic pathways: glycolysis and citric acid cycle (tricarboxylic acid cycle). The breakdown product of lipid especially glycerol also takes part in glycolytic reactions and finally gives rise to pyruvate. This ensures a metabolic link between carbohydrate and fat metabolism. Pyruvate gets converted into acetyl CoA and afterwards enters into the mitochondrial compartment to participate in TCA cycle. Proteins are dissociated into the amino acids that may act as the intermediates of TCA cycle and provide a metabolic link. On the other hand, fatty acid is also enzymatically catalyzed to acetyl CoA to support TCA cycle. By this way carbon from glycolysis and other metabolic sources enters the TCA cycle thus leading to an interminable cycle of energy production in terms of ATP, NADH, FADH2, and NADPH. The entrapped energy of NADH and FADH2 is transferred through the electron transfer system (ETS) of the inner mitochondria thus producing high-energy phosphate bond of ATP by chemo-osmotic reaction with the help of F1F0ATPase enzymatic system (Alberts et al. 2013; Wang and Oster 1998). Any disturbance in energy production imposes imbalance between expenditure and renewal thereby restricting the communication between metabolism and supply of nutrients and essential biomolecules. However, disturbed ETS function within mitochondria may lead to leakage of free electrons that can trigger formation of harmful free radicals especially superoxide anion promoting oxidative stress within the cell.

Chromium: A Transition Metal of Health Concern Detection Chromium is the twenty-fourth element in the periodic table in the group of transition metal with molecular weight of 51.99 g/mol and atomic number 24 (Brandes et al. 1956). It is named so due to its multicolored attribute and from Greek word “chroma.” This element generates many beautiful colored compounds naturally assimilating with different types of metals in different ratios. Chromite is a red-colored chromium compound used as the gem stone; moreover the red color of rubies, pink color of sapphire, and green color of emerald are also due to the

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presence of certain amount of chromium among their metallic components (Holleman et al. 1985). In early 1761 it was discovered as red-colored crystalline pigment of lead chromate (crocolite). According to archaeologists, in the late third century BC, the Qin Dynasty used terracotta army weapons coated with chromium (Cotterell 2004). Johann Gottlob Lehmann, a German mineralogist and geologist, on July 26, 1761, isolated an orange-red-colored compound from the mine of Beryozovskoye of Ural Mountains which he misidentified as lead-iron-selenium compound, but originally it was lead chromate. Later in 1770, Peter Simon Pallas found a unique compound from the same place Lehman visited and noticed that it contained coloring property that may be used as paints (Guertin et al. 2005). In 1797, Louis Nicolas Vauquelin isolated metallic chromium from Peter’s coloring compound with high-temperature oxidation in a charcoal oven, and this finally led to the discovery of chromium metal (Vauquelin 1798). In the eighteenth century, chromium was commonly used in paints and tanning salts, and the production of inorganic compound of chromium salt was initiated (Dennis and Such 1993). Chromium is now commonly used as metal alloys and in industries. Full-fledged uses in stainless steel production and several anticorrosive metals have been practiced now (Nordberg et al. 2014).

The Metal Chromium Metallic Properties Chromium is a steel-gray lustrous metal that can be polished to achieve its shiny texture for commercial purpose. It is a very active metal which merely doesn’t react with water but reacts with most acids. Oxidation of this element continues in the room temperature and subsequently produces Cr(III) oxide. In a stable environment, the oxidation procedure creates a protective layer that prevents further corrosion of this metal (Brandes et al. 1956). Magnetic Properties Elemental chromium contains paramagnetic characteristics. Depending upon the temperature, it shows differing attributes of magnetic properties. Alteration in the temperature affects the electron spinning alignments. Chromium oxide acquires ferromagnetic feature; for this it has the capacity of data tape, a specific way to store information. By differing in the ratio of chromium with other elements, the hybrid compounds can attain magnetic properties. Some stainless steel compounds have magnetic properties that solely correspond to the amount of chromium in them (Ishikawa et al. 1965).

Production, Use, and Human Exposure Chromium, a mineral of earth crust, is mostly excavated through mining. The total amount of chromium produced is million tons per annum. The principal contributor

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of chromium for industrial and other essential purposes is South Africa, India, Kazakhstan, Zimbabwe, Finland, Brazil, and Turkey (Mishra et al. 2016). In India, through opencast mining, maximum amount of chromium is produced in Odisha. Throughout the world, chromium is exported and imported as ferrochrome. Metallic chromium for industrial uses is produced by the reduction of raw chromite ore which is excavated from the mines. Inorganic salts of chromium such as sodium and potassium dichromate are produced by sweltering chromite ore with soda ash, and all these compounds are widely used in leather tanning, wood preserving, photoengraving, and refractory industries. Metallurgical industry mostly uses the metallic chromium for the production of stainless steel, ferrous, and nonferrous alloys (Barnhart 1997) (Fig. 2). Moreover, chromium introduces itself in the nature from combustion processes mainly as Cr(III) oxide. Cr(VI) has been found in fly ash of power plant systems (Stern 1982). Both of these forms of chromium reach to the water sources leaching from the industrial effluents or from the contaminated soil (Pellerin and Booker 2000). Minor amount of chromium is also contributed through the natural phenomenon such as volcanic eruption and dispersal to soil by wind and rain. By these ways Cr(VI) reaches and accumulates in the groundwater and simultaneously contaminates the drinking water. Several studies confirm that Cr(VI) is found in an extensive amount in the groundwater of the nearby areas of different chromium using industries (Saha and Orvig 2010). The industrial effluents from textile, metallurgy, and leather industries have been illegally dumped into the open environment or in deep burials without proper treatment which ultimately release noxious pollutants including chromium in the environment (Kumar et al. 2008). The principal way of exposure of Cr(VI) is inhalation followed by ingestion and dermal absorption as well as by the chronic or accidental acute ingestion (ATSDR 2000).

Most common forms of chromium compounds and their uses

Potassium dichromate K2Cr2O7

Dye industry Corrosion inhibitor Leather tanning Chrome platting Laboratory reagent

Glass cleaning Colored glass making Ceramic glazes Chrome platting Oxidizing agents Instrument repair industry

Chromic acid H2CrO4

Leather tanning industry Cement industry Analytical reagent Wood treatment and processing Photography and photoengraving Chrome platting

Chromium 24

Cr

51.996

Sodium dichromate Na2Cr2O7

Chromium oxide (Cr2O3)

Pigments and ink preparation Colored glass making Polishing Surface of optical devices Paintings

Fig. 2 Different chemical forms of chromium and their sources and uses

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Sources and Assimilation of Chromium into the Body

INHALATION

Sources of hexavalent chromium

EXPOSURE

TANNING INDUSTRY

WELDING

ACCIDENTAL INGESTION

CONTACT ABSORPTION CHROME PLATING

INDUSTRIAL EFFLUENTS

Fig. 3 Exposure of human being to chromium from various environmental sources

Occupational exposure is prominent among the workers of different industries including welders and chrome miners where Cr(VI) is responsible to cause severe health complications to the exposed individuals. Exposure of this toxic metal to the community takes place through stained foods, contaminated water, and polluted air. The maximum exposure of Cr(VI) becomes apparent to the workers involved in chrome plating and chromate production (ATSDR 2000) (Fig. 3).

Absorption, Distribution, and Metabolism The absorption of ingested or inhaled Cr(VI) follows certain characters such as oxidation state, solubility, and particle size of the metallic compound (ATSDR 2012). Outsized particles (>10 μm) of Cr(VI) compounds if inhaled are retained mostly in the upper respiratory tract, whereas smaller particles can move toward the lower respiratory tract. A small amount of accumulated Cr(VI) can be reduced to trivalent form (Cr(III)) in the interstitial lining fluids reacting with cellular biomolecules or antioxidant parameters present within the bronchial tree (Petrilli et al. 1986). At physiological pH Cr(VI) attains membrane-penetrable capacity due to its tetrahedral oxy-anions; on the other hand, Cr(III), being of predominantly octahedral configuration, is practically impermeable to the cell membrane. Cr(VI) enters into the cells via nonspecific anion exchangers of sulfur and phosphorus and subsequently is reduced to form chromium derivatives (Arslan et al. 1987) (Fig. 4). Inhaled Cr(VI) from the lungs reaches to the gastrointestinal tract through mucociliary clearance and gets absorbed there. Within the gastrointestinal tract, the intestinal lining mainly absorbs Cr(VI) compound consumed orally, by dermal contact or by inhalation (Kerger et al. 1997; Kim et al. 2004). According to the recent study, it is found that Cr(VI) is easily absorbed through the oral administration

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Ingestion and Exposure

Absorption

• Through drinking water • Contaminated foods

• Intestinal lining • Contact (dermal) absorption • Upper and lower respiratory tract

Metabolism • Reduced by salivary and gastric

•

Reduced by microsomal enzymes

• Cellular nonspecific anionic exchanger

•

Antioxidant coupling reduction reaction

•

Bio-accumulates in different vital organs like, liver, kidney, muscle, brain, spleen, bones.

• Occupational exposure • Inhalational and dermal contact

juices

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Excretion • Through excretory substances • Through faces

Fig. 4 Pharmacokinetics of hexavalent chromium compounds

to the experimental animal (ATSDR 2000). After entering into the cellular atmosphere, Cr(VI) is sequentially reduced to Cr(V), Cr(III), and Cr(IV) reacting with cellular antioxidants and ascorbic acid (Sehlmeyer et al. 1990). Among these Cr(VI) is more efficiently absorbed than any other form of chromium due to its high diffusional capacity and enters into the cell via nonspecific anion channels; on the other hand, Cr(III) is taken up by the cells through phagocytosis or by passive diffusion thus penetrating in negligible amount inside the cells (ATSDR 2000). Additionally, little amount of Cr(VI) is reduced by salivary and gastric juices by the reductive elements present in their specific compositions and ultimately excreted through the feces. The absorbed amount of Cr(VI) comes to the blood stream and gets distributed throughout the body, deposited among the tissues, and the rest is excreted through urine (Costa 2003). Bioaccumulation of Cr(VI) occurs irrespective to the organs and organisms exposed to this toxic heavy metal though larger amount of Cr(VI) is found to be deposited in the liver, kidney, spleen, and osseous tissue (Kargacin et al. 1993; Mancuso 1997) (Fig. 5).

General Health Effects of Chromium Organic form of chromium comprises of Cr(III) present in dietary vegetables, fruits, and animal sources (Smart and Sherlock 1985). Cr(III) is very much essential for the carbohydrate and fat metabolism and regarded as the trace element of the tissue (Anderson 1998). Oral administration of Cr(III) in trace amount doesn’t show any toxicity, but prolonged medication of it may have some adverse health effect to the individuals (Elbetieha and Al-Hamood 1997). By contrast exposure of Cr(VI) produces severe organ toxicity, carcinogenicity, and toxic exaggeration of metabolic parameters and enzymatic profiles (Abreu et al. 2014).

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Cr(VI) Intracellular reduction

Cr(III)

Cr(IV)

Cr(V)

ROS generation, DNA-Cr, DNA cross links, DNA-cross binding

Effectively Downregulates

Effectively Upregulates

Moderately Upregulates

Moderately Downregulates

Interface with genes of immune regulation

Obstruction in enzyme activity and metabolic pathways

Interfere in the chromosomal level

Interruption in cellular processes

Immunoresponse and disease presentation mechanism

Stress response, energy generation, xenobiotics

Apoptosis, cell cycle, DNA repair and metabolism

Cytoskeleton and cellular biosynthesis process

Fig. 5 Some adverse effects of chromium compounds at a glance

Long-term exposure of hexavalent chromium may result in dermatitis, skin rash, and ulceration of nasal mucous membrane of the exposed individuals (Saha et al. 2011). Chromium-handling workers of different factories may have common complaints of allergic sensitization (Shelnutt et al. 2007). Additionally, Cr(VI) causes organ toxicity depending on its duration of exposure and degree of deposition within organs. In hepatocytes it causes cytoplasmic vacuolization, hepatocellular inflammation, and steatohepatitis of the exposed organism and is also responsible for systemic organ dysfunction (Costa 2003). Nephrotoxic effect of Cr(VI) is observed among occupationally exposed individual in terms of excessive fat deposition in different regions of the kidney with severe deregulation of metabolic pathways (Kumar and Rana 1984). Proteinuria and oliguria associated with nephrocellular abnormality are the distinctive features of Cr(VI)-assisted kidney toxicity (DiazMayans et al. 1986). Hematological parameters, including respiratory pigment and soluble antibodies of body fluid, drastically changed due to the Cr(VI) intoxication (Costa 2003). Intracellular reduction product of Cr(VI) has greater affinity to react with cellular macromolecules causing several biochemical obstructions inside the cells including carcinogenicity, mutagenicity, and chromosomal aberrations (Zhitkovich 2011). Moreover, Cr(VI) seems to bioaccumulate in a great extent in different segments of nervous system and creates irreversible toxic manifestations in a proportion to the percent of Cr(VI) accumulation (Nudler et al. 2009). The hexavalent chromium enters into the cellular structure and produces irreversible toxic derangement to the cellular conformity resulting in the devastating ending of the target tissue (Kirman et al. 2017). Recent studies show that Cr(VI) exerts neurotoxic effect on mature neuronal cell and brain tissue and subsequently fluctuates physiochemical function of the nervous system and brain (Dashti et al. 2014;

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Singh and Chowdhuri 2016). Moreover, Cr(VI) hoarding in rat hypothalamus, anterior pituitary, and hepatic as well as muscular tissue implements toxic infestations among these tissues (Quinteros et al. 2007). Heavy metals including chromium directly or indirectly affect immune system and provoke different forms of immune reactions like neoplasia, autoimmune response, allergy, hypersensitivity reaction, and inflammation. Chromium compounds significantly decrease the T and B lymphocyte proliferation in the cellular level during pathogenic or infected condition and produce immune suppression. Lower doses of Cr(VI) through inhalation elevate phagocytic activity of alveolar macrophages, whereas higher dose of this metal trims down macrophage’s functionality. Therefore, Cr(VI) is responsible for dwindling of humoral immunity among the exposed organism. Antigenic substances or pathogen entry into the cell activates macrophages which result in the formation of nitric oxide and function as cytotoxic, antimicrobial, and antihelminthic agents (Mac-Micking et al. 1997). Study reports exhibit that Cr (VI) deliberately interferes with nitric oxide formation by macrophage cells of the exposed tissue (Srivastava et al. 2014). Moreover, chrome-cobalt alloy proliferates inflammatory mediators from sensitized macrophages subsequently triggering acute inflammation and necrosis (Cohen et al. 1998). Expression of TNF-α (tumor necrosis factor-alpha), NK cells, lymphocytes, neutrophils, and mast cells that make up acutephase reactions is found to be suppressed by hexavalent chromium (Jain and Kannan 2001). Toxic exposure of Cr(VI) produces retardation in the quantity of splenocytes and spleen weights and resultant decline in the blood lymphocytes thus interfering with the blastogenesis and immunoglobulin formation (Borella et al. 1995). Thus Cr (VI) affects different forms of immunogenic molecules and vital immune pathways by suppressing or by reverse modulation of immune response. Moreover, occupational, environmental, or accidental exposure of endocrine disruptor chemicals (EDCs) has been associated with numerous reproductive and physiological abnormalities. Cr(VI) has been recognized as a endocrine disruptor used by several industrial processes which contribute to chemical abundance in the atmosphere. Cr(VI) is evidenced to cause severe reproductive toxicity to the exposed organism irrespective of gender. It produces morphological alteration of the seminiferous tubule with enlarged intracellular spaces, prominent decline in sperm count and motility, and toxic changes in epididymal spermatozoa and Leydig cells’ architecture (Marouani et al. 2017). Reproductive hormones are very much sensitive to EDCs and supposed to act randomly in the presence of these chemicals. In connection with this, it is found that gestational exposure of Cr(VI) causes enhancement of germ cell or oocyte apoptosis, deregulates the hormonal secretion, disrupts fetal development, and produces fetal organ malformation (Gale 1978). Additionally, Cr(VI) can cross the bloodbrain barrier and is also able to pass from mother to offspring through breastfeeding (Samuel et al. 2014). Cr(VI) significantly accumulates in the testis, ovary, and fetus if exposure occurs during gestational periods (Trivedi et al. 1989). It is also responsible for the alteration of steroid and peptide hormones and depletion of antioxidant and enzymatic entity of testicular and ovarian tissues (Banu et al. 2011). Alteration in histoarchitecture and extension of the estrous cycles in the organism are also noted in hexavalent chromium intoxication (Murthy et al. 1996).

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However, metabolic alterations are the important phenomena by which most of the metals or metalloids including chromium exert their toxic manifestations. Few reports are available supporting the metabolic perturbation in chromium intoxication (Tajima et al. 2010; Pan et al. 2012; Xiao et al. 2012). Cr(VI) exerts its harmful effects by altering the metabolic parameters in the organism exposed to this noxious element. It alters enzyme expression by inhibiting enzyme’s activity or by diminishing functional attribution of crucial enzymes. It mimics property of some anions that may influence specific binding capacity of certain enzymes resulting in disturbance of the enzymes’ general biochemical activity. Additionally, Cr(VI) significantly depletes sulfur ions from the cellular compartments and prominently deregulates different metabolic enzymes that are selectively dependent upon the sulfur ions for their active participation in the metabolic reactions (Holland and Avery 2011). It was also responsible for the toxic alteration of enzyme expression at mRNA level in different tissues (Abreu et al. 2014) and actively deteriorates the metabolic functionality of different enzymes.

Chromium as an Oxidative Stress-Producing Molecule Cr(VI) is a potent oxidative stress modulator. Intracellular reduction of the Cr(VI) liberates huge amount of free radicals, reactive oxygen species, reactive nitrogen species, and superoxide anion molecules. All these harmful radicals attack macroand micromolecules of body fluid as well as cellular compartments (Soudani et al. 2012). Surge of free radicals and stress molecules in response to chromium toxicity is responsible for the increased lipid peroxidation in the exposed tissue. Nervous system, being delicate in structure, is very sensitive to the oxidative stress molecules like ROS and RNS and eventually amplified lipid peroxidation and biochemical alteration of nervous tissue (Quinteros et al. 2008). In this connection, an earlier study stated that the treatment with Cr(VI) could lead to lipid peroxidation in mammalian cerebral tissue (Bagchi et al. 2002). Sub-chronic exposure of Cr(VI) shows significant decline in the level of antioxidant parameters of the liver and parallel increase in the lipid peroxidation indicative of oxidative damage in tissue (Bagchi et al. 2001). These findings thus strongly recommend that Cr(VI) somehow disturbs the mitochondrial electron transport chain to leak out free electrons that initiate excess free radical formation.

Chromium: A Potential Metabolic Disruptor Environmental toxicants are well-defined oxidative stress generator, mutagenic, and extremely accountable for the alteration of physiochemical and primary metabolic parameters. Non-biodegradable toxicants including heavy metals such as Cr(VI), arsenic, lead, etc. enter easily into the organisms through inhalation, drinking water,

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or incidental chronic ingestion which ultimately accumulate inside the organisms affecting crucial biomolecules and enzymatic functions (Soudani et al. 2013). Alteration in metabolic integrity among liver and kidney tissue is reported after hexavalent Cr(VI) treatment for 30 days (Shil and Pal 2017a, b). Subacute Cr(VI) exposure does not exhibit significant effect on gain in body weight, whereas mild increase in liver somatic index (LSI) and kidney somatic index (KSI) is reported in Cr(VI)-exposed mice (Shil and Pal 2017a, b). Increase in LSI after Cr(VI) exposure may result from fatty inflammation in hepatic tissue as a result of toxic injury imposed by Cr(VI) (Pedraza-Chaverri et al. 2008). Overaccumulation of Cr(VI), hyalinization of the glomerulus, and nephrocellular hypertrophy may be the possible reasons for increasing the KSI of chromium-exposed animals (Li et al. 2016). Cr(VI) is associated with overproduction of free radicals and other stress molecule lineages, consequently responsible for cellular hypertrophy (Ben-Hamida et al. 2016) that may be one of the important causes of increased LSI and KSI. The liver and kidney being chief detoxifying organs of the body generally take the overload of many environmental toxicants and thus become more vulnerable when there is deficit in cellular defense system. In a previous study, significant bioaccumulation of Cr(VI) had been noted in the renal tissue after subacute chromium toxicity that had been correlated with the metabolic and functional abnormalities of the kidney (Shil and Pal 2017b). It is well reported that biomagnification of Cr(VI) occurs in the organism exposed to this toxic metal either occupational, environmental, or accidental (Batvari et al. 2016). This accounts the viable toxic effects to the respective tissues in the Cr(VI)-treated animal.

Role of Hexavalent Chromium in Carbohydrate Metabolism Carbohydrate depletion is an important toxic manifestation of short-term exposure of hexavalent chromium supported by marked decrease in liver glycogen and pyruvic acid contents in mice model (Shil and Pal 2017a; Vutukuru 2005). Additionally, Cr(VI)-induced hypoglycemia in mice is evident from a dose-dependent study (Shil and Pal 2017a) which may result from renal glycosuria caused by impairment of renal reabsorption of glucose. In this connection, nephrotoxic manifestation of chromium cannot be ignored (Hegazy et al. 2016). It is reported that the kidney is one of the major target organs for many toxic substances, and within it the proximal tubule epithelium is the most important target site of toxicant-induced cell damage (Bashandy et al. 2016). The depressing effect of Cr(VI) on liver and muscle pyruvic acid content may be explained by the fact that glucose that was produced by increased glycogenolytic activity of liver was released immediately to the blood, and this in turn may reduce the accumulation of pyruvic acid in the hepatic tissue by lowering the glycolytic activity (Shil and Pal 2017a). Retardation of glycolytic activity by Cr(VI) may also result from reductive conversion of hexavalent Cr (VI) to trivalent Cr-ATP complex formation that behaves as competitive inhibitor for various ATP-dependent enzymes and several kinases involved in glycolysis as well as TCA cycle (Lippard and Berg 1997; Myers et al. 2011; Ahmad et al. 2011).

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Additionally, this decrease in glycolytic activity in observed tissues may be an effect of hypoglycemia induced by Cr(VI). This is evident from the fact that glucagon, a hormone secreted during hypoglycemia, activates a protein kinase that phosphorylates hepatic L-isozyme of pyruvate kinase to inhibit it and thereby retarding glycolysis (Myers et al. 2011). On the other hand, accumulation of lactic acid in the renal tissue that is associated with enhanced LDH activity following chromium intoxication suggests that utilization of pyruvate as the substrate for LDH is triggered to switch over aerobic to anaerobic pathway (Shil and Pal 2017b).

Hexavalent Chromium as Modulator of Glycolytic, Gluconeogenic, and TCA Cycle Enzymes Alteration in the bioenergetics in the various tissues may be due to abnormal activities of proteins which are directly involved in the metabolic pathways or due to altered substrate availability in the tissue for specific enzyme action (Pauls et al. 1986; Lippard and Berg 1997). Hexavalent chromium is reported to reduce hepatic lactate dehydrogenase (LDH) activity (Shil and Pal 2017a). LDH, being a key enzyme of metabolic link between glycolysis and TCA cycle, is involved in conversion of pyruvate to lactate and thus acts as a good indicator of cellular damage. Suppressed LDH activity by Cr(VI) may result from decreased availability of pyruvic acid as substrate in the hepatic tissue. Another suggestive mechanism of reduced LDH activity may be the leakage of this enzyme from the hepatic tissue as a result of toxic injury imposed by Cr(VI). Toxic damage of the hepatic tissue was indicated by defaming of its actual cellular structure by Cr(VI). Inhibited LDH activity by Cr(VI) was also observed in teleost fish hepatocytes and in renal tissue at sub-chronic exposure (Venugopal and Reddy 1993) and also in hepatocytes of the African catfish at subacute exposure (Kori-Siakpere et al. 2006).On the other hand, elevated LDH activity in the kidney following subacute chromium exposure may result from metabolic shift from aerobic to anaerobic pathway due to less supply of oxygen to the energetically exhausted renal tissue (Shil and Pal 2017b). Additionally, pyruvate dehydrogenase (PDH) activity was found to be decreased in renal tissue of mice exposed to 10 mg per kg body weight of Cr(VI) per day for a period of 30 days. This may in turn perturb the metabolic link between glycolysis and TCA cycle. On the other hand, certain gluconeogenic enzyme activities are also disturbed by short-term chromium exposure in animal model. Glucose 6-phosphatase (G 6-pase) and glucose 6-phosphate dehydrogenase (G 6-PD) activities were downregulated by Cr(VI) in mice kidney (Shil and Pal 2017b). This indicates less supply of calorigenic substrates from noncarbohydrate source in the renal tissue thus making the tissue energy deficient. Effect of Cr(VI) on certain enzymes of TCA cycle has been documented in animal model. Succinate dehydrogenase (SDH), isocitrate dehydrogenase (IDH), and malate dehydrogenase (MDH) are the important enzymes of the TCA cycle that play a crucial role in ATP or energy generation in all vital organs of the individuals. Hexavalent Cr(VI) exposure at subacute dose and duration exerts negative impact on succinate dehydrogenase enzyme activity in the hepatic and renal tissue thus hampering the ATP production because this enzyme not only plays important role in

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TCA cycle but also acts as a potential component of the mitochondrial electron transport chain (Shil and Pal 2017a, b). Suppressed energy production may trigger anaerobic metabolism in the affected tissue following Cr(VI) treatment. That TCA cycle is impaired by Cr(VI) treatment was also evident from the earlier observation (Molina-Jijon et al. 2011). Insufficient compensation by fermentation and inhibition of cellular respiration by hexavalent Cr(VI) cause imbalance in the nucleotide pool and ultimately obliterate the homeostasis of the energy status in Cr(VI)-intoxicated organs of the Cr(VI)-treated animals (Abreu et al. 2014). Moreover, Cr(VI) exerts powerful inhibition on mitochondrial dehydrogenases such as NADH dehydrogenase (mitochondrial complex I) and succinate dehydrogenase (mitochondrial complex II) which in turn causes depletion of NADH pool from the tissue (Ryberg and Alexander 1990; Bianchi et al. 1982; Shil and Pal 2017a, b). Significant decrease of mitochondrial enzymes in all of these target organs due to Cr(VI) toxicity indicates significant deterioration of the intracellular NADH pool and ATP production in that specific tissue. Less availability of NADH may also contribute to retardation of anaerobic conversion of pyruvate to lactate via suppressed activity of LDH (Table 1). Recent studies from our laboratory reveal that the activity of malate dehydrogenase (MDH) in mitochondrial isolate was increased in the liver, whereas it was decreased in the kidney (Shil and Pal 2017a). MDH, being a TCA cycle enzyme, also helps in gluconeogenesis to produce glucose from noncarbohydrate source (Bianchi et al. 1982). In this regard, oxaloacetate, the TCA cycle intermediate which is produced from pyruvate in the mitochondria by pyruvate carboxylase, is reduced to malate before leaving the inner mitochondrial membrane, and mitochondrial MDH helps in this reduction process. As Cr(VI) exposure causes hypoglycemia, the gluconeogenic activity of the liver may be stimulated to replenish the loss of blood glucose level. On the other hand, decreased mitochondrial MDH activity in renal tissue by Cr(VI) exposure indicates suppressed metabolic conversion of malate to oxaloacetate in TCA cycle, which may contribute to low energy supply to that specific tissue as a result of toxic insult (Abreu et al. 2014).

Effect of Hexavalent Chromium on Protein Metabolism Effect of Cr(VI) on Tissue and Serum Protein Content and Free Amino Acid Nitrogen Level Protein depletion is a serious metabolic imbalance in the hepatic and renal tissue due to Cr(VI) toxicity. Protein depletion in the affected tissue indicates the physiological approach to compensate energy demand or get adapted to the changed metabolic system which may lead to the stimulation of degradation processes such as proteolysis and utilization of degraded products for increased energy metabolism (Begam and Vijayaraghavan 1996; Palanisamy et al. 2011). The reduction of total protein content may be due to the breakdown of tissue proteins under the effect of heavy metal which in turn increased the free amino acid nitrogen concentration in various tissues (Shakoori et al. 1994). Earlier observation of Chandravarthy and Reddy

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Table 1 Cr(VI)-induced alteration of carbohydrate metabolism Carbohydrate metabolism Hexavalent chromium intoxication

Blood glucose level decreased (hypoglycemia) (Sastry and Sunita 1983) Pyruvate content decreased in liver and kidney tissue (Shil and Pal 2017a, b) Enhanced glycogen breakdown in experimental fish and mouse (Tewari et al. 1987; Shil and Pal 2017b) Altered TCA cycle enzymes in liver and kidney tissue (Shil and Pal 2017a, b) Decreased NADH dehydrogenase activity or diminished oxidative phosphorylation pathway in the kidney and in cell line (lung cancer cell) (Shil and Pal 2017b; Abreu et al. 2014) Lactic acid content increased in kidney tissue (Shil and Pal 2017b) Increased glycogenolysis in the renal tissue of mouse and fish muscle (Shil and Pal 2017a; Velma and Tchounwou 2013) Decreased glucose 6-phosphatase activity in the kidney tissue (Shil and Pal 2017b) Decreased glucose 6-phosphate dehydrogenase activity in the kidney tissue (Shil and Pal 2017b)

(1994) revealed a remarkable decrease in total protein content in fish gill and brain with increased activities of transaminases and proteases on exposure to heavy transitional metal. Enhancement of free amino acid nitrogen level in the hepatic and renal tissues may be ascribed to the enhanced accumulation of protein degradation products after Cr(VI) intoxication or may be due to mobilization of free amino acids from peripheral tissue like skeletal muscle to the liver and kidney for supplying substrates for the synthesis of new proteins (Shil and Pal 2017a, b). Heavy metals are the potential agents for induction of abnormal glycosylation inside the tissue (Ramamurthy et al. 2016). Irregular glycosylated or post-translated protein molecules lose their biochemical capability to perform programmed function of cellular integrity and conjugation with other plasma membrane forming matrix (Brockington et al. 2001; Peharec-Stefanic et al. 2012). It is reported that there is an appreciable decline in different biochemical constituents in various tissues of freshwater fish under Cr(VI) stress (Vutukuru 2005). A group of workers (Kori-Siakpere et al. 2006) observed that the plasma protein was lowered in Cr(VI)-treated animals exposed to subacute dose. Excretion of protein through urine as a consequence of renal tubular damage may be suggested for lowering serum protein concentration following chromium intoxication. This is supported by the fact that subcutaneous Cr(VI) treatment to rat stimulated urinary excretion of protein, creatinine, and urea nitrogen (Kim and Na 1991).

Effect of Cr(VI) on Proteolytic Enzyme Activities It is established that proteins and proteolytic enzymes are very much sensitive to the heavy metal poisoning (Jacobs et al. 1977), and Cr(VI), being one of them, is widely suspected to impose organ toxicity, genotoxicity, chromosomal aberration, mutational changes, DNA-DNA cross-strand, etc. which may prevent enzyme formation

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Table 2 Changes in the protein metabolic parameters due to Cr(VI) intoxication Protein metabolism Hexavalent chromium (Cr(VI)) intoxication

Tissue protein content decreased in the liver and kidney of mice and in fish muscle (Shil and Pal 2017a, b; Vutukuru 2005) Pronase and cathepsin activities reduced in the kidney of mice; decreased protease activity in fishes (Shil and Pal 2017b; Tulasi and Rao 2013) Trypsin activity increased in the renal tissue of mice (Shil and Pal 2017b) Free amino acid nitrogen level increased in the kidney tissue of mice (Shil and Pal 2017b) Tissue transaminase activity increased in mice and aquatic animals (Shil and Pal 2017b; Satyaparameshwar et al. 2006) Post-translation modification of protein molecules (Bagchi et al. 2001; Abreu et al. 2014) Alteration of cellular protein structure (Mishra and Mohanty 2008, 2009) Rearrangement of proteomics profile (Guo et al. 2013)

or may enhance depletion of proper metabolic intermediates from the respective tissue (Zhitkovich 2011). In this connection, decreased proteolytic enzyme activities such as trypsin, cathepsin, and pronase have been reported in chromium-stressed mice renal tissue (Shil and Pal 2017b). Heavy metal toxicity can lead to alteration of the structure, permeability, and integrity of cell membranes resulting in diffusion of their enzymes outside the cell (Sternlieb and Goldfischer 1976). Decreased proteolytic enzyme activity in the renal tissue by Cr(VI) may be ascribed to less availability of substrate or defective enzyme synthesis. Alteration in physicochemical properties of proteins may involve excess production of reactive oxygen species, and Cr(VI), being a free radical generator (Quinteros et al. 2007), may contribute to them. These in turn may attribute to reduced level of desired tissue proteins for pronase and cathepsin actions. Over-deposition of Cr(VI) in the target tissue may cause symptomatic damage and negative expression of cellular and biomolecular functions in the specific organ and systems (Table 2).

Effect of Cr(VI) on Transaminase Enzyme Activities Aminotransferases contribute an important role in amino acid catabolism and play a key role in nitrogen metabolism and energy mobilization (Calabrese et al. 1977). Transaminases such as GPT and GOT activities in the liver and kidney were markedly increased in subacute Cr(VI) toxicity (Shil and Pal 2017a, b). This may be due to increased accumulation of free amino acid nitrogen in those tissues which may contribute more substrates to compensate hypoglycemia as well as for the activity of transaminases. Increased transaminase activity in Cr(VI) toxicity in specific organ is also reported in earlier studies (Soudani et al. 2012; Kim and Kang 2016) reflecting adverse effects of Cr(VI) at tissue level. Earlier report suggested that leakage of tissue transaminase to the serum due to damage of the

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affected tissue elevated its level in serum sample of the exposed animals (Kim and Na 1991). Cell membrane damage of the exposed tissue by Cr(VI) involves overproduction of lipid peroxides and oxidative stress-mediated degeneration of cellular biomolecules (Bagchi et al. 1995, 1997).

Effect of Cr(VI) on Serum Lipid Profile and Fat Metabolism Cr(VI) appears to have notable effects on lipid metabolism in exposed animals. It has been already established that heavy metal induces the genes responsible for synthesis of the liver enzymes producing cholesterol (Vinodhini and Narayanan 2009; Harabawy and Mosleh 2014). Cr(VI) exhibits its cytotoxic effect by upregulating cholesterol-synthesizing enzymes resulting in increased cholesterol level in cells (Guo et al. 2013). Accumulation of cholesterol and triglyceride in chromiumexposed tissue is supposed to be involved in fatty infiltration of that tissue. This is supported by the morphological studies of the liver carried out in our laboratory which indicated distinct steatosis characterized by overaccumulation of fat in the liver (Shil and Pal 2017a). In continuation with these studies, very recent observation of ongoing research in our laboratory reveals that the activity of fatty acid synthase is stimulated upon subacute chromium exposure. This may result in synthesis of new fatty acids thus enhancing fat depot within the hepatic cell. Not only these, it may also aid lipid substrates for replenishment of energy deficit caused by depletion of hepatic glycogen and protein following chromium intoxication. Additionally, fat deposition in hepatic tissue may be a compensatory mechanism in response to chromium-induced oxidative stress-mediated degeneration of cellular lipids in the form of lipid peroxides. Heavy metal toxicity increased the level of cholesterol, LDL, VLDL, and triglyceride contents in the treated animals that altered the lipid metabolism associated with declined activity of lipoprotein lipase in the hepatic tissue (Yang et al. 2013). High triglyceride content with associated increased mRNA expression of glycerol-3-phosphate acetyltransferase is also noted in heavy metal intoxication (Larregle et al. 2008). Another suggestive mechanism is that heavy metal like Cr(VI) may induce increased expression of fatty acid synthase and also stimulate isocitrate dehydrogenase activity in the liver tissue to promote lipogenesis as a compensatory mechanism of carbohydrate and protein depletion in Cr(VI) intoxication. Moreover, hypoglycemia generally stimulates the secretion of cortisol that causes breakdown of fat in adipose tissue and mobilizes free fatty acids to the liver, thus promoting the synthesis of triglyceride and cholesterol (Wang et al. 2012); and Cr(VI), being a hypoglycemia-inducible factor, may behave like this (Table 3). Further investigation recently carried out in our laboratory reveals that serum lipid profile is significantly disturbed by subacute hexavalent chromium exposure in mice model. The findings reveal that total cholesterol, triglyceride, and LDL cholesterol contents of serum are significantly elevated after Cr(VI) intoxication, whereas Cr(VI) causes a decrease in the level of HDL cholesterol in serum of mice. The present study is partially in compliance with earlier

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Table 3 Alteration in fat metabolic parameters and Cr(VI) accumulation due to Cr(VI) intoxication Fat metabolism and bioaccumulation of Cr(VI) Hexavalent chromium Total cholesterol increased in serum (Kumar and Barthwal 1991; intoxication Soudani et al. 2011) Serum LDL and triglyceride level increased (Soudani et al. 2011; Yousef et al. 2006) Serum HDL level decreased (Soudani et al. 2011) Lipid peroxidation increased in the liver and brain (Huang 1999; Pandey et al. 2005) Fatty infiltration in the hepatic tissue producing steatosis (Shil and Pal 2017a) Enhanced accumulation of chromium in the kidney of mice and muscle of aquatic organism (Shil and Pal 2017b; Loumbourdis et al. 2007)

findings (Soudani et al. 2011). Decreased HDL cholesterol in Cr(VI) toxicity indicates impaired reverse cholesterol transport from blood to the liver which is thought to promote the development of atherosclerosis probably due to lack of defensive effect of HDL to decrease oxidation of other lipoproteins. The increased triglyceride level in serum may be due to increased synthesis of fatty acids in the liver to meet the demand of energy in case of glucose- and protein-deprived situation imposed by Cr (VI). This is supported by the fact that Cr(VI) toxicity is found to stimulate the activity of fatty acid synthase that may contribute to excess accumulation of fatty acids in hepatic tissue resulting in enhancement of lipid synthetic machinery. All these metabolic dysfunctions imposed by chromium have significant impact on metabolic homeostasis that can lead to functional disorder in different organs.

Summary Environmental pollutants are the major cause of serious health complications. People are exposed to those pollutants from different sources like air, water, soil, industries, foods, etc. Certain metals, heavy metals, and metalloids behave as toxic elements when these are contaminating atmosphere and simultaneously produce adverse effects on plant and animal kingdom. Chromium, being a heavy metal, is found to have some impact on human health. As a trace element, the trivalent chromium is needed for function of specific enzymes of carbohydrate and fat metabolism in human, but the hexavalent chromium due to its easy penetrating ability within cell produces certain effects on cellular metabolism and function. Four major sources of chromium have been identified so far: welding, tanning, chrome plating, and industrial effluents. People engaged in those works are being regularly exposed to chromium compounds and may suffer from some common adverse symptoms like skin irritation and rashes, allergic infection of nasal epithelia, and respiratory distress. However, common man may get exposed to this noxious element mainly through food chain and drinking water contaminated with

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chromium. Deposition of chromium within body over the time disturbs normal physiological attribution via alteration of certain parameters of cellular metabolism and bioenergetics. Morphophysiological alteration by hexavalent chromium has been documented in mice model for short-term exposure. Fatty inflammation along with steatosis due to overaccumulation of Cr(VI) in the hepatic tissue is one of the remarkable adverse changes of hexavalent chromium in mice model. In intracellular compartment, Cr(VI) imposes excess production of free radicals and different stress molecules which may lead to form cellular hypertrophy resulting in mild moderation of the organo-somatic indexes among the liver and kidney. The observations revealed earlier illustrate that carbohydrate, protein, and fat metabolic profiles in the liver, kidney, and blood are severely affected by hexavalent chromium exposure. Hypoglycemia is one of the prominent features of Cr(VI) toxicity as evidenced from the earlier study. This might be due to renal glycosuria or impaired nephritic absorption of glucose with altered morphophysiology of the renal tissue. Plunging of muscle and liver glycogen and pyruvate content assures enhancement of glycogen breakdown as well as exhaustion of carbohydrate metabolites from the affected tissue. Decreasing trend of pyruvate indicates retardation of the glycolysis in all the Cr(VI)-exposed tissues. Altered activities of LDH in the hepatic tissue describes the toxic manifestation of Cr(VI) owing to reduced glycolysis with the less production of pyruvate. Increased LDH activity in the renal tissue promotes anaerobic conversion of pyruvate to lactate which may contribute to energy deficit from carbohydrate source. Cr(VI) employs considerable effect on cellular energy generation in different organs including the liver and kidney through domination on the vital metabolic processes like TCA cycle and gluconeogenesis and also by altering protease and oxidative phosphorylation enzymes. It significantly altered intermediatory by-products of various metabolic pathways and seriously hampered the normal physiochemical attribution of specific cell. Moreover, Cr(VI)-induced alteration of proteases in the abovementioned tissue may be either due to depletion of specific substrates or due to changed structural conformity of the enzymatic protein by the toxicant. Lipid profile among the experimental organism was drastically changed that showed a derogative expression in terms of altered fat metabolites and accessory parameters such as LDL, HDL, cholesterol, and triglycerides. To compensate energy deficiency in the hepatic tissue, the fatty acid synthesis was triggered by the fatty acid synthase enzyme and enhanced lipogenesis as indicated by increased IDH activity in that tissue as well as overproduction of cholesterol and triglycerides in serum samples of mice after chromium exposure. In response to these, the reverse transport of cholesterol from blood to the liver was hampered due to less production of HDL cholesterol in Cr(VI) toxicity. Overall it is suggested that Cr(VI) exclusively deteriorates cellular energy generation; collapses homeostatic interrelation of carbohydrate, protein, and fat metabolic pathways; deregulates metabolically linked enzymes; and seriously hampers the structural and functional integrity of the liver and kidney that are supposed to be responsible for their functional imbalance in response to excess chromium exposure (Shil and Pal 2017a, b).

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Tewari H, Gill TS, Pant J (1987) Impact of chronic lead poisoning on the hematological and biochemical profiles of a fish, Barbus conchonius (Ham). Bull Environ Contam Toxicol 38:748–752 Trivedi B, Saxena DK, Murthy RC, Chandra SV (1989) Embryotoxicity and fetotoxicity of orally administered hexavalent chromium in mice. Reprod Toxicol 3:275–278 Tulasi G, Rao KJ (2013) Effect of chromium on protein metabolism in different tissues of fish, Cyprinus carpio. Res J Pharm Biol Chem Sci 4:143–148 Vauquelin LN (1798) Memoir on a new metallic acid which exists in the red lead of Siberia. J Natural Phil Chem Art 3:146 Velma V, Tchounwou PB (2013) Oxidative stress and DNA damage induced by chromium in liver and kidney of goldfish, Carassius auratus. Biomark Insights 28:43 Venugopal NB, Reddy SL (1993) In vivo effects of trivalent and hexavalent Cr(VI) on renal and hepatic ATPases of a freshwater teleost Anabas scandens. Environ Monit Asses 28:131–136 Vinodhini R, Narayanan M (2009) The impact of toxic heavy metals on the haematological parameters in common carp Cyprinus carpio (L.) Iran J Environ Health Sci Eng 6:23–28 Vutukuru SS (2005) Acute effects of hexavalent chromium on survival, oxygen consumption hematological parameters and some biochemical profiles of the Indian major carp, Labeo rohita. Int J Environ Res Pub Health 23:456–462 Wang H, Oster G (1998) Energy transduction in the F1 motor of ATP synthase. Nature 396:279 Wang JC, Gray NE, Kuo T, Harris CA (2012) Regulation of triglyceride metabolism by glucocorticoid receptor. Cell Biosci 2:19–25 White LA (1943) Energy and the evolution of culture. Am Anthropol 45:335–356 Xiao F, Feng X, Zeng M, Guan L, Hu Q, Zhong C (2012) Hexavalent chromium induces energy metabolism disturbance and p53-dependent cell cycle arrest via reactive oxygen species in L-02 hepatocytes. Mol Cell Biochem 371:65–76 Yadav IC, Devi NL, Zhong G, Li J, Zhang G, Covaci A (2017) Occurrence and fate of organophosphate ester flame retardants and plasticizers in indoor air and dust of Nepal: implication for human exposure. Environ Pollut 229:668–678 Yang J, Liu D, Jing W, Dahms HU, Wang L (2013) Effects of cadmium on lipid storage and metabolism in the freshwater crab Sinopotamon henanense. PLoS One 8:1–10 Yousef MI, El-Demerdash FM, Kamil KI, Elaswad FA (2006) Ameliorating effect of folic acid on chromium (VI)-induced changes in reproductive performance and seminal plasma biochemistry in male rabbits. Reprod Toxicol 21:322–328 Zhitkovich A (2011) Cr(VI) in drinking water: sources, metabolism, and cancer risks. Chem Res Toxicol 24:1617–1629

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Contents Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Life Cycle of Nanoparticles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Mode of Entry of Nanoparticles and their Physicochemical Properties . . . . . . . . . . . . . . . . . . . . . . . Size of Au-NPs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Charge on Au-NPs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Shape of Au-NPs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Protein Corona and its Role in Internalization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Capping and Functionalization of Au-NPs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cytotoxicity of Au-NPs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Effect of Au-NPs on Skin Cells . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Effect of Au-NPs on Hepatocytes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Effect of Au-NPs on Renal Cells . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Effect of Au-NPs on Intestinal Cells . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Effect of Au-NPs on Endothelial Cells . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Effect of Au-NPs on Skeletal Muscle Cells . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Effect of Au-NPs on Fibroblast Cells . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Genotoxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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S. P. Bitragunta Department of Biological Sciences, BITS Pilani, Hyderabad Campus, Hyderabad, Telangana, India Biotechnology Division, Environment Protection Training and Research Institute, Hyderabad, Telangana, India e-mail: [email protected] S. Aarathi Menon · P. Sankar Ganesh (*) Department of Biological Sciences, BITS Pilani, Hyderabad Campus, Hyderabad, Telangana, India e-mail: [email protected]; [email protected] © Springer Nature Switzerland AG 2019 C. M. Hussain (ed.), Handbook of Environmental Materials Management, https://doi.org/10.1007/978-3-319-73645-7_59

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Abstract

Application of engineered nanoparticles (ENPs) in biomedical and industrial applications is gaining prominence across the globe. The unique properties of ENPs, such as size, shape, charge, etc., facilitate their specific cellular and subcellular interactions. Among the emerging ENPs, the exceptional characteristics of gold nanoparticles (Au-NPs) including biocompatibility, facile synthesis, and optical properties make them ideal candidate particles for biocatalysis, imaging, and drug delivery. However, widespread use of Au-NPs may result in adverse impacts on environment and health. Thus, toxicity assessment starting from their manufacture to end of the life cycle is an important requisite for risk assessment of Au-NPs. Moreover, understanding the cellular interactions of Au-NPs assists in assessing biochemical effects and establishing methods of toxicity assessment. Therefore, this chapter is aimed at providing an account on cellular interactions of Au-NPs with an emphasis on the role of protein corona in nanoparticle uptake by cells. The chapter highlights the importance of physicochemical properties in evaluating the toxicological profile of Au-NPs. Furthermore, the chapter focuses on effects of Au-NPs on different types of cells such as renal and dermal cells. Thus, main findings of the study will help in divulging cytotoxicity and associated biochemical mechanisms of Au-NPs. Keywords

Gold nanoparticles · Physicochemical properties · Cytotoxicity · Protein corona

Introduction Miniaturization of materials at nanoscale resulted in extensive growth of nanotechnology over the past two decades, thereby leading to widespread application of nanoparticles (NPs) in almost all sectors including cosmetics, therapeutics, and electronics. The size of the nanoparticles gives it a large surface area thereby rendering them more reactive than their bulk counterparts. Among the different NPs available in the market, gold nanoparticles (Au-NPs) are widely used because of their unique properties. The history of application of Au-NPs dates back to fourth century AD, followed by the use of these Au-NPs in medicinal preparation of Swarna Bhasma in seventh century AD (Paul and Sharma 2011) (Fig. 1). Use of Au-NPs in cancer therapy is increasing due to its ability to absorb near-infrared rays. It produces hyperthermic effect on tumor cells due to its enhanced permeability and retention effect (EPR) thereby accumulating more in tumor tissue than in normal tissue (Kumar et al. 2013; Wei 2008). Increased surface area of Au-NPs renders them as the best vehicle for delivery of large biomolecules into cells (Ghosh et al. 2008). Due to their unique optical property, they are also used in biosensors for detecting heavy metals in polluted water (Li et al. 2010; Darbha et al. 2008). They are also used in pollution control for carbon monoxide removal. Colloidal Au-NPs gives

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7th century AD preparation of Swarna Bhasma

17th century AD – used in treating fever 10th – 16th century and syphilis AD – used in Medieval stained glasses

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20th - 21 century AD – Growing application in biomedical field, electronics field, in paint industry, cosmetics etc.

4th century AD Lycurgus Cup was made by the Romans

Fig. 1 The advent of the use of Au-NPs starting from fourth century AD to twenty-first century AD (Reference: Melo 2008; Paul and Sharma 2011; Wilson 2008)

different optical properties depending on size; e.g., 25 nm is red in color, 50 nm is green in color, and 100 nm is orange in color. Hence the mode of interaction of these NPs with various components in the environment will also differ with size. To date, there are different consumer products that are incorporating Au-NPs (Wilson center database – http://www.nanotechproject.org/cpi/). As per the database, Au-NPs were not incorporated in consumer products in 2006; however by 2011 the use of Au-NPs in various commercial products gained momentum. As per the last updated version of the repository in the year 2013, around 19 consumer products that are employing Au-NPs are recorded. However there is no data regarding current status of the use of Au-NPs in consumer products, but the repository includes only those products that claim to use nanotechnology either by the manufacturer or by any third-party source. Thus, there would be many other products in the market with no description about the use of NPs. With increasing production of different-sized Au-NPs and their application, the concern about their impact on the environment and human health is gaining attention across the globe. Though Au-NPs is the least disputed nanoparticle in terms of adverse impact on environment and health, the concern increases due to its increased reactivity at the nanoscale and thus the ability to interact with different components in the environment (Dwivedi and Ma 2013). Therefore, it is important to study how these NPs interact and enter the cells. In this regard, this chapter highlights the possible pathways/modes of Au-NPs interaction with cells and the way forward. Interesting questions such as how does the physiochemical property of the Au-NPs affect the

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mode of entry? and what are the consequences after Au-NPs enter the cell? are the starting point for the present work to shed light on recent advances in toxicology of Au-NPs.

Life Cycle of Nanoparticles See Fig. 2.

Mode of Entry of Nanoparticles and their Physicochemical Properties Understanding the mode of entry of nanoparticles is important to assess their toxicity (Chithrani and Chan 2007). They can enter the cell either by receptor-mediated endocytosis like clathrin-mediated endocytosis or caveolae-mediated endocytosis or phagocytosis (Shang et al. 2014). However, endocytosis is the major mode of entry into phagocytes and endothelial cells or in the presence of opsonins; otherwise, the NPs with size 90%). The fine size of its particles makes it possible to extend the granulometry of the whole granular skeleton of the cement by filling the voids and increasing the compactness. Moreover, its amorphous structure makes it possible to trigger a pozzolanic reaction by the consumption of the lime and the creation of new hydrates. The standard NF P 18–502 distinguishes two classes of silica fumes, A and B; the silica fumes of class A being the richest in silica and are finer. Class A or B silica fume according to the standard EN 206–1 is an addition of type II and is cement substitutable in the sense and under the conditions of this standard. However, given the very high fineness of these additions and their very high reactivity with portlandite released by the hydration of the cement, their proportion is limited to 10% and their use reserved for concretes containing a superplasticizer.

Fly Ash [NF EN 450-1] Fly ash is fine powders consisting mainly of spherical vitreous particles derived from the combustion of pulverized coal in the presence or absence of co-fuels, having pozzolanic properties and consisting essentially of SiO2 and Al2O3; the proportion of reactive SiO2 constituting at least 25% of the mass. In other ways, fly ash is fine particles resulting from the combustion of coal in thermal power plants. Their particles have a spherical shape with a diameter ranging from 1 to 150 μm. Fly ash conforming to the norm NF EN 450–1 is type II additions in accordance with the standard EN 206–1. These additions are substitutable for cement according to the conditions of this standard.

Natural Pozzolan [ASTM C 618] Natural Pozzolan is a natural rock corresponding to volcanic, scoriaceous, essentially stromboli and basic projections – that is to say, a basaltic composition; its color generally varies from red to black according to the degree of oxidation of iron, present respectively as hematite or magnetic. This natural material may have been calcined in an oven and processed, and then ground to obtain a fine powder. It is exploited for the production of composite cement and is found in different varieties; most commonly used at present includes calcined clay, calcined schist, and metakaolin. This Pozzolan is rich in silica SiO2 and in alumina Al2O3 capable of reacting with lime in the presence of water to form, at the end of this reaction, products exhibiting binding properties. Natural Pozzolan conforming to ASTM C 618 are

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Type II additions as defined in EN 206–1/ASTM C 618 and are substitutable for cement according to the terms and conditions of this standard.

Blast Furnace Slag [NF P 18-506] Blast furnace slag is a by-product of iron and steel mills consisting primarily of silicates and calcium, aluminosilicates, and other minerals. In other words, it is a coproduct of the manufacture of cast iron and obtained by quenching the molten blast furnace slag. Once removed from the furnace, it undergoes rapid cooling with a jet of water, which gives it a vitreous structure capable of reacting in the presence of the calcium hydroxide liberated by the hydration of the cement. The standard NF P 18–506 distinguishes two classes of slag A and B; the last one is the most reactive, its fineness being higher. Slags conforming to standard NF P 18–506 are type II additions in accordance with EN 206–1, but only Class B slags are substitutable for cement in the sense and according to the requirements of this standard.

Action of Mineral Additions on Cementitious Materials Mineral additions due to their fineness and reactivity, which are more or less important in the presence of cement, engender significant changes in the physical properties of fresh cement paste and the mechanical performance of mortars and/or concrete in the hardened state. In the fresh state, these additions modify the structure of the granular skeleton (granular effect) and the friction between the solid components in the liquid phase. In the same way, during the setting and hardening, the additions particles interact in the cement hydration process while modifying the structure of the hydrated products and for some may react chemically (physical, chemical, and microreactors effects) to the cementitious media in order to form new hydrated products which have an additional binding character. The mechanisms of these modifications appear to be particularly complex; however, several recent studies (Hemalatha et al. 2016; Nécira et al. 2017; Özbay et al. 2016; Ramezanianpour 2014; Wang et al. 2015) agree to distinguish three main effects of additions in the formulation of a cementitious material. • A granular effect resulting from the modifications made by the addition of the granular structure of the material in the presence of water and optionally of admixtures, which acts on the rheological properties and the compactness of the cementitious materials in the fresh state. • A physical, chemical, and microstructural effect generated by the multiple interactions between the particles of the addition and the process of hydration of cement, which acts on the evolution of the hydration of cement during the setting and the hardening. • A purely chemical effect specific to certain additions in the cementitious medium, which acts during the hydration of cement and which interacts strongly with the physical, chemical, and microstructural effect.

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Granular Effect

Fig. 1 Reduction of water demand brought by introduction of two fly ash in a concrete formulation according to Lewandowski (1983)

Demande en eau (kg/m3)

The granular effect, also known as the filler effect, relates to all changes induced by the presence of mineral additions in the granular skeleton of a fresh cement material. These modifications can result from the stacking capacity of the fine or ultrafine particles of the addition with the other solid grains of the mixture and/or the intensity of the friction between the various particles of the mixture. This granular effect may be favorable or unfavorable due to several factors including morphology, textural surface, granular distribution, and zeta potential of the addition particles used, which influence the rheology of cementitious materials (Benabed et al. 2016; Bingöl and Tohumcu 2013). It acts as soon as the cementitious mixture is kneaded and in all the processing steps on one hand and on the other hand it influences the density of the granular skeleton as well as the stability of the mixtures in the fresh state. When the particles of addition slightly modify the intergranular frictions and succeed in filling the porosities of the granular skeleton (cement and aggregates) by releasing water contained in these pores, the granular effect becomes favorable either by improving consistency of the mixture in the fresh state while keeping the quantity of water constant or either by reducing the quantity of water for a given consistency while improving the compactness of the mixture on one hand and the mechanical performances in the hardened state on the other hand. Several studies have shown the existence of an optimization of the properties of a granular skeleton by incorporation of mineral additions of different natures and quantities. Lewandowski (Lewandowski 1983) has shown that the substitution of part of cement with fly ash in a concrete formulation leads to the progressive reduction of the water content for same consistency because of the nonporous spherical character of the ash particles (Fig. 1). In the same context, Lange et al. (1997) concluded that for a given fluidity, the introduction of a specific quantity of fly ash reduced the amount of water and increased the fluidity of the mixture. This behavior was explained by the spherical 210

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shape of the particles that facilitate granular stacking and reduce interparticle friction (Svatovskaya et al. 2016). The ratio of the surface area of the particle volume also has been reduced by the spherical shape of the particles, which subsequently minimizes the water demand in the mixture (Benabed et al. 2016). The fineness of the addition particles can also be a factor influencing the rheology of fresh cement materials. Indeed, Collins and Sanjayan (1999) reported that in concrete, the fluidity of the mixture was improved by replacing a part of cement by ultrafine additions. Liu et al. (2000) also showed that incorporation of an ultrafine fly ash addition with a surface area of 740 m2/kg improved fluidity and reduced the water requirement of the normal consistency mixture. Similarly, Kronlof (1994) studied the effect of incorporating three additions of quartz of different fineness into the concrete formulation and showed that the effect of ultrafine aggregates on the strength of the concrete leads to decrease in the need of water. This decrease is related to the dosage and the fineness of the substitute material. In the same way, Kwan (2000) studied the influence of the use of silica fume in the manufacture of high strength and self-compacting concrete and showed that the workability of concrete increases with increasing percentage of silica fume for a constant water/ binder ratio (Fig. 2). Similarly, De Larard et al. (1986) studied the effect of silica fume on the improvement of mortars and concrete; they showed that the maneuverability of a mortar varies with the amount of incorporated silica fume (Fig. 3). From the same mine, Yijin et al. (2004) found that the incorporation of fly ash into cement, mortar, and concrete pastes can improve the flow of mixture, but some coarser additions could not reduce the water requirement. Zhang and Han (2000) studied the effect of ultrafine additions on the rheological properties of cement pastes and showed that the yield stress increased with increasing quantity, but the viscosity of the dough varied with the nature and amount of addition. When the rate of substitution of cement by silica fume, fly ash, or limestone additions is less than 15%, the viscosity of the pulp is remarkably reduced. This has not been noticed for slag additions. Indeed, when the particles of the addition considerably modify the granular interstitial frictions in cementitious mixtures or fail to fill the porosities of the granular skeleton, the granular effect, in this case, becomes unfavorable. 300 250

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Gallias et al. (2000) have studied the effect of fine mineral additions on the water requirement of cement pastes and have shown that the introduction of highproportion addition in cement pastes with a standardized consistency leads to a higher water requirement (Fig. 4). Bessa (2004) studied the contribution of mineral additions to the physical, mechanical properties, and durability of mortars; this author, showed that the granular effect of mineral additions on the formulations of without admixtures mortars depends primarily on the fineness and quantity of addition introduced (Fig. 5). The analysis of the various results makes it possible to conclude that the granular effect of mineral additions can be favorable or unfavorable on the behavior of fresh cement materials. This effect depends on several factors characterizing the mineral additions, including nature, quantity, fineness, and morphology, by subsequently influencing the rheology of cementitious materials. Other minerals have an influence on the tightening of the granular skeletal particles and on the hydration processes of the cement that is to say on the development of the physical, chemical, microstructural effects, and chemical properties.

Physical-Chemical and Microstructural Effect The physical-chemical and microstructural effect concerns all the modifications generated by the multiple interactions between the particles of mineral addition on the process of hydration of cement on one hand and on the other hand on the structuring of the hydrated products.

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Fig. 4 Water requirement for CEM II cement pastes according to the proportion of mineral additions, according to Gallias et al. (2000)

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Many authors are unanimous that the introduction of mineral additions into a cementitious mixture modifies the cement hydration process independently of the nature of addition (Berodier and Scrivener 2014; Chuah et al. 2014; Fernández et al. 2016; Luo et al. 2013). Lilkov (1997) showed that the quantity of hydrates formed during the first 24 hours in a cement paste formulated by addition of silica fume and fly ash is greater than that of controls (without additions). Similarly, Hanna (1987) has shown that the incorporation of fine and ultrafine particles of different percentages of silica fume addition into a cementitious material

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modifies the amount of heat released during the first 24 hours. These results were confirmed by Jiang and Van Damme (1996) by studying the action of siliceous and limestone additions on the hydration process of pure C3S. In the same context, Lawrence et al. (2000) also studied the effect of mineral additions; chemically inert, on the hydration of mortars, they showed that the degree of hydration at short-term of mortars containing these types of inert additions was always greater than that of reference mortars without additions. They thus confirmed that these additions improve the hydration of this type of cement (Fig. 6). a. During the first 48 hours at different rates of cement substitution by addition in case of quartz Q24 b. At 1 and 2 days for different rates of cement substitution by the addition in case of quartz of different fineness They also identified the two main physical effects responsible for the hydration of cement and showed that the heterogeneous nucleation, which increases with the fineness of addition, presents an optimum as a function of the rate of substitution of cement. In the same way, Cyr et al. (2005) showed that the incorporation of mineral additions physically influence in the hydration of mortars at short term and that for given an addition this influence depended on the amount of addition incorporated (Fig. 7). They also showed that the mineral adduct particles influence the hydration kinetics only when they are close of cement grains and that this was only possible when the quantity of quartz in the mixture remains small. In the microstructural and microchemical domain, Ramlochan et al. (2003) have shown that the incorporation of Pozzolan and slag in a cementitious material slightly modified the internal product of alite phase compared to the controls (Fig. 8). However, sulfates and aluminates were proportioned differently among the liquid pores and solid phases on one hand, and on the other hand, the microstructure of the cementitious material was modified.

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Fig. 8 Expansion of a standard mortar containing Pozzolan or slag, stored at 95  C, according to Ramlochan et al. (2003)

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Several hypotheses have been put forward to explain the action of additions on the process of hydration of cement. Some authors consider that the presence of mineral additions multiplies the possibilities of germination of the hydrated products of cement and thus facilitates the formation of a solid structure guaranteeing the first mechanical strengths (Adekunle et al. 2015; Colangelo et al. 2015). Other authors explain that the presence of mineral additions (fillers of limestone and/or cementitious dust) in a cementitious material results in an increase in water/ cement ratio and leads to the acceleration of hydration process (Khudhair and Elharfi 2016; Khudhair Mohammed Hussein et al., 2017c). However, the same authors in other works explain that the presence of mineral additions, such as natural Pozzolan and/or combination of natural Pozzolan and limestone fillers in a cementitious material, leads to a decrease in water/cement ratio and leads to slowing and/or decelerating the hydration processor allowing better dispersion of cement grains while leading to a more efficient structuring of the cementitious matrix (Khudhair et al. 2017a, e).

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Today, it is no longer necessary to demonstrate that mineral additions play the role of preferential nucleation sites during cement reactions, allowing a better distribution of hydrated products and thus leading to a more efficient structuring of cementitious matrix (Khudhair et al. 2017a; Mardani-Aghabaglou et al. 2016; Senhadji et al. 2014). For limestone additions, it appears that the presence of calcium carbonate (CaCO3) favors the hydration of C3S from the first moments, the more so because the particles are fine and the quantity of CaCO3 is high (from 15 to 20% in mass). For siliceous additions, quartz particles may constitute preferential nucleation sites especially for the crystallization of portlandite crystals (Alujas et al. 2015; Khudhair et al. 2017a). Thus, the presence of mineral additions provokes an acceleration of the hydration reactions of cement and promotes the properties of the material in the hardened state at the young ages, especially since the particles are fine (Khudhair and Elharfi 2016; Khudhair et al. 2017a). However, this favorable effect seems to fade over time. Indeed, Khudhair (Khudhair et al. 2017e) showed that the use of certain proportions of mineral additions could have a greater retarding effect than the accelerating effect at younger ages. In general, we can conclude that the physical, chemical, and microstructural effect of mineral additions have an important influence on the evolution of mechanical resistance at young ages and on the physical and microstructural properties of cementitious materials in the hardened state.

Chemical Effect While the physical, chemical, and microstructural effect generally concerns all mineral additions independently of their mineralogical nature, the chemical effect is intimately related to their mineralogical composition and relates to the capacity of the additions characterized by Pozzolanic and/or hydraulic to react with water and the anhydrous or hydrated cement constituents to form of a new mineral phases which can contribute to the evolution of the mechanical strengths in the same way as the hydrated cement products. The European standard (EN 206–1) refers the additions chemically active as Type II, taking into account their latent hydraulic activity or pozzolanic activity. Appa Rao (2001) studied the influence of the incorporation of silica fume in a mortar by substituting cement for 30% of this addition and showed that mechanical strengths of compression increase independently of mortar with a constant water/ binder ratio equal to 0.5. Kwan (2000) has also studied the influence of the use of silica fume in manufacture of high strength and self-compacting concrete and has shown that the substitution of cement by 15% of this addition, regardless of the water/binder ratio, leads to an increase in mechanical resistances of compression at long term (28 days). Demirboga (2003) studied the influence of the incorporation of mineral additions, including silica fume, fly ash, and slag, on the mechanical performance of mortars; it showed that the density of mortars formulated by these additions decreased according to the increase in percentage of these additions (silica fume, fly ash, and

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slag) on one hand. On the other hand, he noted that the incorporation of 10% of silica fume resulted in an improvement in mechanical strength at 7, 28, and 120 days. However, the incorporation of 10% of fly ash produced a decrease in resistance mechanical of compression. Similarly, Khudhair et al. (2017c, 2017e) studied the influence of the incorporation of mineral additions, including limestone fillers, natural Pozzolan, and a combination of the two at 40% by weight of cement in a cementitious material; they showed that the density of mortars formulated by these additions decreased as a function of increase in percentage of these additions (limestone fillers and natural Pozzolan) on one hand. On the other hand, they had noted that the incorporation of these additions resulted in an improvement in mechanical strength and durability. Benezet and Benhassine (1999) have studied the influence of quartz particle size in the pozzolanic reaction. They also showed that finely ground crystallized quartz can react chemically with portlandite under certain conditions. However, the quantification of low chemical activity separately from the physical, chemical, and microstructural effect is difficult and uncertain. Lawrence et al. (2005) have studied the effect of nature, the quantity, and fineness of fine mineral additions, such as quartz, limestone, and fly ash, on compressive strengths of mortars; they also showed that for fineness similar to those of cement, and at 7 days, the mortars containing limestone additions had higher compressive strengths than those of mortars formulated by the addition of quartz. This difference became negligible at 28 days for a given cement substitution rate (Fig. 9). These authors had shown that mortars based on limestone and quartz additions had equivalent densities and occluded air volumes, suggesting that the reduction in porosity was not due to the particular behavior of the limestone additions. They argued that it was more logical to think that there are special mechanisms that improve the compressive strengths of mortars formulated by limestone additions at a young age (2 days), but which are negligible in the long term (28 days). 7 days

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This behavior was justified by the formation of aluminates and carboaluminates (Soroka and Stern 1976). The authors also reiterated that the improvement of the mechanical resistance of compression is probably due to the heterogeneous nucleation phenomena observed by several authors (Nalet and Nonat 2016; Nécira et al. 2017). This heterogeneous nucleation is a physical process, generating a chemical activation of the hydration of cement. It depends on several factors among which we find the fineness of particles, the amount of addition, and the affinity of addition powder. A more explicit approach was proposed by Cyr et al. (2006). To quantify the physical and chemical effects of fine additions on mechanical strengths of a mortar containing an amount (p %) of mineral addition is represented as the combination of three effects superimposed on Fig. 10. With: f (dilution) designates to mechanical resistance of compression proportional to the quantity of cement without the physical and chemical effects of mineral additions. Δfφ (Physical) designates to improvement in the mechanical strength of compression due to the physical effect of mineral addition, due to the filler effect and heterogeneous nucleation. Δfpz (Chemical) designates to improvement in mechanical strength of compression which is due to the chemical effect of the mineral addition on one hand and the pozzolanic reaction on the other hand. This approach is consistent with Lawrence’s approach to studying the activity of fly ash and chemically inert mineral additions in cementitious materials (Lawrence 2000).

Fig. 10 Decomposition of mechanical resistance of compression in fractions of physical and chemical effects of mineral addition, according to Cyr et al. (2006)

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The dilution effect (Fig. 11) was explained as the consequence of substitution of cement by the addition. A lower amount of hydrated cement will result in lower mechanical resistance of compression. The physical effect (Fig. 12) is due to the fillers effect and the heterogeneous nucleation. The authors concluded that:

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Fig. 11 Mechanical resistance of compression at 07 days and 28 days of mortars containing quartz additions and fly ash of the same fineness compared to a reference addition, according to Cyr et al. (2006)

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Fig. 12 Improvement of mechanical strength of compression Δfφ at 07 and 28 days of mortars containing increasing quantities of quartz mineral additions, according to Cyr et al. (2006)

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– The heterogeneous nucleation increased with the increase in the fineness of addition and confirmed that the optimum amount of addition was 25% and 35% because beyond this quantity the effect is expected due to the removal of part of the additions to participate in the nucleation of cement hydration process. The chemical effect (Fig. 13) is due to the pozzolanic activity of some additions, which improves the mechanical strength of cement. The authors showed that the chemical effect of natural Pozzolan begins to be significant from 07 days and grows with age. The optimum amount of addition which resulted in maximal pozzolanic activity was 35% and 40%. For limestone additions, Cyr et al. (2006) noted that the limestone addition is not completely inert since it reacts with C3A and C4AF to form carbo-aluminates. They were able to show that the compressive strengths of mortars with limestone additions were higher than those of mortars with quartz additions, but this effect was neglected in comparison with physical effects. Senhadji et al. (2014) have studied the influence of replacement of clinker by natural Pozzolan, silica fume, and limestone on mechanical resistance, acid resistance, and microstructure of mortar. They showed that the incorporation of a natural Pozzolan, silica fume, and limestone in a mortar increases the mechanical resistance of compression and the resistance of chemical attacks in the long term by comparing with the controls. In general, the pozzolanic reaction mainly concerns: – Silica fume, siliceous fly ash (class F), natural Pozzolan, and calcined schists. The amorphous silica present in these different additions reacts in presence of water with the portlandite Ca(OH)2, which is compounded during the hydration of cement to form hydrated calcium silicates C-S-H according to the following reaction: S + x CH + y H CxSHx + y.

Fig. 13 Improvement of mechanical strength of compression Δfpz at 07 and 28 days mortars containing increasing quantities of quartz mineral additions, according to Cyr et al. (2006)

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– Hydraulic activity is more particularly concerned with slags, limestone, and fly ash (class C), which due to the basic nature of cementitious medium can produce HSCs with a different C/S ratio than those resulting from pozzolanic reactions. – Moreover, the limestone additions also show reactivity in presence of hydrated products of cement (Baron and Ollivier 1997). In this case, calcite (CaCO3) reacts with the aluminates of cement (C3A, C4H13) in presence of water to form a calcium mono-carboalumination hydrate of type C3A.CaCO3.11H2O, crystallizing into fine hexagonal plates. However, the chemical activity of limestone additions is significantly lower than that of the pozzolanic siliceous additions. It follows that the chemical effect when it is favorable is complementary to the physical, chemical, and microstructural effect. Its action on properties of the cured material can be quantified by measuring the volume and nature of hydrated formed products. Nevertheless, their strong synergy makes it difficult to distinguish clearly between these two effects and can be associated with a single broader notion which is the contribution of mineral additions of the binding activity of cement (Badreddine-Bessa 2004).

Chemical (Organic) Admixtures as Superplasticizers in the Formulation of Cementitious Materials Effect of Superplasticizer on Rheology Effect of Method of Introduction The time of introduction of a superplasticizer has occupied the minds of several researchers in order to optimize the dispersion effect. Chiocchio and Paolini (1985) have shown that the best time to add a superplasticizer is at the beginning of induction period when any introduction before the first hydration period of C3A results in its broad adsorption by the first hydrates of phase aluminate. The delayed addition effect assumes that the mixture is adsorbed to a lesser extent when added a few minutes after kneading so that there is enough mixture left in the solution to promote dispersion of silicate phases and to lower the viscosity of cement pastes (Yen et al. 1999). Figure 14 shows slump test results where the introduction of superplasticizer was divided into two parts; one half on contacts with mixing water and the other half a few minutes after kneading. The obtained results show that the initial fluidity was considerably increased and that the loss of fluidity was greatly reduced when the second half of superplasticizer was added 3 min after the beginning of mixing. The study of the influence of time of introduction of superplasticizers on the rheological properties of cement pastes shows that the superplasticizers based on naphthalene and melamine increase the rheological properties of cement pastes at short and long term. The reduction of the shear threshold and plastic viscosity depends on the composition of cement and time of introduction of superplasticizer, the optimal time of which was 10–15 min after the start of the kneading process (Kim et al. 2000).

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Fig. 14 Variation of fluidity of a group for several methods of introduction according to Chiocchio and Paolini (1985)

Effect of Superplasticizer Dosage Several researchers (Adjoudj et al. 2013; Khudhair et al. 2017g; Yen et al. 1999) considering that the higher the dosage of superplasticizer and the W/C are high, more the rheological behavior is maintained over time. Similarly, Khalil and Ward (1980) have shown that when we use superplasticizers with very high dosages, this enveloping effect can delay the hydration of cement grains, whatever the nature of superplasticizer, cement, or the average size of its grains. On another hand, the presence of superplasticizer in excess makes it possible of compensating the consumption of the polymer by the grains of cement and its hydrates (Flatt and Houst 2001). Sugamata et al. (2000) studied the influence of various superplasticizers and their dosages on the maintenance of rheological behavior of a mortar. Their results show that increasing the dosage of a superplasticizer no longer influences the fluidity of mortar from a certain value. Similarly, Shindoh and Matsuoka (2003) have also shown that the addition of a superplasticizer helps to reduce the viscosity (Fig. 15) from a certain dosage. This characteristic is now well known, that is to say, the saturation dosage, the dosage beyond which the admixture no longer allows the rheology of the mixture to be significantly modified. Currently, superplasticizers are used at dosages relative to saturation assays in order to limit the phenomenon of loss of rheology over time. Hu (1995) studied the effect of superplasticizer dosage on the rheological parameters of a concrete with constant water dosage. The results presented in Fig. 16 show that the superplasticizer decreases the shear threshold and plastic viscosity. On another hand, its effect on viscosity remains modest after a certain dosage. Effect of Type of Superplasticizer The chemical nature of superplasticizer plays a major role in its adsorption on cement grains. Malhotra and Malanka (2004) found that in order to increase the slump from 50 to 260 mm, it was necessary to add 0.6% of poly-naphthalene

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sulfonates (PNS) or poly-melamine sulfonates, “PMS,” while this could be accomplished with only 0.4% of PNS. Similar results confirmed that the PNS superplasticizer is better than PMS (Yeh 1998). Another comparative study (Uchikawa et al. 1992), of the effect of PNS and lignosulfonate (LS) on properties of pastes made with eight different types of cement, showed that cement pastes containing PNS are more fluid than those containing of the LS due to the strong affinity of PNS of cement grains. By studying the performance of polycarboxylate superplasticizers, Falikman et al. (2005) showed that these superplasticizers provided the same rheological and mechanical performance with dosages of 2.7 to 3.3 times lower than the conventional superplasticizers based on poly-naphthalene. In another study (Uchikawa 1994), it was shown, after kneading with mixing water, that the shear threshold of cement paste containing AS (amino-sulphonic) was the lowest and increased in order with the use of PNS and LS. However, the plastic viscosity of cement pastes with PNS and LS is higher than that of polycarboxylate PC and AS. Lachemi et al. as well as Khayat (Khayat 1998; Lachemi et al. 2004) studied the effect of the several viscosity and thickening agents (polysaccharides of microbial origin such as welan gum, acrylic copolymers) on rheological properties of mortars.

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This category of admixtures increases the apparent viscosity of materials, according to the authors; they reduce the segregation between the various compounds of mortar. A study of the influence of superplasticizers on the rheological behavior of fresh cement mortars using a rotational rheometer was conducted by Golaszewski and Szwabowski (2004). He found that superplasticizers based polycarboxylate were more effective than superplasticizers based naphthalene in improving the rheological properties of mixtures. Other studies of the effect of mineral additions on rheological properties of cement pastes have confirmed that the shear threshold and plastic viscosity decrease when a PNS superplasticizer is used (Park et al. 2005). By measuring the shear threshold of various types of superplasticizer, Husson (1991) found that the addition of superplasticizer reduces the shear threshold and interaction between cement flocs. A reduction in 90% shear threshold was recorded with only 0.2% polycarboxylate (PC), while dosages of up to 8% poly-naphthalene sulfonates (PNS) were necessary to achieve such a decrease as illustrated in Fig. 17.

Effect of Molecular Weight of the Superplasticizer The retarding effect of superplasticizers has been well documented and has been attributed to the capacity of superplasticizer adsorbed on a surface of cement particles and their hydrates. However, the efficiency or severity with which a superplasticizer can delay the rate of hydration varies according to its nature. The influence of molecular weight of superplasticizer on rheological properties of cement pastes is the main parameter in the study of cement-superplasticizers interaction. Uchikawa et al. (1992) have studied the effect of molecular weight of PC superplasticizers (polycarboxylate) on the fluidity of cement pastes. They concluded that

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PC superplasticizers have an optimum molecular weight for fluidity at a given W/C ratio. Using three PCs having different molecular weights, the maximum value of the subsidence is observed for a molecular weight of 25,000, 21,900, and 16,000 g/mol with W/C ratios of 0.5, 0.3, and 0.2, respectively. Andersen et al. (1987) have demonstrated that superplasticizers made of polymers with longer chains have higher dispersion possibilities, while shorter chain polymers have a more pronounced influence on retardation. Khudhair et al. (2017g, f) have studied the influence of a water-reducing superplasticizer and sitting accelerator (SP402)/or retarder (SP103) at various percentages ranging from 0.5 to 4% by weight of cement with a step of 0.5% on physical properties and mechanical performance of mortars and concrete. They showed that the physical properties of fresh cement paste including maneuverability, fluidity, setting (depending on the climatic conditions of its installation in winter and/or summer), the water content, and the mechanical performances, namely porosity, capillary absorption, the mechanical strength of compression, and durability of mortars and/or concrete in the hardened state, have been improved. They concluded that better physical and mechanical properties of the cementitious material formulated by this superplasticizer SP402 are between 0.5% and 3.5% by weight of the cement. Beyond this percentage, the properties decrease considerably. This shows that the 3.5% percentage of this superplasticizer is the saturation point (3.5% SP402). However, they concluded that better physical and mechanical properties with SP104 are between 0.5% and 2.5% by weight of cement. Beyond these percentages, the mechanical properties decrease considerably. This shows that the 2.5% percentage of this superplasticizer is the saturation point (2.5% SP103).

Properties of a Cementitious Material in the Presence of Superplasticizers The addition of superplasticizers to fresh concrete improves its rheological properties. Indeed, in the presence of superplasticizers, the forces of attractive interactions between cementitious particles clamping deprived, this allows them to disperse rather than flocculate. The fluidity of system then becomes greater. In addition, we can use superplasticizers to reduce the amount of mixing water while keeping the consistency constant, which reduces the porosity of concrete and then improves its mechanical properties in the hardened state. An increase in the concentration of superplasticizer relative to the weight of cement further improves its fluidity up to a maximum concentration (saturation state) where from which each additional addition has no significant influence on the fluidity of system. This is the saturation point. At this point, it is not possible to adsorb more superplasticizers on the surface of cement grains. This may be due to the fact that the superplasticizer layer is complete or of steric repulsion too large due of the small distance between the particles. Temkhajornkit and Nawa (2004) have studied the fluidity of cement pastes containing a naphthalene sulfonate superplasticizer. They have shown that the introduction of superplasticizer results in a reduction in flow stress which tends of

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zeta potential. They have also suggested that the adsorption of superplasticizers on the surface of cement particles changes the sign of zeta potential of particle surface which becomes negative (Ramachandran et al. 1998). In this case, the cement particles having the same sign of zeta potential are electrostatically repelled. Aitcin (Saric-Coric and Aïtcin 2003) and Ramachandran (1987) have shown that the greater superplasticizer dosage, irrespective of nature of superplasticizer and type of cement or average size of its grains, the more dispersing effect and retarding effect of the setting are pronounced. Pallière and Briquet (1982) and Hanna et al. (1989) have studied the influence of fineness of cement on the effect of thinners of superplasticizers. They have shown that the higher the Blaine fineness of cement, the lower fluidizing effects.

Combined Effect Superplasticizers and Mineral Additions on Properties of Cementitious Material When the superplasticizer being dosed of saturation, the introduction of silica fume only slightly increases the viscosity (Huynh 1996). On another hand, the viscosities of silica fume containing slurries increase much more rapidly over time than those of the silica-free grouts. Uchikawa et al. (1987) studied the influence of superplasticizers in presence of certain mineral additions on behavior of cementitious mixtures and showed that the addition of naphthalene superplasticizer produced a very large dispersion of all cement particles and mineral addition in aqueous solution and that the mineral additions disperse well without excessive flocculation within the cement paste. Kadri (1998) studied the workability of high-performance concrete and showed that the chemical composition of cement and in particular the C3A content plays a major role on the consistency of concrete by absorbing the molecules of superplasticizers. Kara Ali (2002) has studied the action of superplasticizer on the reduction of water requirement of mortars with additions. He shown that the reduction in the water requirement of mortars increases with the increase of the fluidifying admixutres dosage, this increase is independent of the nature of the admixutres. Khudhair et al. (2017b, d) have studied the combined effect of limestone and various superplasticizers; sitting accelerator (SP402); and/or retarder (SP103) in formulation of mortar/concrete, partially substituting clinker by the limestone fillers (5% to 40% by weight of cement with a step 5%) on the physical and mechanical properties in the fresh and hardened state. Indeed, the results of physical and mechanical analyze of cement formulated by the combination of limestone fillers and two types of superplasticizer (SP402/SP03) showed that up to 40% of limestone fillers added the physical and mechanical properties of new cement materials (composite cement) are suitable according to national and international standards. Other work done by Bessa (2004) he studied the effect of mineral additions on the need of additives; he has shown that the need of mortars and the granular effect of mineral additions depend primarily on quantity and fineness of additive, addition incorporated, independently of the type of cement.

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Some interaction between superplasticizer and various fillers has been observed by Sheinn et al. (2003). In the absence of a superplasticizer, the granularity and geometry of the particles of different mineral additions or different types of cement have an influence on performances of cementitious matrix. In the presence of superplasticizer in this matrix, the rheological properties appear to vary according to the reactivity of particles and affinity between superplasticizer and mineral additions and/or type of cement. Burgos-Montes et al. (Chloup-Bondant 1996) studied the combined effect of limestone and various superplasticizers. They found that the shear threshold of limestone cement decreased with even small superplasticizer dosages, such as 0.16% PC decreases the shear threshold by 78% for cement at limestone and 53% for cement ordinary. The results on fresh BAPs, using the slump flow, J-Ring, and L-Box test, show that the addition of fillers having a large surface area of Blaine improves the fluidity of BAP. Similarly, it appears that fineness of fillers influences the demand of superplasticizers in a significant way. An experimental study was carried out by Saada et al. (1990), where variable fineness cement was recomposed by varying proportions of Portland cement with limestone or siliceous fillers. The results, presented in Fig. 18, show that the shear threshold, measured for several specific surfaces and different superplasticizers, increases with fineness and that efficiency of admixtures depends on the nature of the filler. The slag has interesting characteristics as a mineral additive especially in relation of consistency of its chemical composition (Hinrichs and Odler 1989). According to Park et al. (2005), the replacement part of cement by blast furnace slags in the presence of superplasticizer generally reduces the shear threshold and the viscosity of cement pastes. Grzeszczyk and Janowska-Renkas (2012) have shown that the granulometry of slag has a considerable influence on theology. They have found that slag with fine particles (