Advanced Oxidation Processes for Effluent Treatment Plants 0128210117, 9780128210116

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ADVANCED OXIDATION PROCESSES FOR EFFLUENT TREATMENT PLANTS

ADVANCED OXIDATION PROCESSES FOR EFFLUENT TREATMENT PLANTS Edited by

MAULIN P. SHAH

Elsevier Radarweg 29, PO Box 211, 1000 AE Amsterdam, Netherlands The Boulevard, Langford Lane, Kidlington, Oxford OX5 1GB, United Kingdom 50 Hampshire Street, 5th Floor, Cambridge, MA 02139, United States © 2021 Elsevier Inc. All rights reserved. No part of this publication may be reproduced or transmitted in any form or by any means, electronic or mechanical, including photocopying, recording, or any information storage and retrieval system, without permission in writing from the publisher. Details on how to seek permission, further information about the Publisher’s permissions policies and our arrangements with organizations such as the Copyright Clearance Center and the Copyright Licensing Agency, can be found at our website: www.elsevier.com/permissions. This book and the individual contributions contained in it are protected under copyright by the Publisher (other than as may be noted herein). Notices Knowledge and best practice in this field are constantly changing. As new research and experience broaden our understanding, changes in research methods, professional practices, or medical treatment may become necessary. Practitioners and researchers must always rely on their own experience and knowledge in evaluating and using any information, methods, compounds, or experiments described herein. In using such information or methods they should be mindful of their own safety and the safety of others, including parties for whom they have a professional responsibility. To the fullest extent of the law, neither the Publisher nor the authors, contributors, or editors, assume any liability for any injury and/or damage to persons or property as a matter of products liability, negligence or otherwise, or from any use or operation of any methods, products, instructions, or ideas contained in the material herein. Library of Congress Cataloging-in-Publication Data A catalog record for this book is available from the Library of Congress British Library Cataloguing-in-Publication Data A catalogue record for this book is available from the British Library ISBN: 978-0-12-821011-6 For information on all Elsevier publications visit our website at https://www.elsevier.com/books-and-journals

Publisher: Susan Dennis Acquisitions Editor: Kostas KI Marinakis Editorial Project Manager: Liz Heijkoop Production Project Manager: Joy Christel Neumarin Honest Thangiah Cover Designer: Matthew Limbert Typeset by SPi Global, India

Editor’s biography Maulin P. Shah is currently a research scientist at the Environmental Biology Lab, India. Dr. Shah has served as an assistant professor at Godhra, Gujarat University since 2001. He has more than 160 research publications in highly reputed national and international journals. He is an active editorial board member in 75 highly reputed journals in the field of Environmental and Biological Sciences. He has been appointed as an editor-in-chief of two journals: the Research Journal of Microbiology and the Journal of Biotechnology and Biomaterials. His work has been focused on assessing the impact of industrial pollution on the microbial diversity of wastewater following cultivation-dependant and cultivation-independent analysis. Maulin’s work involves isolation, screening, identification, and genetic engineering of the high impact of microbes for the degradation of hazardous materials.

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Contributors Komal Agrawal Bioprocess and Bioenergy Laboratory, Department of Microbiology, Central University of Rajasthan, Ajmer, Rajasthan, India

J. Anandkumar Department of Chemical Engineering, National Institute of Technology Raipur, Raipur, Chhattisgarh, India

Aditi Banerjee Centre for Nanotechnology, Indian Institute of Technology Guwahati, Guwahati, Assam, India

Jay Bergi Department of Biotechnology, Shree Ramkrishna Institute of Computer Education and Applied Sciences, Surat, Gujarat, India

Navneeta Bharadvaja Plant Biotechnology Laboratory, Delhi Technological University, Delhi, India

Q.Q. Cai Department of Civil & Environmental Engineering, Faculty of Engineering, National University of Singapore, Singapore, Singapore

Jayashankar Das SOA Deemed to be University, Bhubaneswar, Odisha, India

Shivika Datta Department of Zoology, Doaba College, Jalandhar, Punjab, India

Sushma Dave Jodhpur Institute of Engineering and Technology, Jodhpur, Rajasthan, India

S.H. Deng Department of Civil & Environmental Engineering, Faculty of Engineering, National University of Singapore, Singapore, Singapore

Daljeet Singh Dhanjal Department of Biotechnology, Lovely Professional University, Phagwara, Punjab, India

Anurag Garg Environmental Science and Engineering Department, Indian Institute of Technology Bombay, Mumbai, India

Shashwati Ghosh Sachan Department of Bio-Engineering, Birla Institute of Technology, Mesra, Ranchi, Jharkhand, India

J.Y. Hu Department of Civil & Environmental Engineering, Faculty of Engineering, National University of Singapore, Singapore, Singapore

Pyla Jayasree Indian Institute of Technology Bhubaneswar, Argul, Odisha, India

L. Jothinathan Department of Civil & Environmental Engineering, Faculty of Engineering, National University of Singapore, Singapore, Singapore

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Contributors

Sukhmanpreet Kaur Department of Chemistry, Rayat Bahra Institute of Engineering and Nano-Technology, Hoshiarpur, Punjab, India

Prateek Khare Department of Chemical Engineering, Madan Mohan Malaviya University of Technology, Gorakhpur, Uttar Pradesh, India

Prashant Kumar Department of Bio-Engineering, Birla Institute of Technology, Mesra, Ranchi, Jharkhand, India

Vijay Kumar Regional Ayurveda Research Institute for Drug Development, Gwalior, Madhya Pradesh, India

Vineet Kumar Environmental Microbiology and Biotechnology Laboratory, School of Environmental Sciences, Jawaharlal Nehru University, New Delhi, Delhi; Department of Environmental Microbiology, Babasaheb Bhimrao Ambedkar (Central) University, Lucknow, Uttar Pradesh, India

Lal Mohan Kundu Centre for the Environment; Department of Chemistry, Indian Institute of Technology Guwahati, Guwahati, Assam, India

Sandhya Mishra Applied Phycology and Biotechnology Division, CSIR— Central Salt and Marine Chemicals Research Institute, Bhavnagar, India

H.Y. Ng Department of Civil & Environmental Engineering, Faculty of Engineering, National University of Singapore, Singapore, Singapore

S.L. Ong Department of Civil & Environmental Engineering, Faculty of Engineering, National University of Singapore, Singapore, Singapore

Lalit M. Pandey Centre for the Environment; Bio-Interface & Environmental Engineering Lab, Department of Biosciences and Bioengineering, Indian Institute of Technology Guwahati, Guwahati, Assam, India

Ratnesh Kumar Patel Department of Chemical Engineering, Madan Mohan Malaviya University of Technology, Gorakhpur, Uttar Pradesh, India

V. Prashanth Indian Institute of Technology Bhubaneswar, Argul, Odisha, India

Parth Rajput Indian Institute of Technology Bhubaneswar, Argul, Odisha, India

Shristi Ram Applied Phycology and Biotechnology Division, CSIR—Central Salt and Marine Chemicals Research Institute, Bhavnagar; Academy of Scientific and Innovative Research (AcSIR), Ghaziabad, India

Neelancherry Remya Indian Institute of Technology Bhubaneswar, Argul, Odisha, India

Romina Romero Technological Development Unit (UDT), University of Concepcion, Coronel, Chile

Contributors

Arpita Roy Plant Biotechnology Laboratory, Delhi Technological University, Delhi, India

Biju Prava Sahariah University Teaching Department, Chhattisgarh Swami Vivekanand Technical University, Bhilai, Chhattisgarh, India

Monalisa Satapathy Department of Chemical Engineering, National Institute of Technology Raipur, Raipur, Chhattisgarh, India

Maulin P. Shah Research Scientist, Environmental Microbiology Lab, Ankleshwar, Gujarat, India

Ravi Shankar Department of Chemical Engineering, Madan Mohan Malaviya University of Technology, Gorakhpur, Uttar Pradesh, India

Shambhoo Sharan Department of Chemical Engineering, Madan Mohan Malaviya University of Technology, Gorakhpur, Uttar Pradesh, India

Joginder Singh Department of Biotechnology, Lovely Professional University, Phagwara, Punjab, India

Simranjeet Singh Department of Biotechnology, Lovely Professional University, Phagwara; Punjab Biotechnology Incubator (PBTI); RAWTL, Department of Water Supply and Sanitation, SAS Nagar, Punjab, India

Swati Singh Environmental Science and Engineering Department, Indian Institute of Technology Bombay, Mumbai, India

Sushma Rani Tirkey Applied Phycology and Biotechnology Division, CSIR—Central Salt and Marine Chemicals Research Institute, Bhavnagar; Academy of Scientific and Innovative Research (AcSIR), Ghaziabad, India

Ratna Trivedi Department of Environmental Science, Shree Ramkrishna Institute of Computer Education and Applied Sciences, Surat, Gujarat, India

Ayushi Verma Plant Biotechnology Laboratory, Delhi Technological University, Delhi, India

Pradeep Verma Bioprocess and Bioenergy Laboratory, Department of Microbiology, Central University of Rajasthan, Ajmer, Rajasthan, India

Rahul Verma Centre for the Environment, Indian Institute of Technology Guwahati, Guwahati, Assam, India

Aparna Yadu Department of Chemical Engineering, National Institute of Technology Raipur, Raipur, Chhattisgarh, India

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Preface Water treatment technologies such as adsorption or membrane filtration, advanced oxidation processes offer the possibility of the complete conversion of hazardous substances to carbon dioxide, water, and salt-free producing residues. However, one challenge is the combination of advanced oxidation processes with other unit operations to enhance the overall process efficiency. The most significant disadvantages of advanced oxidation processes are the high energy consumption and the production of critical intermediates. New knowledge of oxidant production with higher yields, reactions paths, reactor design, a combination of processes, the determination of dangerous intermediates such as genetic test methods, and water applications has been reused. All these parameters are considered keys to exploiting the potential of advanced oxidation processes for water treatment and reuse. Advanced oxidation processes are purification technologies aimed at degrading and mineralizing unsurpassed organic matter from wastewater by reaction with hydroxyl radicals. Recently, these technologies were proposed as a solution for the treatment of emerging contaminants, in particular pharmaceuticals and personal care products. The advanced oxidation process reactions can further be supported by iron catalysts and UV radiation, resulting in the formation of photopentone systems. The optimum design, operation, and control of these processes can be driven by sophisticated process system design tools that combine advanced oxidation process science, photo-Fenton chemistry, and cutting-edge technology with model-based optimization strategies. However, the use of optimization tools requires the availability of reliable models. The following contents are discussed: The basic principles of oxidation reactions are fundamental for the successful application of advanced oxidation processes in complex matrices such as water and wastewater. The investigations of reaction paths and modeling promote understanding the basics. The Fenton reaction is one of the oldest induced iron oxidation processes and is used worldwide in its entirety. However, there are some drawbacks such as high sludge production, iron leaching, acidic pH, etc. We try to overcome these drawbacks through a better understanding of the process, using reaction promoters and heterogeneous catalysts. Photocatalysis was

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Preface

discovered more than 30 years ago by Japanese scientists. Although a huge number of investigations have been carried out to achieve a better understanding of the process and the reaction pathways, no commercial application is available worldwide. Other catalysts under development with an emphasis on higher photonics efficiency, a shift of the absorption spectrum to the visible light range, and higher photoreactor efficiency could help overcome constraints. Oxidation processes are used in many plants around the world to remove inadequate substances. New applications suggest opportunities for further development. Integrating advanced oxidation processes into processes offers other economic benefits. Combinations with biological processes and in particular membrane bioreactors are continually investigated and supported by toxicological tests. This work is a useful reference for researchers and students involved in wastewater technologies, including analytical and environmental chemistry, chemical and environmental engineering, toxicology, biotechnology, biochemical engineering, liquid waste management, and related fields. It is intended to encourage industrial and public-health scientists as well as decision-makers to accelerate the application of AOPs as technological alternatives for the improvement of wastewater treatment plants.

1 Advanced oxidation processes for complex wastewater treatment Vineet Kumara,b, Maulin P. Shahc a

Environmental Microbiology and Biotechnology Laboratory, School of Environmental Sciences, Jawaharlal Nehru University, New Delhi, Delhi, India. b Department of Environmental Microbiology, Babasaheb Bhimrao Ambedkar (Central) University, Lucknow, Uttar Pradesh, India. cResearch Scientist, Environmental Microbiology Lab, Ankleshwar, Gujarat, India

1

Introduction

Complex wastewater with high COD, BOD, and color is discharged from various industries such as distilleries, breweries, tanning, leather manufacturing, pulp-paper, etc. (Chandra and Kumar, 2015, 2018). Among them, the alcohol distilleries are one of the major sources generating high-strength complex wastewater (Nataraj et al., 2006). A medium-scale distillery using sugarcane molasses, a by-product of sugar production that is the most commonly used raw material, can generate an average of 12–15 L of complex wastewater per liter of alcohol produced (De Vrieze et al., 2014; Kumar and Chandra, 2020). This highstrength complex wastewater is also known as spent wash, stillage, raw effluent, or vinasses. Distillery spent wash is an undesirable viscous, hydrophilic, and residual brown colored liquid waste, which contains high levels of biological oxygen demand (40,000–60,000 mg L
1), chemical oxygen demand (90,000–190,000 mg L
1), and total dissolved solids (90,000– 150,000 mg L
1) with an acidic pH (3.0–4.07) (Acharya et al., 2011; Singh and Dikshit, 2011; Chandra and Kumar, 2017a). Spent wash also contains residual reducing sugars, phenolics, lipids, proteins, amino acids, and volatile organic acids generated by Advanced Oxidation Processes for Effluent Treatment Plants. https://doi.org/10.1016/B978-0-12-821011-6.00001-3 # 2021 Elsevier Inc. All rights reserved.

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Chapter 1 Advanced oxidation processes for complex wastewater treatment

the yeasts during fermentation process. Besides the organic con
2
, potassium (K+), phostent, a high mineral load of sulfate SO 4
2+ + 2
phorous PO4 , calcium (Ca ), and sodium (Na ) has been reported in generated spent wash ( Jain and Srivastava, 2012; Chandra and Kumar, 2017a). Moreover, it also contains large concentrations of heavy metals viz. iron (Fe), zinc (Zn), nickel (Ni), manganese (Mn), lead (Pb), mercury (Hg), copper (Cu), and chromium (Cr) ( Jack et al., 2014; Mahar et al., 2013). Recently, spent wash was reported to contain high amounts of butanedioic acid bis(TMS)ester; 2-hydroxysocaproic acid; benzenepropanoic acid, α-[(TMS)oxy], TMS ester; vanillylpropionic acid, bis(TMS), and other recalcitrant organic pollutants such as 2-furancarboxylic acid, 5-[[(TMS)oxy] methyl], TMS ester; benzoic acid 3-methoxy-4-[(TMS)oxy], TMS ester; and tricarballylic acid 3TMS (Chandra and Kumar, 2017a). According to the USEPA (2012), most of these identified organic compounds are toxic, carcinogenic, mutagenic, and endocrine disruptors in nature while some organic compounds in spent wash are recalcitrant to biodegradation (Chowdhary et al., 2018). The pollutants present in generated spent wash are reacting to each other and make the wastewater more toxic and complex (Arimi et al., 2014; Kumar and Sharma, 2019). Depending on the sugarcane origin and opted industrial process for alcohol production, the intrinsic composition of spent wash can vary significantly. The acidic pH (from 3.8 to 5) of spent wash is associated mainly with the presence of organic acids produced by the yeasts during the alcoholic fermentation process. The acidity of the spent wash causes the dissolution of metals in water bodies. Moreover, the dark brown color of spent wash is imparted by complex compounds such as melanoidin, caramel, hexose alkaline degradation products (HADP), furfurans (from acid hydrolysis), lignin, and polyphenols as well as plant pigments such as carotenoids, chlorophyll, anthocyanins, and tannins, which make spent wash more complex (Hatano et al., 2008; Dai and Mumper, 2010; Arimi et al., 2014; Zhang et al., 2017). These colorants are concentrated in molasses after the crystallization of sugar and are further transferred to the spent slurry during molasses fermentation. Among the color-causing pigments, melanoidin is the major dark brown color pigment of spent wash, and it is produced through nonenzymatic browning reactions such as the Maillard reaction, alkaline degradation reactions, or sugar degradation occurred at medium temperature ( >50 °C) and in a basic pH medium (Hodge, 1953; Coca et al., 2004; Kumar and Chandra, 2018a). Melanoidin polymers are generally regarded to be heterogeneous and acidic with a high molecular weight (5000 and 40,000 Da) along with similar chemical properties to humic

Chapter 1 Advanced oxidation processes for complex wastewater treatment

substances (i.e., humic and fulvic acids) (Liang et al., 2009; Liu et al., 2013; Kumar and Chandra, 2020). They are composed of highly dispersed colloids, which are negatively charged due to the dissociation of carboxylic acids and phenolic groups. Melanoidins are difficult to characterize due to their varying sizes and the types of reducing sugars and amino acids involved in their formation (Chandra et al., 2008; Arimi et al., 2014; Hatano and Yamatsu, 2018). It has been reported that melanoidins have net negative charges. Therefore, different heavy metals such as Cu, Cr, Cd, Fe, Zn, Ni, and Pb strongly bind with melanoidins to form organometallic complexes in distillery spent wash (Migo et al., 1997; Hatano et al., 2013, 2016; Chandra et al., 2018a,b,c). Over the last decades and due to its high inorganic loads, spent wash has been widely used as a liquid fertilizer for sustainable agriculture (Kumar and Chopra, 2012; Jain and Srivastava, 2012; Kumari et al., 2015). However, regulations have made spreading spent wash more difficult because of its low pH and high organic and inorganic content, which may be responsible for groundwater contamination and soil compaction (Ansari, 2014). Some studies have indicated that spent wash negatively affects the physical properties of soil, such as hydraulic conductivity and redox potential (Alves et al., 2015). Thus, the safe disposal of spent wash is becoming a serious problem throughout the world (Chowdhary et al., 2018). Different conventional processes such as anaerobic digestion (biomethanation), anaerobic lagoons, and activated sludge are available to treat spent wash (Kumar and Sharma, 2019; Kumar et al., 2020). Among them, biomethanation is a popular, cost-effective, first-step conventional biological treatment of spent wash that produces methane to meet part of the power requirement in distilleries (Khairnar et al., 2013; Sankaran et al., 2017). This methane gas is mainly utilized for running steam boilers to generate electricity. On average, 1 m3 of spent wash produces 38–40 m3 of biogas (Sankaran et al., 2014). Moreover, in most instances during biomethanation, the melanoidin compounds repolymerize, thereby intensifying the color of spent wash and making the decolorization of wastewater even more difficult (Zhang et al., 2017). The polymerization may also extend to different levels, and it occurs in complex ways. After biomethanation, the distillery wastewater still retains a high COD (40,000– 52,000 mg L
1), BOD (8000–12,000 mg L
1), and substantial color (Saner et al., 2014; Naveen and Premalatha, 2016). Most of the distilleries employ a direct two-stage aerobic process for further treatment of the biomethanated spent wash. Often, it becomes recalcitrant [biodegradability index (BI) 95% and color removal >80%. This study has shown that through ozonolysis, the complex aromatic compounds are broken down into simple aliphatic compounds, resulting in an increased oxidation state and improved biodegradability. Apart from color abatement, the ozonolysis process also achieved 88% sludge solubilization. The use of ozonation as a posttreatment has an advantage of low sludge formation but its application is limited by the high installation and operation costs. The problem becomes even more complicated if the process involves high daily volumes of effluents. Another limitation of the ozonation process is the low COD removal, especially where the influent has reasonably high COD (Coca et al., 2005). However, it is still interesting to study the chemical pretreatment (partial oxidation) to increase the microbial biodegradability (posttreatment) of molasses distillery wastewater colorants to find the optimal balance of cost and performance.

2.3

Photocatalytic treatment

Photocatalytic degradation is an attractive, efficient, and costeffective treatment technology to enhance the biodegradability of hazardous and nonbiodegradable contaminants, such as persistent organic pollutants (Mabuza et al., 2017; Takle et al., 2018). This process allows the transformation of chemical pollutants into less toxic substances and/or with structural features that are more readily biodegradable. Photocatalytic treatment process is based on the combination of oxidizing agents with an appropriated catalyst and/or light. These processes may be of particular interest for treating effluent containing highly toxic compounds, and for which the biological processes might not be pertinent (Catalkaya and Sengul, 2006). The basis of photocatalysis is the photoexcitation of a semiconductor that is solid as a result of the absorption of electromagnetic radiation, often but not exclusively in the near-UV spectrum. Under near-UV irradiation, a suitable semiconductor material may be excited by photons possessing energies of sufficient magnitude to produce conduction band electrons (e
) and valence band holes (h+). The e
and the h+ migrate from their respective bands to the surface of the semiconductor. They react with a suitable redox species in the environment, which could lead to H2O splitting (H2 and O2 generation) and the formation of OH and O2
radicals. These radicals are able to degrade a wide range of recalcitrant organic compounds and detoxify HM ions (e.g., Cr6+ and As3+). The results achieved in recent years show that technologies based on photochemical and photocatalytic approaches seem to be very

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Chapter 1 Advanced oxidation processes for complex wastewater treatment

promising as viable alternatives for wastewater treatments (Matthews, 1991; Poulopoulos et al., 2019; Takle et al., 2018). The photocatalytic degradation process using solar light as an irradiation source showed potential application for the removal of the color and/or COD from distillery effluent (Vineetha et al., 2013). The maximum color removal of the distillery wastewater achieved was 79% at an H2O2 concentration of 0.3 M, a pH of 6, an effluent COD concentration of 500 ppm, and a catalyst dosage of 0.1 g/L. The TiO2/H2O2 system seems to be more efficient in comparison to the synergetic action that appears when using H2O2 and TiO2 (Vineetha et al., 2013). Recently, a vanadium-doped TiO2 (V-TiO2) photocatalyst was studied via the degradation of spent wash and Jakofix red dye (HE 8BN) under natural sunlight as well as a standard artificial solar energy source (Xe lamp) (Takle et al., 2018). The highest activity was observed for a 1% V-TiO2 photocatalyst for the degradation of spent wash and Jakofix red dye under natural sunlight. Further, the degradation of colored compounds in the spent wash was monitored by gel permeation chromatography, which showed the degradation of high molecular weight compounds into low molecular weight fractions. The catalyst decomposed 90% of the Jakofix red dye (HE 8BN) in 3.5 h and 65% of the spent wash in 5 h under irradiation with natural sunlight. A cycling stability test showed the high stability and reusability of the photocatalyst for degradation reactions, with a recovery of around 94%–96%. David et al. (2015a) reported the decolorization and degradation of spent wash using a nano-Al2O3/kaolin photocatalyst. In this study, the degradation of organic pollutants in the form of color was performed using a nanophotocatalyst prepared using an aluminum oxide (Al2O3) nanoparticle and kaolin clay. The process parameters such as dosage, pH, temperature, and agitation were optimized using a Taguchi orthogonal array design to attain the maximum decolorization efficiency. Optimization of the process parameters resulted in a maximum of 80% spent wash decolorization. Using an artificial neural network (ANN), a two-layered feed-forward back propagation model offered the best performance and predictive model for spent wash decolorization. The experimental data were found to be in excellent agreement with the predicted results from the ANN model. The photocatalytic degradation of molasses wastewater has been reported to be effective for the complete removal of the recalcitrant melanoidins contained in molasses wastewater, leading to the color disappearance. This process, however, is ineffective for the removal of the high organic content of molasses wastewater. Therefore, it has been recommended for application as a plausible posttreatment to anaerobic digestion (Otieno et al., 2016).

Chapter 1 Advanced oxidation processes for complex wastewater treatment

2.4

UV-based oxidation treatment

In UV direct photolysis, the degradation process through the absorption of incident radiation from the UV light is the main removal mechanism. Therefore, the application usually focuses on the contaminants that strongly absorb UV radiation. Most UV light absorbers contain double bonds or conjugated double bonds, which includes carbon, nitrogen or oxygen atoms and is characterized by delocalized π-electrons (Deng and Zhao, 2015). UV photolysis can selectively reduce some organic compounds, but it alone is not efficient enough for pollutant removal. However, many researchers have found that UV photolysis can enhance the oxidation potential of some oxidation processes. When UV photolysis is combined with ozonation (UV/O3 oxidation), more OH is produced via the photolysis of H2O2 as a reaction intermediate, and organic pollutants are decomposed more completely. The generation of OH is fundamental to the process, as the OH is largely responsible for the success of this process. This process combines both H2O2 and UV light in a synergistic effect to degrade the contaminants and pathogenic microorganisms in the contaminated water. In wastewater treatment processes, UV/O3 oxidation can be used as a pretreatment for biological processes in which complex pollutants are decomposed into more biodegradable substances. Therefore, more organic compounds can be removed in the subsequent biological treatment process. The degradation of dissolved organic nitrogen (DON)-associated color from wastewater containing melanoidins by using a UV/H2O2 AOP was investigated by Jason Dwyer (2008). The oxidation process was much more capable of removing 99% color, 50% dissolved organic carbon (DOC), and 25% DON at the optimal applied dose of UV/H2O2 for the system (3300 mg L
1). This indicated that color and DON removal were decoupled problems for the purpose of treating melanoidin by an AOP; therefore, color removal cannot be used as an indication of DON removal. Color was caused by organic molecules with molecular weights greater than 10 kDa. Oxidation caused a partial reduction of the DON (41%–15% of the total dissolved nitrogen) and DOC (29%– 14% of the DOC) associated with the large molecular weight fraction (>10 kDa) as well as almost complete color removal. The degraded DON was mostly accounted for by the formation of ammonia (31% of the nitrogen removed from the large fraction) and small molecular weight compounds (66% of the nitrogen removed from the large fraction). The degraded DOC appeared to be mostly mineralized (to CO2) with only 20% of the degraded compounds appearing as small molecular weight DOC (Dwyer

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Chapter 1 Advanced oxidation processes for complex wastewater treatment

et al., 2008). UV photodegradation, on the other hand, can degrade the color-causing melanoidins, leading to complete color removal after a short duration of irradiation. However, the UV process is energy-intensive and incapable of effective TOC reduction. For effective molasses wastewater treatment, these two processes can be integrated such that UV photodegradation is applied as a posttreatment to anaerobic digestion. In this system, the anaerobic digestion first removes the high organic content, followed by color removal by the UV process. Desk-scale anaerobic digestion and photodegradation processes were carried out in batch reactors. A hybrid photocatalyst consisting of titanium dioxide (TiO2) and zinc oxide (ZnO) was used for photocatalytic degradation. Biodegradation of wastewater during anaerobic digestion at thermophilic conditions in the bioreactor achieved high TOC and COD reductions of 80% and 90%, respectively, but with increased color intensity. Contrastingly, UV photodegradation achieved a high color reduction of 92% with an insignificant 6% TOC reduction after 30 min of irradiation. During photodegradation, the mineralization of the recalcitrant organic compounds led to color disappearance. The author suggested that the UV process was only suitable for color reduction. Therefore, there is a possible synergy when the two processes are integrated with anaerobic digestion preceding UV, where anaerobic digestion removes the high COD/TOC while UV removes the recalcitrant color at a reduced cost (Mabuza et al., 2017). The main drawback of using the above-mentioned oxidation processes is their high operational cost.

2.5

Sonication/sonolysis

Ultrasound is increasingly being seen as having potential in the treatment of water, wastewater, and sewage sludge. Sonochemical oxidation employs the use of ultrasound, resulting in the cavitation phenomenon. That phenomenon is defined as the formation, growth, and subsequent collapse of microbubbles or cavities occurring in extremely small intervals of time (microseconds) and releasing large magnitudes of energy at millions of such locations in the reactor. Sonochemical reactions are induced upon high-intensity acoustic irradiation of liquids at frequencies that produce cavitation (typically in the range of 20–1000 kHz). It is generally believed that there are three potential sites for chemical reactions in ultrasonically irradiated liquids: the cavitation bubble itself, the interfacial region between the bubble and the surrounding liquid, and the solution bulk. Ultrasound irradiation effectively destroys the contaminants in water because of localized

Chapter 1 Advanced oxidation processes for complex wastewater treatment

high concentrations of OH and H2O2 in the solution as well as high localized temperatures and pressures. Sangave and Pandit (2004) sonicated distillery wastewater as a pretreatment step to convert complex molecules into a more utilizable form by cavitation. Samples exposed to 2 h ultrasound pretreatment displayed 44% COD removal after 72 h of conventional aerobic oxidation compared to 25% COD reduction shown by untreated samples. These results are contrary to those of Mandal et al. (2003), who concluded that ultrasonic treatment was ineffective for distillery spent wash treatment. The combined processes of sonolysis, enzymatic hydrolysis, and aerobic biological oxidation were found to increase the biodegradation efficiency of distillery wastewater (Sangave et al., 2007a,b). However, the time scale and the dissipated power necessary to obtain complete mineralization of the pollutants in the case of sonolysis are not economically acceptable. Hence, ultrasound is found more effective when used in combination with other conventional treatment processes than as a stand-alone process. Besides, the application of sonolysis in the treatment of complex effluents is limited by the high installation/process costs as well as low COD removal.

2.6

Electrochemical oxidation

In recent years, new AOPs based on electrochemical technology, that is, electrochemical advanced oxidation processes (EAOPs), have been developed. Electrochemical oxidation is widely used to treat contaminates present in distillery effluent (Canizares et al., 2009; Asaithambi et al., 2012a,b; Sahu and Chaudhari, 2015). These processes use electrons as the main reagent, but also require the presence of supporting electrolytes. In the electrochemical process, the pollutants are destroyed by either direct oxidation or indirect oxidation (Thakur et al., 2009; Martı´nez-Huitle and Panizza, 2018). In a direct anodic oxidation process, the pollutants are first adsorbed on the anode surface and then destroyed by the anodic electron transfer reaction. In an indirect oxidation process, strong oxidants such as hypochlorite/chlorine, O3, or H2O2 are electrochemically generated by either the anodic or cathodic process. The organic pollutants are then destroyed in the bulk solution by the oxidation reaction of the generated oxidant. The oxidative performances of three different anode materials-planar graphite (Gr), lead dioxide-coated titanium (PbO2-Ti), and ruthenium oxide-coated titanium (RuO2Ti)-on the electrooxidation of industry-treated distillery effluent were studied by Manisankar et al. (2003). Maximums of 92% of COD reduction, 98.1% of BOD reduction, and 99.5% of absorbance

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Chapter 1 Advanced oxidation processes for complex wastewater treatment

reduction were obtained in the set up in which RuO2-Ti as the anode and stainless steel as the cathode were used. Electrochemical oxidation provides several advantages for prevention and remediation such as high energy efficiency, amenability to automation, easy handling because of the simple equipment required, and safety. Despite the substantial effectiveness of this process in removing organics, it has some drawbacks such as the production of large quantities of sludge, the consumption of chemicals, and the necessity of acidic conditions.

2.7

Wet air oxidation

Wet oxidation, also called wet air oxidation (WAO), refers to the process of oxidizing suspended or dissolved organic and inorganic material in the liquid phase with dissolved oxygen at elevated temperatures (125–320°C) and pressures (0.5–20 MPa) using a gaseous source of oxygen (O2)/air (Zou et al., 2007; Bhargava et al., 2006). At elevated temperatures and pressures, the solubility of O2 in the aqueous solution increases and provides a strong driving force for oxidation of pollutants. Generally, elevated pressures are maintained to keep aqueous solutions in a liquid state (Mishra et al., 1995). WAO is a well-established pretreatment technique for hazardous complex industrial wastewater that is too dilute to incinerate and too concentrated for biological treatment. A few studies also explored the possibilities of complex wastewater treatment by the WAO, which can be followed as a pretreatment step to enhance the biodegradability and facilitate biogas generation in subsequent biological treatment processes (Padoley et al., 2012a,b; Goto et al., 1998). A lab-scale WAO reactor with biomethanated distillery wastewater (COD: 40,000 mg L
1; initial BI of 0.17) was used to demonstrate the proof of concept (Padoley et al., 2012a,b). The studies were conducted using a designed set of experiments; reaction temperatures (150–200°C), air pressure (6–12 bar), and reaction time (15– 120 min) were the main process variables of concern for WAO process optimization. They found that optimum WAO pretreatment conditions were 175°C, 0.6 MPa air pressure, and 0.5 h (final BI of 0.4, residual COD: 60%). This pretreatment process is beneficial because it facilitated the degradation of melanoidin, leading to BI enhancement and reduced toxicity of wastewater. WAO-induced enhanced biodegradability of distillery effluent was reported by Malik et al. (2014). Initially, the distillery effluent was pretreated by WAO at different process conditions (pressure, temperature, and time) to facilitate enhancement in the BI. The biodegradability of WAO pretreated effluent was evaluated by subjecting it to aerobic

Chapter 1 Advanced oxidation processes for complex wastewater treatment

biodegradation and anaerobic followed by aerobic biodegradation. The biodegradation of pretreated effluent (BI:0.4–0.8) through aerobic process enhanced COD reduction of up to 67.7%, whereas the untreated effluent (BI:0.17) indicated poor COD reduction of only 22.5%. Anaerobic followed by aerobic biodegradation of pretreated effluent has shown up to 87.9% COD reduction while the untreated effluent has shown only 43.1% COD reduction. This study demonstrated that the WAO pretreatment enhanced biooxidation/ biodegradation of complex distillery effluent. Daga et al. (1986) reported WAO treatment of distillery spent wash targeting COD reduction in the temperature range of 150–230°C and at an O2 partial pressure range of 0–2.5 MPa. Shah et al. (1989) reported the use of WAO for the treatment of alcohol distillery wastewater followed by aerobic treatment. Recently, Tembhekar et al. (2015) investigated the kinetics of COD destruction and BI improvement during the WAO. In alcohol distilleries, the treatment of biomethanated spent wash is crucial. After biogas recovery, spent wash has a high COD and resists further oxidation. Single-stage anaerobic digestion processes are insufficient to achieve higher removal of organic matter; hence, they can be combined with single- or multistage aerobic or chemical oxidation treatments. The coupling of treatment processes can achieve the improved efficiency of distillery wastewater treatment with improved COD and color removal. A hybrid pretreatment technique was explored for the selective improvement in biodegradability of biomethanated spent wash (Bhoite and Vaidya, 2018a,b). In this work, the biodegradability of biomethanated distillery spent wash (BOD5 8100 and COD 40,000mg/L) was improved by oxidation using an FeSO4 catalyst at 175°C for 1 h. After oxidation, adsorption over activated carbon (loading 5 g/100 mL wastewater) was resulted 73% reduction in COD and a substantial increase in the BOD5/COD ratio (from 0.2 to 0.45). This BI further rose to 0.52 when anaerobic digestion was performed, using 1% acclimatized biomass. After a final polishing aerobic treatment step, the BI and COD were reduced up to 0.58 and 91%, respectively. Clearly, this work has provided a useful solution to the effective pretreatment and valorization of biomethanated distillery wastewaters. Additionally, this hybrid pretreatment technique provides the opportunity to generate more methane from the distillery spent wash (Bhoite and Vaidya, 2018b). In another study, the iron-catalyzed WAO of biomethanated spent wash exhibited high BOD5 (8100 mg/L) and COD (40,000 mg/L) for enhanced biogas recovery, as reported by Bhoite and Vaidya (2018a). In this study, for the catalytic oxidation process, two ironbased heterogeneous catalysts were employed: Fe2O3 and Fe/C. In a batch slurry reactor, the catalyst performance was investigated at the denoted temperature (175°C), the oxygen partial pressure

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(0.69 MPa), and the catalyst loading (33 mg/L) for 1 h. After WAO for 1 h, the BI value increased from 0.2 to 0.47 and the COD reduction was 71%. It was found that the activity of the catalysts reduced in the order: Fe/AC-N > Fe/AC-T > Fe2O3. The results were more encouraging, including the COD conversion (87%), the color reduction (88%), and the BI value (0.71) when carbon adsorption (5% w/v) followed WAO over Fe/AC-N. Clearly, our novel hybrid process for pretreatment, wet oxidation-carbon adsorption, showed potential. Postbiomethanation, around 1.2 N m3 biogas (CH4 72%) was formed per cubic meter of the wastewater; without pretreatment by catalytic WAO and carbon adsorption, the yield of biogas (CH4 11%) was just 1 Nm3 for every cubic meter of wastewater. Finally, 97% COD was removed and the BI value was 0.84 during polishing by aerobic treatment. The author recommended a useful option for the effective treatment of spent wash and additional energy recovery in the form of biogas.

3 Combined oxidation treatment approaches Distillery wastewater often exhibits different properties depending upon its origin. Therefore, a “universal” treatment method for its degradation and detoxification is not available (Kumar and Chandra, 2020; Kumar and Sharma, 2019). However, the combined AOPs are the appropriate technique, as they provide economically and technically flexible options and have been examined by different groups of researchers worldwide for the effective treatment of distillery wastewater. The treatment of distillery wastewater for the percentage of color and COD removal was investigated through various AOPs such as Fenton and ozone with different combinations of UV and H2O2 systems. The ozonephoto-Fenton system resulted in 100% color and COD removal compared with other process (Asaithambi et al., 2015). Besides, several combinations of chemical/biological treatment processes have been reported for the degradation of contaminants present in distillery wastewater. Kumar et al. (2006) carried out an experiment to explore strategies for the mineralization of refractory compounds in distillery spent wash by anaerobic biodegradation/ozonation/aerobic biodegradation. The treatment of distillery spent wash by anaerobic-aerobic biodegradation was shown in an overall COD removal of 70.8%. Ozonation of the anaerobically digested distillery spent wash was carried out as-is (phase

Chapter 1 Advanced oxidation processes for complex wastewater treatment

I experiments) and after pH reduction and the removal of inorganic carbon (phase II experiments). The ozonation step resulted in an increase in overall COD removal (95%) obtained when an ozone dose of approximately 5.3 mg ozone absorbed/mg initial total organic carbon was used. The author concluded that the combination of chemical and biological processes led to the greater destruction of organic contaminants present in the effluent (Kumar et al., 2006). Similarly, laboratory-scale experiments were performed to degrade highly concentrated organic matter in the form of color in distillery spent wash through batch oxidative methods such as electrocoagulation (EC), electro-Fenton (EF), and the Fenton process (David et al., 2015a,b). For EC, 79% color removal was achieved using iron electrodes whereas in EF, 44% spent wash decolorization was observed using carbon (graphite) electrodes with optimum conditions. By the Fenton process, 66% decolorization was attained at optimized conditions. Likewise, a combination of flocculation-ozonation-Fenton established an effective alternative for treating distillery wastewater (Martins et al., 2011); however, these processes are very expensive to operate.

4

Advantages and limitations of advanced oxidation treatment processes

AOPs offer several advantages: It is widely known that conventional wastewater treatment systems have serious shortcomings that can be addressed by AOPs. This method not much required the filtration of the sample it treats the whole distillery wastewater.  A major advantage of AOPs is related to the transformation of organic compounds to simpler stable inorganic compounds such as water, carbon dioxide, and salts with little/no sludge production, which erases the need for another treatment stage.  AOPs can be useful in pretreatment to transform recalcitrant pollutants so they can be biologically treated or as a posttreatment before their discharge into the environment.  A rapid reaction rate of AOPs for distillery wastewater treatment is the use of HO. Despite its advantages, there are several limitation/disadvantages of AOPs:  AOPs, especially Fenton oxidation, are usually accomplished in acid conditions, resulting in large quantities of acid and base consumption. 

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AOP system utilizing H2O2 may be harmful to humans. This process is also a dosage-dependent process, so the appropriate amounts of HO molecules are formed to achieve the desired level of treatment. The major limitation of AOPs is a relatively high-cost process due to the use of costly chemicals and increased energy consumption as well as the formation of unknown recalcitrant by-products that, in some cases, may well be a lot of nephrotoxic than the parent compounds stay unsolved. All AOPs are used to treat the scavenging of group radicals by nontarget substances while they are not suitable for certain categories of toxic compounds that resist attack by HO.

5 Conclusions and future outlook The treatment and safe disposal of distillery wastewater has been a challenge for modern technologies and extensive research done by the scientific community due to its potential danger to the aquatic and soil environments, especially with respect to human and animal health. However, several AOPs and their different combinations for treating alcohol distillery wastewater have been explored for the complete removal of suspended solids and the biodegradation of undecomposed polyphenols and other refractory compounds present in distillery wastewater. They have been implemented extensively in both the bench and pilot scales. However, these processes are expensive and energy-intensive in terms of operation and maintenance for long-term applications and commercial implementation. Therefore, further efforts are required for piloting these technologies to truly evaluate the advantages and drawbacks of these systems in real-life applications. The reliability of such processes for the assured treatment of wastewater in the long run to meet stipulated discharge norms also needs to be assessed. AOP systems are not widely used in industry. This is not because of their inability to treat wastewater, but rather because of a general lack of awareness and understanding about the current advanced oxidation technologies and their applications to different cases. Overall, the future for the applicability of AOPs processes is bright because their combination with conventional wastewater treatment systems is conceptually feasible. Future developments will rely on the close collaboration of analytical chemists, engineers, and electrochemists to ensure the effective application and exploitation of these oxidative technologies.

Chapter 1 Advanced oxidation processes for complex wastewater treatment

References Acharya, B.K., Pathak, H., Mohana, S., Shouche, Y., Singh, V., Madamwar, D., 2011. Kinetic modeling and microbial community assessment of anaerobic biphasic fixed film bioreactor treating distillery spent wash. Water Res. 45, 4248–4259. Ahmed, B., Limem, E., Abdel-Wahab, A., Nasr, B., 2011. Photo-Fenton treatment of actual agro-industrial wastewaters. Ind. Eng. Chem. Res. 50, 6673–6680. Alves, P.R.L., Natal-da-Luz, T., Sousa, J., Cardoso, E.J.B.N., 2015. Ecotoxicological characterization of sugarcane vinasses when applied to tropical soils. Sci. Total Environ. 526, 222–232. Ansari, F., 2014. Environmental impact of distillery effluent on vertical soil horizon due to leaching effect: an experimental approach. Int. J. Chem. Environ. Eng. 5, 223. Arimi, M.M., 2017. Modified natural zeolite as heterogeneous Fenton catalyst in treatment of recalcitrants in industrial effluent. Prog. Nat. Sci. Mater. Int. 27 (2), 275–282. Arimi, M.M., Zhang, Y., Goetz, G., Kiriamiti, K., Geissen, S.U., 2014. Antimicrobial colorants in molasses distillery wastewater and their removal technologies. Int. Biodeterior. Biodegrad. 87, 34–43. Asaithambi, P., Garlanka, L., Anantharaman, N., Matheswaran, M., 2012a. Influence of experimental parameters in the treatment of distillery effluent by electrochemical oxidation. Sep. Sci. Technol. 47 (3), 470–481. Asaithambi, P., Susree, M., Saravanathamizhan, R., Matheswaran, M., 2012b. Ozone assisted electrocoagulation for the treatment of distillery effluent. Desalination 297, 1–7. Asaithambi, P., Saravanathamizhan, R., Matheswaran, M., 2015. Comparison of treatment and energy efficiency of advanced oxidation processes for the distillery wastewater. Int. J. Environ. Sci. Technol. 12, 2213–2220. Ayyasamy, P.M., Yasodha, R., Rajakumar, S., Lakshmanaperumalsamy, P., Rahman, P.K.S.M., Lee, S., 2008. Impact of sugar factory effluent on the growth and biochemical characteristics of terrestrial and aquatic plants. Bull. Environ. Contam. Toxicol. 81, 449–454. Battimelli, A., Loisel, D., Garcia-Bernet, D., Carrere, H., Delgenes, J.P., 2010. Combined ozone pretreatment and biological processes for removal of colored and biorefractory compounds in wastewater from molasses fermentation industries. J. Chem. Technol. Biotechnol. 85, 968–975. ´ lvarez, P.M., 1999. Wine distillery wastewater degBeltra´n, F.J., Garcı´a-Araya, J.F., A radation. 1. Oxidative treatment using ozone and its effect on the wastewater biodegradability. J. Agric. Food Chem. 47 (9), 3911–3918. ´ lvarez, P.M., 2000. Continuous flow integrated Beltra´n, F.J., Garcı´a-Araya, J.F., A chemical (ozone)-activated sludge system treating combined agroindustrialdomestic wastewater. Environ. Progr. 19 (1), 28–35. Beltran, F.J., Alvarez, P.M., Rodriguez, E.M., Garcia-Araya, J.F., Rivas, J., 2001. Treatment of high strength distillery wastewater (cherry stillage) by integrated aerobic biological oxidation and ozonation. Biotechnol. Prog. 17, 462–467. Benitez, F., Beltran-Heredia, J., Real, F., Acero, J.L., 1999. Enhancement of the ozonation of wine distillery wastewaters by an aerobic pretreatment. Bioprocess Eng. 21 (5), 459–464. Bharagava, R.N., Chandra, R., 2010. Effect of bacteria treated and untreated postmethanated distillery effluent (PMDE) on seed germination, seedling growth and amylase activity in Phaseolus mungo L. J. Hazard. Mater. 180, 730–734. € ger, K., Akolekar, D.B., Grocott, S.C., 2006. Bhargava, S.K., Tardio, J., Prasad, J., Fo Wet oxidation and catalytic wet oxidation. Ind. Eng. Chem. Res. 45 (4), 1221–1258.

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Bhoite, G.M., Vaidya, P.D., 2018a. Iron-catalyzed wet air oxidation of biomethanated distillery wastewater for enhanced biogas recovery. J. Environ. Manag. 226, 241–248. Bhoite, G.M., Vaidya, P.D., 2018b. Improved biogas generation from biomethanated distillery wastewater by pretreatment with catalytic wet air oxidation. Ind. Eng. Chem. Res. 57, 2698–2704. Bhoite, G.M., Vaidya, P.D., 2019. Fenton oxidation and adsorption pretreatment for superior biogas recovery from biomethanated spent wash. Chem. Eng. Commun. https://doi.org/10.1080/00986445.2019.1650031. Blonskaja, V., Zub, S., 2009. Possible ways for post treatment of biologically treated wastewater from yeast factory. J. Environ. Eng. Landsc. Manag. 17, 189–197. Caderby, E., Baumberger, S., Hoareau, W., Fargues, C., Decloux, M., Maillard, M., 2013. Sugar cane stillage: a potential source of natural antioxidants. J. Agric. Food Chem. 61, 11494–11501. Canizares, P., Herna´ndez-Ortega, M., Rodrigo, M.A., Barrera-Diaz, C.E., RoaMorales, G., Sa´ez, C., 2009. A comparison between conductive-diamond electrochemical oxidation and other advanced oxidation processes for the treatment of synthetic melanoidins. J. Hazard. Mater. 164, 120–125. Catalkaya, E.C., Sengul, F., 2006. Application of Box-Wilson experimental design method for the photodegradation of bakery’s yeast industry with UV/H2O2 and UV/H2O2/Fe(II) process. J. Hazard. Mater. 128, 201–207. Chain, E.S.K., Dewalle, F.B., 1977. Treatment of high strength acidic wastewater with completely mixed anaerobic filter. Water Res. 11, 295. Chandra, R., Kumar, V., 2015. Biotransformation and biodegradation of organophosphates and organohalides. In: Chandra, R. (Ed.), Environmental Waste Management. CRC Press, Boca Raton, pp. 475–524. Chandra, R., Kumar, V., 2017a. Detection of Bacillus and Stenotrophomonas species growing in an organic acid and endocrine-disrupting chemicals rich environment of distillery spent wash and its phytotoxicity. Environ. Monit. Assess. 189, 26. Chandra, R., Kumar, V., 2017c. Phytoextraction of heavy metals by potential native plants and their microscopic observation of root growing on stabilized distillery sludge as a prospective tool for in-situ phytoremediation of industrial waste. Environ. Sci. Pollut. Res. 24, 2605–2619. Chandra, R., Kumar, V., 2018. Phytoremediation: a green sustainable technology for industrial waste management. In: Chandra, R., Dubey, N.K., Kumar, V. (Eds.), Phytoremediation of Environmental Pollutants. CRC Press, Boca Raton. Chandra, R., Bharagava, R.N., Rai, V., 2008. Melanoidins as major colourant in sugarcanemolasses based distillery effluent and its degradation. Bioresour. Technol. 99, 4648–4660. Chandra, R., Yadav, S., Kumar, V., 2015. Microbial degradation of lignocellulosic waste and its metabolic products. In: Chandra, R. (Ed.), Environmental Waste Management. CRC Press, Boca Raton, pp. 249–298. Chandra, R., Kumar, V., Tripathi, S., 2018a. Evaluation of molasses-melanoidin decolourisation by potential bacterial consortium discharged in distillery effluent. 3 Biotech 8, 187. Chandra, R., Kumar, V., Tripathi, S., Sharma, P., 2018b. Heavy metal phytoextraction potential of heavy metal phytoextraction potential of native weeds and grasses from endocrine-disrupting chemicals rich complex distillery sludge and their histological observations during in situ phytoremediation. Ecol. Eng. 111, 143–156. Chandra, R., Kumar, V., Tripathi, S., Sharma, P., 2018c. Phytoremediation of industrial pollutants and life cycle assessment. In: Chandra, R., Dubey, N.K., Kumar, V. (Eds.), Phytoremediation of Environmental Pollutants. CRC Press, Boca Raton.

Chapter 1 Advanced oxidation processes for complex wastewater treatment

Chauhan, J., Rai, J.P.N., 2010. Monitoring of impact of ferti-irrigation by postmethanated distillery effluent on groundwater quality. Clean: Soil, Air, Water 38 (7), 630–638. Chidankumar, C.S., Chandraju, S., 2008. Impact of irrigation of distillery spentwash on the nutrients of pulses in untreated and treated soil. Sugar Tech 10 (4), 314–318. Chowdhary, P., Raj, A., Bharagava, R.N., 2018. Environmental pollution and health hazards from distillery wastewater and treatment approaches to combat the environmental threats: a review. Chemosphere 194, 229–246. Coca, M., Garcıa, M.T., Gonzalez, G., Pena, M., Garcıa, J.A., 2004. Study of coloured components formed in sugar beet processing. Food Chem. 86, 421–433. Coca, M., Pena, M., Gonzalez, G., 2005. Variables affecting efficiency of molasses fermentation wastewater ozonation. Chemosphere 60, 1408–1415. Daga, N.S., Prasad, C.V.S., Joshi, J.B., 1986. Kinetics of hydrolysis and wet air oxidation of alcohol distillery waste. Ind. Eng. Chem. Res. 28, 22–31. Dai, J., Mumper, R.J., 2010. Plant phenolics: extraction, analysis and their antioxidant and anticancer properties. Molecules 15 (10), 7313–7352. David, C., Arivazhagan, M., Ibrahim, M., 2015a. Spent wash decolourization using nano-Al2O3/kaolin photocatalyst: taguchi and ANN approach. J. Saudi Chem. Soc. 19, 537–548. David, C., Arivazhagann, M., Tuvakara, F., 2015b. Decolorization of distillery spent wash effluent by electrooxidation (EC and EF) and Fenton processes: a comparative study. Ecotoxicol. Environ. Saf. 121, 142–148. De Vrieze, J., Hennebel, T., Van den Brande, J., Bilad, R.M., Bruton, T.A., Vankelecom, I.F.J., Verstraete, W., Boon, N., 2014. Anaerobic digestion of molasses by means of a vibrating and non-vibrating submerged anaerobic membrane bioreactor. Biomass Bioenergy 68, 95–105. Deng, Y., Zhao, R., 2015. Advanced oxidation processes (AOPs) in wastewater treatment. Cur. Pollut. Rep. 1 (3), 167–176. Derakhshan, M., Fazeli, M., 2018. Improved biodegradability of hardly decomposable wastewaters from petrochemical industry through photo-Fenton method and determination of optimum operational conditions by response surface methodology. J. Biol. Eng. 12, 10. Dwyer, J., Kavanagh, L., Lant, P., 2008. The degradation of dissolved organic nitrogen associated with melanoidin using a UV/H2O2 AOP. Chemosphere 71, 1745–1753. Ebrahiem, E.E., Al-Maghrabi, M.N., Mobarki, A.R., 2017. Removal of organic pollutants from industrial wastewater by applying photo-Fenton oxidation technology. Arab. J. Chem. 10 (2), S1674–S1679. Ghosh, M., Verma, S.C., Mengoni, A., Tripathi, A.K., 2004. Enrichment and identification of bacteria capable of reducing chemical oxygen demand of anaerobically treated molasses spent wash. J. Appl. Microbiol. 96, 1278–1286. Goto, M., Nada, T., Ogata, A., Kodama, A., Hirose, T., 1998. Supercritical water oxidation for the destruction of municipal excess sludge and alcohol distillery wastewater of molasses. J. Supercrit. Fluids 13 (1), 277–282. Gulyas, H., von Bismarck, R., Hemmerling, L., 1995. Treatment of industrial wastewaters with ozone/hydrogen peroxide. Water Sci. Technol. 2 (7), 127–134. Hadavifar, M., Zinatizadeh, A.A., Younesi, H., Galehdar, M., 2009. Fenton and photo-Fenton treatment of distillery effluent and optimization of treatment conditions with response surface methodology. Asian-Pacific J. Chem. Eng.. https://doi.org/10.1002/apj.313. Hatano, K., Yamatsu, T., 2018. Molasses melanoidin-like products enhance phytoextraction of lead through three Brassica species. Int. J. Phytorem. 20, 552. https://doi.org/10.1080/15226514.2017.1393397.

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Hatano, K., Kikuchi, S., Miyakawa, T., Tanokura, M., Kubota, K., 2008. Separation and characterization of the colored material from sugarcane molasses. Chemosphere 71, 1730–1737. Hatano, K., Komatsu, I., Aoyagi, N., Takahashi, K., Kubota, K., 2013. A study on the self-assembly behavior of dark materials from molasses. Environ. Sci. Pollut. Res. 20, 4009–4017. Hatano, K., Kanazawa, K., Tomura, H., Yamatsu, T., Tsunoda, K., Kubota, K., 2016. Molasses melanoidin promotes copper uptake for radish sprouts: the potential for an accelerator of phytoextraction. Environ. Sci. Pollut. Res. 23, 17656–17663. Hodge, J.E., 1953. Chemistry of browning reactions in model systems. J. Agric. Food Chem. 1, 928–943. Jack, F., Bostock, J., Tito, D., Harrison, B., Brosnan, J., 2014. Electrocoagulation for the removal of copper from distillery waste streams. J. Inst. Brew. 120, 60–64. Jain, R., Srivastava, S., 2012. Nutrient composition of spent wash and its impact on sugarcane growth and biochemical attributes. Physiol. Mol. Biol. Plants 18 (1), 95–99. Jain, N., Bhatia, A., Kaushik, R., Kumar, S., Joshi, H.C., Pathak, H., 2005. Impact of post-methanation distillery effluent irrigation on groundwater quality. Environ. Monit. Assess. 110, 243–255. Jason Dwyer, L., 2008. The degradation of dissolved organic nitrogen associated with melanoidin using a UV/H2O2 AOP. Chemosphere 71 (9), 1745–1753. Juwarkar, A., Dutta, S.A., 1989. Impact of distillery effluent application to land on soil microflora. Environ. Monit. Assess. 15, 201–210. Kaushik, A., Nisha, R., Jagjeeta, K., Kaushik, C.P., 2005. Impact of long and short term irrigation of a sodic soil with distillery effluent in combination with bioamendments. Bioresour. Technol. 96, 1860–1866. Kaushik, A., Basu, S., Raturi, S., Batra, V.S., Balakrishnan, M., 2018. Recovery of antioxidants from sugarcane molasses distillery wastewater and its effect on biomethanation. J. Water Proc. Eng. 25, 205–211. Khairnar, P., Chavan, F., Diware, V.R., 2013. Generation of energy from distillery waste water. Int. J. Sci. Spiritual Bus Technol. 2, 30–35. Kim, S.B., Hayase, F., Kato, H., 1985. Decolourisation and degradation products of melanoidin on ozonolysis. Agric. Biol. Chem. 49, 785–792. Kumar, V., Chandra, R., 2018a. Characterisation of MnP and laccase producing bacteria capable for degradation of sucrose glutamic acid-Maillard products at different nutritional and environmental conditions. World J. Microbiol. Biotechnol. 34, 32. Kumar, V., Chandra, R., 2018b. Bacterial assisted phytoremediation of industrial waste pollutants and eco-restoration. In: Chandra, R., Dubey, N.K., Kumar, V. (Eds.), Phytoremediation of Environmental Pollutants. CRC Press, Boca Raton. Kumar, V., Chandra, R., 2020. Bioremediation of melanoidins containing distillery waste for environmental safety. In: Saxena, G., Bharagava, R.N. (Eds.), Bioremediation of Industrial Waste for Environmental Safety. Vol II—Microbes and Methods for Industrial Waste Management. Springer, Berlin. Kumar, V., Chopra, A.K., 2012. Fertigation effect of distillery effluent on agronomical practices of Trigonella foenum-graecum L. (Fenugreek). Environ. Monit. Assess. 184, 1207–1219. Kumar, S., Gopal, K., 2001. Impact of distillery effluent on physiological consequences in the freshwater teleost Channa punctatus. Bull. Environ. Contam. Toxicol. 66, 617–622. Kumar, V., Sharma, D.C., 2019. Distillery effluent: pollution profile, eco-friendly treatment strategies, challenges and future prospects. In: Arora, P.K. (Ed.), Microbial Metabolism of Xenobiotic Compounds. Microorganisms for Sustainability, vol. 10. https://doi.org/10.1007/978-981-13-7462-3_17.

Chapter 1 Advanced oxidation processes for complex wastewater treatment

Kumar, A., Saroj, D.P., Tare, V., Bose, P., 2006. Treatment of distillery spent-wash by ozonation and biodegradation: significance of pH reduction and inorganic carbon removal before ozonation. Water Environ. Res. 78 (9), 994–1004. Kumar, V., Chandra, R., Thakur, I.S., Saxena, G., Shah, M.P., 2020. Recent advances in physicochemical and biological treatment approaches for distillery wastewater. In: Shah, M., Banerjee, A. (Eds.), Combined Application of Physicochemical & Microbiological Processes for Industrial Effluent Treatment Plant. Springer Singapore. https://doi.org/10.1007/978-981-15-0497-6_6. Kumari, K., Ranjan, N., Kumar, S., Sinha, R.C., 2015. Distillery effluent as a liquid fertilizer: a win–win option for sustainable agriculture. Environ. Technol. 37 (3), 381–387. Liakos, T.I., Lazaridis, N.K., 2014. Melanoidins removal from simulated and real wastewaters by coagulation and electro-flotation. Chem. Eng. J. 242, 269–277. Liang, Z., Wang, Y., Zhou, Y., Liu, H., 2009. Coagulation removal of melanoidins from biologically treated molasses wastewater using ferric chloride. Chem. Eng. J. 152, 88–94. Liu, J., Zhang, X., 2014. Comparative toxicity of new halophenolic DBPs in chlorinated saline wastewater effluents against a marine alga: halophenolic DBPs are generally more toxic than haloaliphatic ones. Water Res. 65, 64–72. Liu, M., Zhu, H., Dong, B., Zheng, Y., Yu, S., Gao, C., 2013. Submerged nanofiltration of biologically treated molasses fermentation wastewater for the removal of melanoidins. Chem. Eng. J. 223, 388–394. Mabuza, J., Otieno, B., Apollo, S., Matshediso, B., Ochieng, A., 2017. Investigating the synergy of integrated anaerobic digestion and photodegradation using hybrid photocatalyst for molasses wastewater treatment. Euro-Mediterr. J. Environ. Integr. 2, 17. Mahar, M.T., Khuhawar, M.Y., Baloch, M.A., Jahangir, T.M., 2013. Health risk assessment of heavy metals in groundwater, the effect of evaporation ponds of distillery spent wash: a case study of southern Punjab Pakistan. World Appl. Sci. J. 28 (11), 1748–1756. Malik, S.N., Saratchandra, T., Tembhekar, P.D., Padoley, K.V., Mudliar, S.L., Mudliar, S.N., 2014. Wet air oxidation induced enhanced biodegradability of distillery effluent. J. Environ. Manage. 136, 132–138. Malik, S.N., Ghosh, P.C., Vaidya, A.N., Mudliar, S.N., 2019. Ozone pre-treatment of molasses-based biomethanated distillery wastewater for enhanced biocomposting author links open overlay panel. J. Environ. Manag. 246, 42–50. Mandal, A., Ojha, K., Ghosh, D.N., 2003. Removal of color from distillery wastewater by different processes. Ind. Chem. Eng. Sect. B 45, 264–267. Manickavachagam, M., Sillanpaa, M., Swaminathan, M., Ahmmad, B., 2014. Advanced oxidation processes for wastewater treatment. Int. J. Photoenergy. https://doi.org/10.1155/2014/682767. Manisankar, P., Viswanathan, S., Rani, C., 2003. Electrochemical treatment of distillery effluent using catalytic anodes. Green Chem. 5, 270–274. Mantzavinos, D., Psillakis, E., 2004. Enhancement of biodegradability of industrial wastewaters by chemical oxidation pre-treatment. J. Chem. Technol. Biotechnol. 79, 431–454. rez-Parra, J., Suay, R., 2011. Use of ozone in wastewater treatment Martı´nez, S.B., Pe to produce water suitable for irrigation. Water Resour. Manag. 25 (9), 2109–2124. Martı´nez-Huitle, C.A., Panizza, M., 2018. Electrochemical oxidation of organic pollutants for wastewater treatment. Curr. Opin. Electrochem. 11, 62–71. Martins, R.C., Quinta-Ferreira, R.M., 2014. A review on the applications of ozonation for the treatment of real agro-industrial wastewaters. Ozone Sci. Eng. 36 (1), 3–35.

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Martins, R.C., Pinto, F.L., Castro-Silva, S., Quinta-Ferreira, R.M., 2011. Flocculation, ozonation, and fenton’s process in the treatment of distillery effluents. Environ. Eng. 139, 110–116. Martı´n Santos, M.A., Ferna´ndez Bocanegra, J.L., Martı´n Martı´n, A., Garcı´a Garcı´a, I., 2003. Ozonation of vinasse in acid and alkaline media. J. Chem. Technol. Biotechnol. 78 (11), 1121–1127. Matthews, R.W., 1991. Environment: photochemical and photocatalytic processes. degradation of organic compounds. In: Pelizzetti, E., Schiavello, M. (Eds.), Photochemical Conversion and Storage of Solar Energy. Springer, Dordrecht. Migo, V.P., Del Rosario, E.J., Matsumura, M., 1997. Flocculation of melanoidins induced by inorganic ions. J. Ferment. Bioeng. 83, 287–291. Mishra, V.S., Mahajani, V.V., Joshi, J.B., 1995. Wet air oxidation. Ind. Eng. Chem. Res. 3412–3448. Moraes, B.S., Zaiat, M., Bonomi, A., 2015. Anaerobic digestion of vinasse from sugarcane ethanol production in Brazil: challenges and perspectives. Renew. Sust. Energ. Rev. 44, 888–903. Naik, N., Jagadeesh, K.S., Noolvi, M.N., 2010. Enhanced degradation of melanoidin and caramel in biomethanated distillery spentwash by microorganisms isolated from mangroves. Iran. J. Energy Environ. 1 (4), 347–351. Nataraj, S.K., Hosamani, K.M., Aminabhavi, T.M., 2006. Distillery wastewater treatment by the membrane-based nanofiltration and reverse osmosis processes. Water Res. 40 (12), 2349–2356. Naveen, C., Premalatha, M., 2016. Impact of blend ratio on the co-firing of postmethanated distillery effluent solid waste and low rank Indian coal via analysis of oxidation kinetics through TGA. RSC Adv. 6, 26121–26129 Otieno, B.O., Apollo, S.O., Naidoo, B.E., Ochieng, A., 2016. Photodegradation of molasses wastewater using TiO2–ZnO nanohybrid photocatalyst supported on activated carbon. Chem. Eng. Commun. 6445, 1443–1454. https://doi.org/ 10.1080/00986445.2016.1201659. Otieno, B., Apollo, S., Kabuba, J., Naidoo, B., Ochieng, A., 2019. Ozonolysis posttreatment of anaerobically digested distillery wastewater effluent. Ozone: Sci. Eng. https://doi.org/10.1080/01919512.2019.1593818. Oturan, M.A., Aaron, J., 2014. Advanced oxidation processes in water/wastewater treatment: principles and applications. A review. Crit. Rev Environ. Sci. Technol. 44 (23), 2577–2641. Padoley, K.V., Saharan, V.K., Mudliar, S.N., Pandey, R., Pandit, A.B., 2012a. Cavitationally induced biodegradability enhancement of a distillery wastewater. J. Hazard. Mater. 219–220, 69–74. Padoley, K.V., Tembhekar, P.D., Saratchandra, T., Pandit, A.B., Pandey, R.A., Mudliar, S.N., 2012b. Wet air oxidation as a pretreatment option for selective biodegradability enhancement and biogas generation potential from complex effluent. Bioresour. Technol. 120, 157–164. Pandey, S.K., Tyagi, P., Gupta, A.K., 2008. Physico-chemical analysis of treated distillery effluent irrigation response on crop plant pea (Pisum sativum) and wheat (Triticum aestivum). Life Sci. J. 84–89. Poulopoulos, S.G., Yerkinova, A., Ulykbanova, G., Inglezakis, V.J., 2019. Photocatalytic treatment of organic pollutants in a synthetic wastewater using UV light and combinations of TiO2, H2O2 and Fe(III). PLoS One 14 (5), e0216745. Ramakritinan, C.M., Kumaraguru, A.K., Balasubramanian, M.P., 2005. Impact of distillery effluent on carbohydrate metabolism of freshwater fish, Cyprinus carpio. Ecotoxicology 14, 693–707. Ray, S.G., Ghangrekar, M.M., 2015. Enhancing organic matter removal, biopolymer recovery and electricity generation from distillery wastewater by combining fungal fermentation and microbial fuel cell. Bioresour. Technol. 176, 8–14.

Chapter 1 Advanced oxidation processes for complex wastewater treatment

Rice, R.G., 1995. Applications of ozone for industrial wastewater treatment—a review. Ozone: Sci. Eng. 18 (6), 477–515. Sahu, O.P., Chaudhari, P.K., 2015. Electrochemical treatment of sugar industry wastewater: COD and color removal. J. Electroanal. Chem. 739, 122–129. Sanchez-Galvan, G., Bolanos-Santiago, Y., 2018. Phytofiltration of anaerobically digested sugarcane ethanol stillage using a macrophyte with high potential for biofuel production. Int. J. Phytorem. 20 (8), 805–812. Saner, A.B., Mungray, A.K., Mistry, N.J., 2014. Treatment of distillery wastewater in an upflow anaerobic sludge blanket (UASB) reactor. Desalin. Water Treat. 57 (10), 4328–4344. Sangave, P.C., Pandit, A.B., 2004. Ultrasound pre-treatment for enhanced biodegradability of the distillery wastewater. Ultrason. Sonochem. 11 (3–4), 197–203. Sangave, P.C., Gogate, P.R., Pandit, A.B., 2007a. Combination of ozonation with conventional aerobic oxidation for distillery wastewater treatment. Chemosphere 68 (1), 32–41. Sangave, P.C., Gogate, P.R., Pandit, A.B., 2007b. Ultrasound and ozone assisted biological degradation of thermally pretreated and anaerobically pretreated distillery wastewater. Chemosphere 68, 42–50. Sankaran, K., Premalatha, M., Vijayasekaran, M., Somasundaram, V.T., 2014. DEPHY project: distillery wastewater treatment through anaerobic digestion and phycoremediation—a green industrial approach. Renew. Sustain. Energy Rev. 37, 634–643. Sankaran, K., Premalatha, M., Vijayasekaran, M., 2017. Characterization of distillery wastewater—an approach to retrofit existing effluent treatment plant operation with phycoremediation. J. Clean. Prod. https://doi.org/10.1016/j. jclepro.2017.02.045. Shah, V.B., Joshi, J.B., Kulkarni, P.R., Joshi, D.S., 1989. Aerobic biological treatment of alcohol distillery waste: kinetics and microbiological analysis. Indian Chem. Eng. 31, 61. Singh, S.S., Dikshit, A.K., 2011. Decolourization of anaerobically digested and polyaluminium chloride treated distillery spentwash in a fungal stirred tank aerobic reactor. Biodegradation 22, 1109–1117. Takle, S.P., Naik, S.D., Khore, S.K., Ohwal, S.A., Bhujbal, N.M., Landge, S.L., Kale, B.B., Sonawane, R.S., 2018. Photodegradation of spent wash, a sugar industry waste, using vanadium-doped TiO2 nanoparticles. RSC Adv. 8, 20394. Tateda, M., Fujita, M., Ike, M., Kokubo, S., 2004. Effect of preozonation on improvement of settleability of solid in highly concentrated organic wastewater of Japanese wheat and sweet potato spirit distillery. J. Environ. Sci. 16 (2), 230–233. Tembhekar, P.D., Padoley, K.V., Mudliar, S.L., Mudliar, S.N., 2015. Kinetics of wet air oxidation pretreatment and biodegradability enhancement of a complex industrial wastewater. J. Environ. Chem. Eng. 3 (1), 339–348. Thakur, C., Srivastava, V.C., Mall, I.D., 2009. Electrochemical treatment of a distillery wastewater: parametric and residue disposal study. Chem. Eng. J. 148, 496–505. Tsioptsias, C., Bantia, D.C., Samarasa, P., 2016. Experimental study of degradation of molasses wastewater by biological treatment combined with ozonation. J. Chem. Technol. Biotechnol. 91, 857–864. United States Environmental Protection Agency (USEPA), 2012. U.S. Environmental Protection Agency Endocrine Disruptor Screening Program Universe of Chemicals. United States Environmental Protection Agency, Washington, DC. https://www.epa.gov/endocrine-disruption/universe-chemicals-potentialendocrine-disruptor-screening-and-testing.

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Vineetha, M.N., Matheswaran, M., Sheeba, K.N., 2013. Photocatalytic colour and COD removal in the distillery effluent by solar radiation. Sol. Energy 91, 368–373. Wang, J.L., Xu, L.J., 2011. Advanced oxidation processes for wastewater treatment: formation of hydroxyl radical and application. Crit. Rev. Environ. Sci. Technol. 42 (3), 251–325. Zhang, M., Wang, Z., Li, P., Zhang, H., Xie, L., 2017. Bio-refractory dissolved organic matter and colorants in cassava distillery wastewater: characterization, coagulation treatment and mechanisms. Chemosphere 178, 259–267. Zhang, M., Dong, H., Zhao, L., Wang, D., Meng, D., 2019. A review on Fenton process for organic wastewater treatment based on optimization perspective. Sci. Total Environ. 670, 110–121. Zou, L.Y., Li, Y., Hung, Y.T., 2007. Wet air oxidation for waste treatment. In: Wang, L.K., Hung, Y.T., Shammas, N.K. (Eds.), Advanced Physicochemical Treatment Technologies. Handbook of Environmental Engineering. vol. 5. Humana Press.

Further reading Beltrain de Heredia, J., Domingues, J.R., Partido, E., 2005. Physico-chemical treatment for depuration of wine distillery wastewater (vinasses). Water Sci. Technol. 51 (1), 159–166. Benitez, F.J., Real, F.J., Acero, J.L., Garcia, J., Sanchez, M., 2003. Kinetics of the ozonation and aerobic biodegradation of wine vinasses in discontinuous and continuous processes. J. Hazard. Mater. B 101, 203–218. Chandra, R., Kumar, V., 2017b. Detection of androgenic-mutagenic compounds and potential autochthonous bacterial communities during in situ bioremediation of post methanated distillery sludge. Front. Microbiol. 8, 887. Chaudhari, P.K., Mishra, I.M., Chand, S., 2005. Catalytic thermal treatment (catalytic thermolysis) of a biodigester effluent of an alcohol distillery plant. Ind. Eng. Chem. Res. 44 (15), 5518–5525. Chaudhari, P.K., Mishra, I.M., Chand, S., 2008. Effluent treatment of alcohol distillery: catalytic thermal pretreatment (Catalytic thermolysis) with energy recovery. Chem. Eng. J. 136, 14–24. Dhale, A.D., Mahajani, V.V., 2000. Treatment of distillery waste water after bio-gas generation: wet oxidation. Indian J. Chem. Technol. 7 (1), 11–18. Fito, J., Tefera, N., Van Hulle, S.W.H., 2018. Physicochemical properties of the sugar industry and ethanol distillery wastewater and their impact on the environment. Sugar Tech. https://doi.org/10.1007/s12355-018-0633-z. 13. Gengec, E., Kobya, M., Demirbas, E., Akyol, A., Oktor, K., 2012. Optimization of baker’s yeast wastewater using response surface methodology by electrocoagulation. Desalination 286, 200–209. Hayase, F., Kim, S.B., Kato, H., 1984. Decolorization and degradation products of the melanoidins by hydrogen peroxide. Agric. Biol. Chem. 48 (11), 2711–2717. Kabir, E.R., Rahman, M.S., Rahman, I., 2015. A review on endocrine disruptors and their possible impacts on human health environmental toxicology and pharmacology. https://doi.org/10.1016/j.etap.2015.06.009. Kobya, M., Gengec, E., 2012. Decolourization of melanoidins by a electrocoagulation process using aluminium electrodes. Environ. Technol. 33 (21), 2429–2438. Pena, M., Coca, M., Gonzalez, G., Rioja, R., Garcıa, M.T., 2003. Chemical oxidation of wastewater from molasses fermentation with ozone. Chemosphere 51, 893–900.

Chapter 1 Advanced oxidation processes for complex wastewater treatment

Prajapati, A., Chaudhari, P., 2014. Electrochemical treatment of rice grain based distillery effluent: chemical oxygen demand and color removal. Environ. Technol. 35, 242–248. Prasad, R.K., Srivastava, S.N., 2009. Electrochemical degradation of distillery spent wash using catalytic anode: factorial design of experiments. Chem. Eng. J. 146, 22–29. Sreethawong, T., Chavadej, S., 2008. Color removal of distillery wastewater by ozonation in the absence and presence of immobilized iron oxide catalyst. J. Hazard. Mater. 155, 486–493.

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2 Technological model on advanced stages of oxidation of wastewater effluent from food industry Sushma Davea, Jayashankar Dasb a

Jodhpur Institute of Engineering and Technology, Jodhpur, Rajasthan, India. SOA Deemed to be University, Bhubaneswar, Odisha, India

b

1

Introduction

Wastewater from a treatment plant is free from 80% total suspended solids (TSS) and biological oxygen demand (BOD) but it needs advanced treatment to make the water reusable. Tertiary treatment is a synonym for advanced oxidation processes (AOPs). It was first used in the 1980s for potable water treatment and is very popular nowadays for wastewater. There is a lot of difference between the wastewater generated from domestic use and that from industries. It also varies from industry to industry. Broadly, industries produce large quantities of highly polluted wastewater containing toxic substances that are resistant to both organic and inorganic biological treatment methods, such as aromatic and aliphatic hydrocarbons, phthalates, plasticizers, pesticides, and phenols (Meric et al., 2005; Gogate and Pandit, 2004). When compared among the industrial sector, it is the food industry along with other agro-industries such as fruit processing, coffee, dairy, palm oil, and pulp mill industries, is also arising as important in various countries, thus requiring to be addressed as far as wastewater treatment is concerned. These activities are the biggest water consumers among the industrial sector, which is seen in the development of the environmental ´ lvarez et al., 2011). They include loads problem (Mavrov, 2000; A of organic pollutants such as proteins and fats and are rich in chemicals used in various food processes. These wastewaters, distinguished by their large pollutant content, are an environmental issue if they are discharged without proper treatment. Advanced Oxidation Processes for Effluent Treatment Plants. https://doi.org/10.1016/B978-0-12-821011-6.00002-5 # 2021 Elsevier Inc. All rights reserved.

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Chapter 2 Technological model on advanced stages of oxidation

For example, there could be a severe risk of contagion to surface and groundwater. Because the problem is of extreme environmental conditions, the pollutants need to be either treated or recovered. A wide variety of pollutants released from the food industry are recalcitrant in nature and dead set against bioremediation. Due to the adverse impacts upon the disposal of treated wastewater, there has been an increase in threats to aquatic life and humans as well as the environment (Fig. 1). Before we undergo complete strategic planning for an advanced treatment method for food industry wastewater, there is a need to assess the characteristics of wastewater. As per Metcalf and Eddy (2003), wastewater after treatment should be free of organic matter, TSS, total BOD COD, odor, nitrogen (Ntot), and phosphorus (Ptot) before it is dispersed into any water body. After primary and secondary treatment, between 80% and 90% of TSS and BOD and a small amount of nitrogen and phosphorus are eradicated. Traces of heavy metals are also removed. Tertiary treatment or advanced treatment removes the organic matter left after secondary treatment. The main objective of tertiary treatment is to increase the effluent quality before it is dispersed in surface water bodies or absorbed in the groundwater table so that it may be reused. For a competent and cost-effective method for wastewater treatment in the food industry, advanced oxidation methods are required because just secondary treatment does not make the water fit for disposal in water bodies. A promising method for such treatment is to apply oxidants so as to oxidize all types of contaminants present in the secondary effluent. These are advanced oxidation methods (AOPs), including the photo-Fenton reaction, UV/O3, UV/H2O2, UV/ultrasound, UV/persulfate, electrochemical oxidation, wet air oxidation, photochemical oxidation, etc.

Electrochemical oxidation

Electrocoagulation

Photocatalytic oxidation

Pollutants in waste water

Electro-Fenton

Ozonation

Fig. 1 Advanced oxidation processes in a nutshell.

Wet air oxidation

CO2 and H2O

Chapter 2 Technological model on advanced stages of oxidation

35

Table 1 Standard potentials of common oxidants (Pera-titus et al., 2004; Turhan and Turgut, 2009).

Oxidant

Electrochemical oxidation potential [V]

Atomic oxygen Fluorine (F2) Hydroxyl radical (HO•) Persulfate Permanganate Ozone (O3) Hydrogen peroxide (H2O2) Hypobromous acid (HOBr) Chlorine dioxide (ClO2) Chlorine (Cl2) Oxygen (molecular) Bromine (Br)

2.42 3.03 2.80 2.12 1.68 2.07 1.77 1.59 1.50 1.36 1.23 1.09

The AOPs involve the generation of free radicals due to chemical, photochemical, and photocatalytic reactions. Treatment of these refractory and recalcitrant organic pollutants is not possible without these free radicals (Pera-titus et al., 2004). The AOPs in water are capable of degrading organic pollutants. This improves the quality of wastewater by mineralizing the pollutants along with the conventional biological process. This began back in the 1980s, initially using the hydroxyl radical whereas the application of sulfate was also combined with the hydroxyl radical. AOPs are applied for decontamination because in the process of disinfection, these radicals are known to have a too-short half-life. Therefore, the disinfection is not effective due to very low concentrations (Table 1).

2

Processes for food industry wastewater treatment

AOPs have been used to trim organic matter in wastewater from the food industry. The types of oxidation methods are classified as follows:

2.1

Electrochemical oxidation

Electrochemical oxidation (EC) is an advantageous method of oxidation because it is a treatment type that does not produce chemical compounds that are hazardous in nature after

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Chapter 2 Technological model on advanced stages of oxidation

wastewater treatment. The formation of hydroxyl radicals is the main characteristic feature of electrochemical oxidation (Rizzo, 2011) while it also helps to reduce the coagulant quantity. The efficiency of the electrochemical process depends on the applied current, the solution pH, and the nature and concentration of the contaminants in the water (Rizzo, 2011). A wide variety of electrodes have been used in the electrochemical oxidation of wastewater such as graphite, Pt, TiO2, IrO, titanium, etc. Recently, there has been an enhancement in the direction of use of electrochemical techniques for the treatment of industry effluent so as to reduce organic matter because this technique provides an environmentally benign platform in the wastewater reuse industry. The pollutants in wastewater are recalcitrant and are difficult to be removed by traditional treatment technologies. Therefore, more efficient technologies for removal are required. As an extension of these technologies, an electrochemical AOP (ECAOP) is to mineralize these pollutants. The electrochemical oxidation (EO) of organic matter for wastewater treatment can be achieved in different ways. The electrochemical technique involves redox reactions that take place at both electrodes, the anode and the cathode. At the anode, the usual oxidation of organic matter takes place whereas at the cathode, a reduction of metal takes place. Two different pathways—direct and indirect oxidation—occur in the electrolytic cell and lead to the oxidation of the organic matter. During direct anodic oxidation or electrolysis reaction occurs directly on the anode and involves direct charge transfer reactions between the anode surface and the organic pollutants involved. The mechanism involves the mediation of the electrons, which are capable of oxidizing some organic pollutants at defined potentials more negative the oxygen evolution reaction (OER) potential. Direct electrolysis usually requires prior adsorption of pollutants onto the anode surface, which is the rate-limiting process and doesnot lead to the overall combustion of organic pollutants. When direct electrolysis is conducted at applied potentials lower than the potential of the water oxidation reaction, the electrodes are susceptible to surface poisoning and further inhibiting the ECOAP process. On the other hand, the indirect ECOAP processes are mediated by the in situ electrogeneration of highly oxidant species at the electrode surface. Different kinds of oxidant species can be generated by the ECOAP process such as the electrogeneration of reactive oxygen species, active chlorine species, and some other compounds such as sulfate, phosphate, and hydrogen carbonate, yielding peroxodisulfate, peroxodiphosphate, and

Chapter 2 Technological model on advanced stages of oxidation

37

peroxodicarbonate. The reactions of the generation of oxidants are as follows: Mð
OHÞ ! MO + H + + e 2Cl ! Cl2 ðaqÞ + 2e Cl2 ðaqÞ + H2 O ! HClO + Cl + H + HClO $ H + + ClO H2 SO4  ! S2 O8 2 + 2H + + 2e 2HCO3  ! C2 O6 2 + 2H + + 2e 2PO4 3 ! P2 O8 4 + 2e In indirect electrolysis, the redox reagent can be electrogenerated by either the anodic or the cathodic process. The most popular technique among cathodic processes is the electro-Fenton (EF), in which H2O2 is generated at the cathode with O2 or air feeding while an iron catalyst is also regenerated on the cathode surface. In recent times, other promising technologies based on the electro-Fenton method such as photo-electro-Fenton (PEF), solar-photo-electro-Fenton (SPEF), coagulation based on the dissolution of iron anodes (peroxicoagulation (PC)), ultrasound irradiation, dissolution of heterogeneous catalysts that supply Fe2+ (heterogeneous-EF), and bioremediation, have also received a great deal of attention. The process selection depends on the nature and structure of the electrode material, the experimental conditions, and the electrolyte composition (Table 2).

Table 2 Summary of a few ECAOPs.

Type of treatment/electrode

Treated parameters

References

Electrocoagulation/Al

COD BOD total solids and fecal coliforms COD, BOD, TOC, and turbidity

Roa-morales et al. (2007)

Fenton process/electrocoagulation (electrode Fe and Al) Anaerobic electrocoagulation (AE) and (AAE) anaerobic-aerobic electrocoagulation Conductive-diamond electrochemical oxidation

Color, COD, and TOC

Tezcan et al. (2009) and Kobya et al. (2006) Gengec et al. (2012)

COD

Herna´ndez-ortega et al. (2009)

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Chapter 2 Technological model on advanced stages of oxidation

As an exceptional and ecofriendly method for the removal of persistent organic contaminants, ECAOPs are usually used with electrons as reagents. However, this technique suffers the primary disadvantages of high energy consumption and low current efficiency. Direct or indirect oxidation (including chlorine and S2 O8 2 ) could cause damage to the organic pollutants. Carbon anodes are rarely used because they have poor efficiency in oxidant formation (such as hydroxyl radicals) and a low O2 evolution overpotential. Positive and negative applied potentials were used for the electro-oxidation or electroreduction of the pollutants, respectively. In recent years, the extensive use of electrochemical oxidation techniques in wastewater treatment has been of great interest. Increasing attention has been paid to the employment of DSA in indirect organic pollutant degradation, due to its high current efficiency in chloride ion oxidation. However, the dyes would only be partially mineralized by the active chlorine (provided by the mixture of OCl, HOCl, and Cl2). Fortunately, via water oxidization, a substantial amount of hydroxyl radicals (OH•) with significant oxidant features could form on the surface of the boron-doped diamond electrodes. Here, boron-doped diamond electrodes could be used for thorough mineralization and color removal. Panizza and Cerisola (2007) presented a comparison of the catalytic activities of four electrode materials (BDD, PbO2, Pt, and Ti Ru SnO2) in the anodic oxidation of methyl red. As indicated in the bulk electrolysis, the use of BDD and PbO2 led to the thorough removal of color and COD, whereas only a partial methyl red oxidation was observed using Pt and Ti Ru SnO2. Chen et al. (2005) further confirmed the better stability and activity of Ti/BDD electrodes compared with the Ti/SnO2Sb2O5 electrodes in the oxidization of the pollutants. At present, the electrodes commonly used during electrochemical oxidation include graphite electrodes, iron electrodes, PbO2 electrodes, noble metal electrodes, dimensionally stable anodes (DSAs), and boron-doped diamond electrodes (BDDs), out of which the last two are considered the best. The food industry consumes large amounts of water for many purposes such as production, cleaning, transportation, and refrigeration. According to the product variety, the source, quantity, and composition of the wastewater range significantly. The main streams of wastewater in the pastry industry are from washing the egg-crusher, blender, filler, and wooden box. The pastry wastewater effluent’s characteristic consist of large amounts of total suspended solids, different nitrogen compounds, fats, proteins, oils, resistant organic pollutants, phosphorus, chlorine, and other chemicals used in washing and sanitizing. Depending on the type

Chapter 2 Technological model on advanced stages of oxidation

of pollutant present in wastewater, several treatment methods have been applied to treat wastewater such as adsorption, electrochemical and biological processes, AOPs, etc. Biological processes are commonly used for the treatment of wastewater containing high concentrations of biodegradable organic matter. Although biological processes are effective, they may not be feasible due to a long hydraulic retention time and a large area requirement. Because resistant organic pollutants are fragmented, it is difficult and not feasible to apply traditional treatment methods. Therefore, the methods generally require modification with other processes. The EC process has been successfully conducted for the treatment of food effluents at the industrial scale, but was not sufficient at meeting the discharge limits without applying cotreatment. Some studies have been reported as a combination of EC processes with membrane technologies, thermolysis, and chemical coagulation. AOPs supported by electrochemical processes (EF, photo-EC, and photo-EF) have also been recently considered as alternative methods for wastewater treatment. Using a synergetic effect, the efficiency of the removal of resistant organic pollutants was maximized with minimal operating costs, versatility, high energy efficiency, and simple equipment design. Electrochemical AOPs can also be performed at room temperature and pressure. A combination of electrochemical and AOPs easily oxidize and lead to the mineralization of most organic and inorganic pollutants to produce H2O, CO2, and inorganic ions by the production of hydroxyl radicals. Hydroxyl radicals can be produced by various methods such as chemical, electrochemical, photo-assisted electrochemical, photocatalysis, Fenton, and ozonation.

2.2

Electrocoagulation

Metal electrodes connected to a direct current (DC) power source are dipped in a solution and the electrical current passes through the electrodes in an electrochemical reactor. Thus, various processes occur in the electrochemical reactor: the formation of anodic metal and metal hydroxide cations in the aqueous phase by electrolytic reactions at the electrode surface; the adsorption of colloidal or soluble pollutants onto the surface of the metal hydroxides; and the removal of pollutants, adhesion to bubbles, and eventually sedimentation (Garcia-segura et al., 2017). On the other hand, when pH is acidic, the electrode is attacked by H+ ions, which leads to electrode dissolution. Then, oxygen in the aqueous phase oxidizes ferrous ions (Fe2+) to ferric ions.

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Chapter 2 Technological model on advanced stages of oxidation

Meanwhile, the oxygen evolution reaction occurs at the anode. The Fe2+ and Fe3+ ions are released in the aqueous phase, such as hydrated and hydrolyzed monomeric and polymeric iron forms, that is, Fe(OH)2+, FeOH2+, Fe2 ðOHÞ2 4 + , Fe(OH)4, Fe(H2O)2+, Fe(H2O)5OH2+, and Fe(H2O)4(OH)2+ (Rajkumar et al., 2007). Negatively charged colloidal particles in the effluent get neutralized with hydroxyl species and settle due to the agglomeration of heavy molecules. Organic ligands are also eradicated through a comprehensive process by combining in the way of settling heavily mass chemicals (Garcia-segura et al., 2017; Mohan and Balasubramanian, 2006).

2.3

Electro-Fenton

When H2O2 is added into the aqueous phase during electrocoagulation by using the Fe anode as the Fe2+ source, several competing reactions that contain Fe2+, Fe3+, and H2O2 are involved. They eventually form hydroxyl radicals, and this process is defined as electro-Fenton (Mohan and Balasubramanian, 2006). The hydroxyl radicals are capable of quickly decomposing organic substrates (RH) and cause the chemical degradation of these organic compounds. Meanwhile, ferrous ions are depleted as rapidly as they are produced. On the other hand, ferrous ions can also react with hydroxyl radicals in aqueous solution (Oliveira et al., 2007). For that reason, more ferrous ions are required to maintain hydroxyl radical production, or this process should be carried out intermittently. Several integrated or sequential treatment processes were applied for different wastewater resources. Yao et al. (2017) applied catalytic thermal treatment and coagulation to treat desizing wastewater. A comparative study was made by Gulshin et al. (2017) using methods such as electrocoagulation, electrochemical Fenton, electro-Fenton, and peroxi-coagulation for the decolorization of real textile wastewater also. They studied the energy and removal efficiency of electrochemical wastewater treatment for the leather industry. In this study, sequential and integrated processes, including EF and EC, were applied for the treatment of pastry industry wastewater. The effect of the process period of EF and the current density of the EC process using iron plate electrodes were examined in terms of TOC removal as well as electrode and energy consumption.

Chapter 2 Technological model on advanced stages of oxidation

2.4

Fenton process

Pollution in water bodies caused by organic wastewater has become a serious problem worldwide. AOPs using the Fenton method are among the most effective and suitable methods for the abatement of organic pollutants. Nevertheless, the process has three apparent limitations: the narrow working pH range; the elevated costs and threats associated with the handling, transportation, and storage of reagents (H2O2 and catalyst); and the significant iron sludge-related secondary pollution. To overcome these limitations, the application of the coupled Fenton optimization processes (photoelectro-Fenton, heterogeneous electro-Fenton, heterogeneous photo-electro-Fenton, three-dimensional electro-Fenton) for organic wastewater treatment is more popular. The oxidation mechanism for the Fenton process has been studied for nearly 90 years. Studies show that the Fenton process includes more than 20 chemical reactions (Duesterberg and Mylon, 2008; Pawar and Gawande, 2015) and its generally accepted core reaction is shown in equation below. The highly oxidative hydroxyl radical (•OH) formed from the reaction of H2O2 with Fe2+ under strong acid can quickly and nonselectively degrade most stubborn organic pollutants to carbon dioxide
ska et al., 2015). The oxidation mechanism and water (Krzemin for the Fenton process has been widely used in various kinds of organic wastewater treatment. Fe2 + + H2 O2 + H + ! Fe3 + + H2 O +
OH FeðOHÞ + + hʋ ! Fe2 + + HO
The degradation efficiency of organic pollutants in the Fenton process depends on operation parameters such as wastewater pH, concentration of the Fenton reagent, and the initial organic pollutant concentration, out of which pH is a very important factor because the species catalyst deactivation of organic pollutants in wastewater cannot be done using Fenton at both low and high pH. The data so far studied show that a pH range of 2–4 is capable of achieving the best treatment efficiency (Rodriguez et al., 1999; Pignatello et al., 2007). However, most pH values of organic wastewater are not within this range. In order to achieve the optimal pH range in the Fenton process, high amounts of chemicals are spent for adjusting organic wastewater at pH 2–4 before decontamination, which increases the costs of organic wastewater treatment.

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Chapter 2 Technological model on advanced stages of oxidation

The classical Fenton reagents consist of H2O2 and a homogeneous solution of iron ions. Both are unstable in chemical properties and easy to lose activity, resulting in the waste of reagents. In addition, concentrated H2O2 is explosive and toxic. Therefore, the storage and transportation of concentrated H2O2 and the homogeneous solution of iron ions not only increases the cost of organic wastewater treatment, but also endangers the operators (Huang et al., 2016). Fenton catalytic reactions basically consist of the oxidation of Fe2+ to Fe3+ along with •OH generation and the reduction of Fe3+ to Fe2+. However, the reaction rate of the former is about 6000 times that of the latter (Song et al., 2006), which interrupts the efficient cycling of Fe3+ and Fe2+ and leads to the accumulation of Fe3+ in the solution. The Fe3+ begins to precipitate above pH 3 in the form of oxyhydroxide, which is the so-called iron sludge (Xue et al., 2017). It is difficult to separate and recover iron sludge, which not only causes a serious loss of the iron species and a reduction of catalytic activity, but also causes secondary pollution to the environment (Panizza and Cerisola, 2007). In summary, the Fenton process has three shortcomings: the narrow working pH range; the high costs and risks associated with the handling, transportation, and storage of reagents (H2O2 and homogeneous solution of iron ions); and the significant iron sludge-related secondary pollution. In order to overcome these shortcomings, the Fenton process is optimized and improved unceasingly to form various optimized Fenton processes. Some important operation parameters include the wastewater pH and catalyst, H2O2, and the organic pollutant concentration in various optimized sludge production (Luan, 2017). The results show that the iron ions leaching in the heterogeneous Fenton system are extremely low and the concentrations are much lower than the legal limit of 2 mg/L imposed by the directives of the European Union (Wang et al., 2014; Yang and Tang, 2018). The main reasons were that the iron species with catalytic activity are immobilized on some supports such as zeolite by the hydrothermal method and other preparation methods. The protective effects of the supports prevent the iron species from immersing in the solution, which reduces the loss of the iron species ( Jin et al., 2017). The structure and composition of the heterogeneous Fenton catalyst itself facilitate the electron transfer from the electron donor to Fe3+, accelerating the reduction of Fe3+ to Fe2+ and achieving the efficient cycling of Fe3+ and Fe2+ on the heterogeneous Fenton catalyst. The key to the heterogeneous Fenton reaction is developing long-term stable heterogeneous Fenton catalysts with high catalytic activity that can be used at a wide pH range and are easily

Chapter 2 Technological model on advanced stages of oxidation

H2O2

•OH Pollutant

Fe2+

Oxidative intermediates

CO2+ H2O

Fe3+

separated without the aid of extra energy. To sum up, when the conventional Fenton process was compared with the heterogeneous Fenton process, the latter had many advantages such as low iron ion leaching, efficient cycling of Fe3+ and Fe2+, low iron sludge production, wide working pH range, reusability, and the long-term stability of the catalysts. However, most studies on the heterogeneous Fenton process are conducted at the lab scale, and its industrialization promotion is limited by many disadvantages such as relatively harsh synthetic conditions, complicated synthesis routes, high synthesis costs of heterogeneous Fenton catalysts, and the design of the heterogeneous Fenton reactor. The photo-Fenton process has drawn much attention to the abatement of refractory organic pollutants due to the low iron sludge production (Fig. 2).

2.5

43

Ozonation

The ozonation of water is a well-known technology and the strong oxidative properties of O3 and its ability to effectively oxidize many organic compounds in aqueous solution have been well documented. Contrasting other oxidizing agents such as Cl2, oxidation with O3 leaves no toxic residues that have to be removed or disposed. Ozonation is one of the most versatile and environmentally benign AOP processes used for food wastewater treatment. Although the cost of ozone production is high, attention toward the use of ozone in wastewater treatment has increased considerably in the last few years due to the abundant advantages of this process (Meric et al., 2005). Ozonation can eliminate toxic substances and increase biodegradable organic pollutants while also having strong potential for the decolorization of wastewater. Ozone is very reactive toward compounds incorporating conjugated double bonds (such as C]C, C]N, N]N), which are associated with colored compounds. Ozone is capable of reacting with solutes with strong oxidative properties

Fig. 2 Fenton oxidation process diagram.

44

Chapter 2 Technological model on advanced stages of oxidation

of (O) (Meric et al., 2005; Pera-titus et al., 2004; Oliveira et al., 2007; Turhan and Turgut, 2009). O3 + H2 O ! 2HO
Ozone has many properties desirable for wastewater treatment, including no residue sludge, less risk, single-step decolorization and degradation, ease in handling, and needs less space. At a low pH, ozone entirely reacts with compounds with specific functional groups through selective reaction mechanisms such as electrophilic, nucleophilic, or the dipolar addition reaction mechanism. When the pH range is high, ozone decomposes, yielding hydroxyl radicals, a highly oxidizing species that reacts nonselectively with a wide range of organic and inorganic. Ozonation has been successfully applied for the treatment of winery and distillery wastewater, olive mile wastewater, meat
ska industry wastewater, and molasses wastewater (Krzemin et al., 2015).

2.6

Photocatalytic process

The photocatalytic and photochemical degradation processes are gaining importance in wastewater treatment because these processes result in complete mineralization with operation at mild conditions of temperature and pressure. In the process, the semiconductor material is excited by electromagnetic radiation possessing energy of sufficient magnitude to produce conduction band electrons and valence band holes. The selection of the adequate catalyst must consider the following properties: chemical action, stability, accessibility, usefulness, cost, and lack of toxicity. The surface area and the number of active sites offered by the catalyst (thus the nature of catalyst, that is, crystalline or amorphous is important) for the adsorption of pollutants play an important role in deciding. Several catalytic materials have been studied in photocatalysis (various oxides such as TiO2, ZnO, SnO2, WO3, ZrO, CeO, etc., or some sulfides are used such as CdS, ZnS, etc.) Among the semiconductors reported so far, the outstanding stability and oxidative power make TiO2 the best semiconductor photocatalyst for environmental remediation and energy conversion processes (Gogate and Pandit, 2004; Urtiaga and Ortiz, 2009). This arises from the fact that pH affects the double action of ozone on the organic matter, which may be a direct or indirect (free radical) oxidation pathway (Gulshin et al., 2017). Although there are wide potential applications of TiO2, there are major drawbacks associated with the use of TiO2, leading to toxicity in human beings and reduction in reactivity due to the

Chapter 2 Technological model on advanced stages of oxidation

45

Table 3 Combination of various AOPs for treating food industry wastewater.

Biological and chemical treatment

Target wastewater

References

Diamond electrochemical oxidation Fenton-ozonation/UV/H2O2 Sonication/anaerobic fermentation UV/H2O2/O3

Synthetic melanoidins solution Cassava wastewater Coffee and wine industry Meat industry

Herna´ndez-ortega et al. (2009)

Dissolved air flotation (DAF) with UV/H2O2 or photo-Fenton UV/O3/modified photo-Fenton/ozonation

Olive production

Lean˜o and Babel (2012) Zhang et al. (2019) and Krzemi
nska et al. (2015) De Sena et al. (2009) Lafi et al. (2009) and Andreozzi et al. (2008)

complex behavior of water. TiO2 has an energy bandgap of 3.2 eV, which can be activated by UV radiation with a wavelength up to 387.5 nm. It only requires 1 W/m of light. Photocatalytic degradation occurs through a multistep process that involves the formation of reactive species on the surface of the photocatalyst and the subsequent generation of hydroxyl radicals that results in the mineralization of most organic compounds. Photocatalysis can be explained by the following simplified reaction (with TiO2) (Stasinakis, 2008; Giannakis, 2019) (Table 3). TiO2 + hʋ ! e HO ! H + + HO


2.7

Wet air oxidation

Wet air oxidation, also known as wet oxidation, is referenced for a process in which the oxidation of suspended or dissolved material in the liquid phase is performed by using dissolved oxygen at high temperature. This method is applicable to waste streams that are too diluted to incinerate and too concentrated for biological treatment. Organics + O2 ! CO2 + H2 O + RCOOH∗ Sulfur species + O2 ! SO4 2 Organic Cl + O2 ! Cl + CO2 + RCOOH∗

46

Chapter 2 Technological model on advanced stages of oxidation

Organic N + O2 ! NH3 + CO2 + RCOOH∗ Phosphorus + O2 ! PO4 3 It was proposed and developed by Hollender et al. (2009), and is one of the most economically and technologically viable advanced oxidation processes for wastewater treatment. WAO is proper to a high organic loading at a high flow rate and can partially cover the application range of incineration and biological methods. It has great potential for the treatment of effluent containing a high content of organic matter. By using WAO, the organic pollutants are either partially oxidized into biodegradable intermediates or mineralized to carbon dioxide, water, and innocuous end products. Complex organic compounds are mostly oxidized into carbon dioxide and water along with simpler forms that are biodegradable. The efficiency of aqueous phase oxidation can be largely improved by the use of catalysts, either heterogeneous or homogeneous. Thermal and catalytic wet air oxidation has better energy output than incineration (for sludge or wastewater).

3

Conclusion

The food industry uses large amounts of water for many different purposes, including cooling, cleaning, raw materials, sanitary water for food processing, transportation, cooking, dissolving, auxiliary water, etc. In principle, the water used in the food industry may be used as process and cooling water or boiler feed water. The advanced oxidation processes are becoming popular as a prime technology that, in combination with activated sludge treatment, fulfills strict environmental requirements. It works at high temperature and high pressure and often has a corrosive reaction mixture; therefore, its intensification is a permanent need. The most promising solutions are catalytic wet oxidation and the combination of WO with AOP techniques, both of which generate the most reactive OH ∙ radicals.

References ´ lvarez, P.M., Pocostales, J.P., Beltra´n, F.J., 2011. Granular activated carbon proA moted ozonation of a food-processing secondary effluent. J. Hazard. Mater. 185, 776–783. https://doi.org/10.1016/j.jhazmat.2010.09.088. Andreozzi, R., et al., 2008. Effect of combined physico-chemical processes on the phytotoxicity of olive mill wastewaters. Water Res. 42, 1684–1692. https://doi. org/10.1016/j.watres.2007.10.018.

Chapter 2 Technological model on advanced stages of oxidation

Chen, X., Gao, F., Chen, G., 2005. Comparison of Ti/BDD and Ti/SnO2–Sb2O5 electrodes for pollutant oxidation. J. Appl. Electrochem. 185–191. https://doi.org/ 10.1007/s10800-004-6068-0. De Sena, R.F., et al., 2009. Treatment of meat industry wastewater using dissolved air flotation and advanced oxidation processes monitored by GC–MS and LC–MS. Chem. Eng. J. 152, 151–157. https://doi.org/10.1016/j.cej.2009.04.021. Duesterberg, C.K., Mylon, S.E., 2008. pH effects on iron-catalyzed oxidation using Fenton’s reagent. Environ. Sci. Technol. 42 (22), 8522–8527. Garcia-segura, S., Ocon, J.D., Chong, M.N., 2017. Electrochemical oxidation remediation of real wastewater effluents—a review. Process Saf. Environ. Prot. https://doi.org/10.1016/j.psep.2017.09.014. Gengec, E., et al., 2012. Optimization of baker’s yeast wastewater using response surface methodology by electrocoagulation. Desalination 286, 200–209. https://doi.org/10.1016/j.desal.2011.11.023. Giannakis, S., 2019. A review of the concepts, recent advances and niche applications of the (photo) Fenton process, beyond water/wastewater treatment: surface functionalization, biomass treatment, combatting cancer and other medical uses. Appl. Catal. B Environ. 248, 309–319. https://doi.org/10.1016/j. apcatb.2019.02.025. Gogate, P.R., Pandit, A.B., 2004. A review of imperative technologies for wastewater treatment I: oxidation technologies at ambient conditions. Adv. Environ. Res. 8 (03), 501–551. https://doi.org/10.1016/S1093-0191(03)00032-7. Gulshin, I., et al., 2017. Comparison of various advanced oxidation processes used in remediation of industrial wastewater laden with recalcitrant pollutants. IOP Conf. Ser. Mater. Sci. Eng. 206. https://doi.org/10.1088/1757-899X/206/1/ 012089. Herna´ndez-ortega, M., et al., 2009. A comparison between conductive-diamond electrochemical oxidation and other advanced oxidation processes for the treatment of synthetic melanoidins. J. Hazard. Mater. 164 (1), 120–125. https://doi.org/10.1016/j.jhazmat.2008.07.134. Hollender, J., Zimmermann, S.G., Koepke, S., et al., 2009. Elimination of organic micropollutants in a municipal wastewater treatment plant upgraded with a full-scale post-ozonation followed by sand filtration. Environ. Sci. Technol. 43 (20), 7862–7869. Huang, D., et al., 2016. Science of the total environment combination of Fenton processes and biotreatment for wastewater treatment and soil remediation. Sci. Total Environ. https://doi.org/10.1016/j.scitotenv.2016.08.199. Jin, H., et al., 2017. Oxygen vacancy promoted heterogeneous Fenton-like degradation of ofloxacin at pH 3.2-9.0 by Cu substituted magnetic Fe3O4@FeOOH nanocomposite. Environ. Sci. Technol. 51 (21), 12699–12706. https://doi.org/ 10.1021/acs.est.7b04503 (Epub 2017 Oct 17). Kobya, M., et al., 2006. Treatment of potato chips manufacturing wastewater by electrocoagulation. Desalination 190 (1–3), 201–211. https://doi.org/10.1016/ j.desal.2005.10.006.
ska, D., Neczaj, E., Borowski, G., 2015. Advanced oxidation processes for Krzemin food industrial wastewater decontamination. J. Ecol. Eng. 16 (2), 61–71. https://doi.org/10.12911/22998993/1858. Lafi, W.K., et al., 2009. Treatment of olive mill wastewater by combined advanced oxidation and biodegradation treatment of olive mill wastewater by combined advanced oxidation and biodegradation. Sep. Purif. Technol. 70 (2), 141–146. https://doi.org/10.1016/j.seppur.2009.09.008. ˜ o, E.P., Babel, S., 2012. Effects of pretreatment methods on cassava wastewater Lean for biohydrogen production optimization. Renew. Energy 39 (1), 339–346. https://doi.org/10.1016/j.renene.2011.08.030.

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Chapter 2 Technological model on advanced stages of oxidation

Luan, M., 2017. Treatment of refractory organic pollutants in industrial wastewater by wet air oxidation. Arab. J. Chem. 10, S769–S776. https://doi.org/10.1016/j. arabjc.2012.12.003. Mavrov, V., 2000. Reduction of water consumption and wastewater quantities in the food industry by water recycling using membrane processes*. Desalination 131 (1–3), 75–86. https://doi.org/10.1016/S0011-9164(00)90008-0. Meric, S., et al., 2005. Decolourisation and detoxifying of Remazol Red dye and its mixture using Fenton’s reagent. Desalination 173 (3), 239–248. https://doi.org/ 10.1016/j.desal.2004.09.002. Metcalf, L., Eddy, H.P., 2003. Wastewater Engineering: Treatment and Reuse, fourth ed. McGraw-Hill, New York. Mohan, N., Balasubramanian, N., 2006. In situ electrocatalytic oxidation of acid violet 12 dye effluent. J. Hazard. Mater. 136 (2), 239–243. https://doi.org/ 10.1016/j.jhazmat.2005.11.074. Oliveira, F.H., Osugi, M.E., Paschoal, F.M.M., et al., 2007. Electrochemical oxidation of an acid dye by active chlorine generated using Ti/Sn(1–x)IrxO2 electrodes. J. Appl. Electrochem. 37 (5), 583–592. https://doi.org/10.1007/s10800-006-9289-6. Panizza, M., Cerisola, G., 2007. Electrocatalytic materials for the electrochemical oxidation of synthetic dyes. Appl. Catal. B Environ. 75 (1–2), 95–101. https:// doi.org/10.1016/j.apcatb.2007.04.001. Pawar, V., Gawande, S., 2015. An Overview of the Fenton Process for Industrial Wastewater. JSPM’S Rajarshi Shahu College of Engineering, Pune, pp. 127–136. Pera-titus, M., et al., 2004. Degradation of chlorophenols by means of advanced oxidation processes: a general review. Appl. Catal. B Environ. 47 (4), 219–256. https://doi.org/10.1016/j.apcatb.2003.09.010. Pignatello, J.J., Oliveros, E., Mackay, A., 2007. Advanced oxidation processes for organic contaminant destruction based on the Fenton reaction and related chemistry. Crit. Rev. Environ. Sci. Technol. 36 (1), 1–84. https://doi.org/ 10.1080/1064338050032656. Rajkumar, D., Song, B.J., Kim, J.G., 2007. Electrochemical degradation of Reactive Blue 19 in chloride medium for the treatment of textile dyeing wastewater with identification of intermediate compounds. Dyes Pigments 72 (1), 1–7. https:// doi.org/10.1016/j.dyepig.2005.07.015. Rizzo, L., 2011. Bioassays as a tool for evaluating advanced oxidation processes in water and wastewater treatment. Water Res. 45 (15), 4311–4340. https://doi. org/10.1016/j.watres.2011.05.035. Roa-morales, G., et al., 2007. Aluminum electrocoagulation with peroxide applied to wastewater from pasta and cookie processing. Sep. Purif. Technol. 54 (1), 124–129. https://doi.org/10.1016/j.seppur.2006.08.02554. Rodriguez, J., et al., 1999. Pulp mill effluent treatment by Fenton-type reactions catalyzed by iron complexes. Water Sci. Technol. 40 (11–12), 351–355. https://doi. org/10.1016/S0273-1223(99)00738-6. Song, W., et al., 2006. Decomposition of hydrogen peroxide driven by photochemical cycling of iron species in clay. Environ. Sci. Technol. 40 (15), 4782–4787. Stasinakis, A.S., 2008. Use of selected advanced oxidation processes (AOPs) for wastewater treatment—a mini review. Glob. Nest J. 10 (3), 376–385. € ˘u € € Savas, A., Og € tveren, U.B., 2009. Hybrid processes for the treatment of Tezcan, U., cattle-slaughterhouse wastewater using aluminum and iron electrodes. J. Hazard. Mater. 164 (2–3), 580–586. https://doi.org/10.1016/j.jhazmat. 2008.08.045. Turhan, K., Turgut, Z., 2009. Decolorization of direct dye in textile wastewater by ozonization in a semi-batch bubble column reactor. Desalination 242 (1–3), 256–263. https://doi.org/10.1016/j.desal.2008.05.005.

Chapter 2 Technological model on advanced stages of oxidation

Urtiaga, A., Ortiz, I., 2009. Contributions of electrochemical oxidation to wastewater treatment: fundamentals. J. Chem. Technol. Biotechnol. 84 (12), 1747–1755. https://doi.org/10.1002/jctb.2214. Wang, Q., Tian, S., Ning, P., 2014. Degradation mechanism of methylene blue in a heterogeneous Fenton-like reaction catalyzed by Ferrocene. Ind. Eng. Chem. Res. 53 (2), 643–664. https://doi.org/10.1021/ie403402q. Xue, W., et al., 2017. Nanoscale zero-valent iron coated with rhamnolipid as an effective stabilizer for immobilization of Cd and Pb in river sediments. J. Hazard. Mater.. https://doi.org/10.1016/j.jhazmat.2017.06.028. Yang, B., Tang, J., 2018. Electrochemical oxidation treatment of wastewater using activated carbon electrode. Int. J. Electrochem. Sci. 13, 1096–1104. https:// doi.org/10.20964/2018.01.78. Yao, H., et al., 2017. Controllable preparation and catalytic performance of heterogeneous Fenton-like α-Fe2O3/crystalline glass microsphere catalysts. Ind. Eng. Chem.. https://doi.org/10.1021/acs.iecr.7b03440. Zhang, M., et al., 2019. Science of the total environment a review on Fenton process for organic wastewater treatment based on optimization perspective. Sci. Total Environ. 670, 110–121. https://doi.org/10.1016/j.scitotenv.2019.03.180.

Further reading Martı´nez-Huitle, C.A., Panizza, M., 2018. Electrochemical oxidation of organic pollutants for wastewater treatment. Curr. Opin. Electrochem. https://doi.org/ 10.1016/j.coelec.2018.07.010. Pang, S.-Y., Jiang, J., Ma, J., 2011. Oxidation of sulfoxides and arsenic (III) in corrosion of nanoscale zero valent iron by oxygen: evidence against ferryl ions (Fe(IV)) as active intermediates in Fenton reaction. Environ. Sci. Technol. 45 (1), 307–312. https://doi.org/10.1021/es102401d.

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3 Exploring microbes as bioremediation tools for the degradation of pesticides Prashant Kumar, Shashwati Ghosh Sachan Department of Bio-Engineering, Birla Institute of Technology, Mesra, Ranchi, Jharkhand, India

1

Introduction

The global population is increasing rapidly and there is a lot of pressure on the agricultural sector to feed this vast population. The agricultural sector is taking every measure to ensure that it meets the ever-increasing demand for agricultural products. To ensure crop safety, several steps are taken by individual farmers and corporations. One of them is to use pesticides to protect the crop from pests, as they are the greatest threat to crops. Hence, the global usage of pesticides has increased exponentially. Pesticides are chemicals that are used to control pests (insects, weeds, fungi, bacteria). They are used to kill, repel, and prevent pests as they are made up of various toxic chemicals. Most of these chemicals are designed in such a way as to disturb the physiological activities of the target organism, leading to dysfunction and reduced vitality. Pesticides largely include herbicides, insecticides, nematicides, molluscicides, rodenticides, and fungicides. The greater use of pesticides for high agricultural production as well as pesticide characteristics such as high lipophilicity, bioaccumulation, long half-life, and the potential to travel long range with water have increased the chances of contaminating the air, water, and soil many-fold, even after many years of application (Rajendran, 2003). Pesticides are now becoming a major pollutant and their proper disposal and degradation are matters of great concern. The pesticide manufacturing industries need to treat their wastewater effluents containing pesticides and their byproducts before releasing them back into the environment. Advanced Oxidation Processes for Effluent Treatment Plants. https://doi.org/10.1016/B978-0-12-821011-6.00003-7 # 2021 Elsevier Inc. All rights reserved.

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Chapter 3 Exploring microbes as bioremediation tool

There is a need for better management of pesticide degradation. The fate of pesticides in the environment after application may be photodegradation, leaching, root uptake by plants, runoff by water, chemical degradation, or microbial degradation. In this context, microbes can play a vital role in providing an ecofriendly and cost-effective solution. Microbes are considered among the first life forms on this planet. Microbes are distributed in different habitats and have a much diversified metabolic capability that gives them an advantage in surviving harsh conditions. The nutritional sources of microbes are also very variable, as microbes can survive on different types of material that include toxic products. Diverse metabolic pathways and varied enzymatic capabilities allow microorganisms to break down toxic products into their elementary form, known as complete mineralization. This property of microbes can be exploited as the tool for degrading pesticides that are polluting water bodies and soils.

1.1

Pesticide problem: An introduction and overview

The use of pesticides is not new, as it has been known for thousands of years that the Greeks and Romans intentionally used sulfur, mercury, arsenic, and other plant products as pesticides. The application of pesticides is now global. Every major commercial agricultural firm and even individual farmers are applying pesticides in their field; some are using them in maintaining their lawns and gardens. There are approximately 1.8 billion people that are engaged in the agricultural sector due to extreme pressure to increase productivity; to safeguard the produced crops, the use of pesticides has increased significantly. After World War II, some chemicals were very popular as pesticides, such as DDT (dichlorodiphenyltrichloroethane), BHC (benzene hexachloride), aldrin, dieldrin, endrin, and 2,4-D (2,4-dichlorophenoxyacetic acid). Currently, the global use of pesticides is approximately 2 million tons per year, of which the major share of 45% is used by Europe followed by the United States (24%) while the rest of the world’s share is 25%. In Asia, pesticide use is increasing at an alarming rate. Currently, China is the largest consumer of pesticides in Asia followed by Korea, Japan, and India. In India, pesticide use is approximately 0.5 kg/hectare, and a large portion of that is from organochlorine pesticides owing to the warm, humid climatic conditions (Kandpal, 2014). Based on their chemical nature, pesticides are classified as organochlorines, organophosphates, carbamates, pyrethroids,

Chapter 3 Exploring microbes as bioremediation tool

53

phenyl amides, phenoxyalkonates, triazines, benzoic acid, phthalimides, and dipyrids; their general characteristics are listed in Table 1. The review statistics on the use of different pesticides show that 40% of all pesticides used belong to the organochlorine class of chemicals (Gupta, 2004) and overall, organophosphorus compounds account for 38% of the total pesticides used globally. An examination of the effects of different classes of pesticides leads to the conclusion that many of them are responsible for hypertension, cardiovascular disorders, and other health-related problems in humans. Many of the organochlorine molecules are carcinogenic and neurotoxic (Kaiser, 2000). There are basically three methods to remove pollutants in different matrices, depending on the characteristics of the remediation processes used. Currently, biological, chemical, and physical methods are used. Biological remediation or bioremediation is an attractive technology that results in either the complete mineralization of xenobiotic compounds or their conversion to less-toxic products. It is considered to be low cost and environmentally friendly as compared to physical or chemical methods for removing contaminants (Nwankwegu and Onwosi, 2017; Sinha et al., 2011).

Table 1 General characteristics of pesticides (Badii and Landeros, 2007).

Pesticides

Characteristics

Examples

Organochlorines

Lipid soluble, accumulate in the fatty tissue of animals, are transferred through the food chain; toxic to a variety of animals, long-term persistent Soluble in organic solvents but also in water. They infiltrate and reach groundwater, less persistent than chlorinated hydrocarbons; some affect the central nervous system. They are absorbed by plants, transferred to leaves and stems, which are supplied to leaf-eating insects as they feed on leaves Carbamate acid derivatives; kill a limited spectrum of insects, but are highly toxic to vertebrates. Relatively low persistence Affect the nervous system; are less persistent than other pesticides; are the safest in terms of their use, some are used as household insecticides

DDT, aldrin, lindane, chlordane, mirex

Organophosphates

Carbamates Pyrethroids

Malathion, methyl parathion, diazinon

Sevin, carbaryl Pyrethrins

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Chapter 3 Exploring microbes as bioremediation tool

1.2

Pesticide toxicity in humans

Pesticides are toxic substances that are poisonous for pests but according to a study, only about 3% of pesticides reach the targeted organism (George Tyler Miller, 2004). The remainder misses the targeted organism and then either remains in the soil or mixes with running water and groundwater, contaminating it. Humans are most likely to be exposed to the remaining pesticides either by direct contact or by indirect contact through consumption of contaminated food and water. The route of exposure can also be different, through dermal, gastrointestinal, or respiratory routes. Pesticide exposure can cause various hazardous effects on human health. Pesticides can also cause toxicity in humans that can lead to several disease conditions. The toxicity can be acute or chronic. When the toxicity is acute or a single exposure through ingestion, the results can be respiratory tract irritation, sore throat and/or cough, allergic sensitization, eye and skin irritation, nausea, vomiting, diarrhea, headache, loss of consciousness, extreme weakness, seizures, and/or death. When the toxicity is chronic for a longer duration of time at a low concentration, it can cause Parkinson’s disease, asthma, depression and anxiety, attention deficit hyperactivity disorder (ADHD), and cancer, including leukemia and nonHodgkin’s lymphoma. Several pesticides such as organophosphates are meant to disrupt the nervous system of pests but can also cause a similar effect in humans when exposed. Pesticides such as organochlorines are very stable in the environment and take a long time to degrade naturally. Microbes can help in speeding the degradation processes of these pesticides.

2 Current remediation methods The complete removal of pesticides either from industrial effluents or from sites contaminated by agricultural activities is quite challenging. The current remediation methods include the use of ion exchange resins or biosorbents for the removal of pesticides and toxic pollutants, but these processes are sensitive to external environmental conditions and need much more effort to maintain optimally (Wang and Chen, 2009). The conventional methods of remediation consist of chemical precipitation, coagulation, adsorption on matrices, adsorption by activated charcoal, ion exchange, and reverse osmosis (US EPA, 2007). They are only applicable, however, when the pollutant or toxic chemical is present in a high amount, whereas pesticide pollution can occur at a much lower concentration that cannot be solved by conventional methods. Some of the methods such as adsorption by activated

Chapter 3 Exploring microbes as bioremediation tool

charcoal can be useful to treat wastewater, but the process is too slow and the regeneration process is costly (Mohan et al., 2001). The ion exchange methods are based on different resins that are also costly, but they have the advantage of regeneration of resins so they can be used several times (Chiarle et al., 2000). There are some physiochemical methods for the degradation of pesticides that can be applied to the contamination site to check the spread of pesticides to other sites. Photodegradation of pesticides can be accomplished by direct photolysis, wherein the pesticide absorbs light energy, then is excited and can be transformed depending on the availability of activation energy. In indirect photodegradation, the pesticide reacts with other species that have been produced photochemically, causing degradation or conversion into other products. Photodegradation occurs in shallow surface layers of soil (Rani and Sud, 2015). Chemical treatments involve the application of chemical agents to promote the extraction or conversion of hazardous chemicals such as pesticides to harmless or less harmful chemicals through chemical reactions (Hamby, 1996). Depending on the matrix to remedy, the economic costs can be very high. However, all these methods have several disadvantages such as the use of chemical catalysts like titanium dioxide, and the use of expensive technology in the case of ozone. For some pesticides, the alkaline hydrolysis method is used, such as in the case of organophosphates. This method must include a rigorous control of the conditions under which the experiments are performed, such as maintenance of alkaline pH, as well as control on the presence of complexes formed with metal ions, which involves the formation of secondary pollutants. These conventional physicochemical approaches are generally expensive and the remediation process is often incomplete due to the conversion of the parent compound to metabolites, which are more persistent and equally or more toxic to nontarget organisms. Microbes can be used as tools for pesticide degradation, and this can be a very promising method. The microbial method of pesticide degradation needs to be optimized and new capable microbes with a high efficiency of degradation need to be discovered. The microbial method is cheap, sustainable, ecofriendly, and less labor-intensive; it can also be applied for in situ remediations.

3

Bioremediation of pesticides: Importance of microorganisms

Biological remediation or bioremediation is an attractive technology that results in the complete conversion of organic compounds into less harmful end products such as CO2 and H2O. It

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is considered low-cost and environmentally friendly compared to physical or chemical methods for removing contaminants (Nwankwegu and Onwosi, 2017). There are basically three types of bioremediation with microorganisms: remediation through improved natural attenuation (taking advantage of the natural capacities of the microorganisms present in the matrix); bioaugmentation (introduction of nonnative and/or genetically modified microorganisms); and biostimulation (addition of electron acceptors or nutrients) (Helbling, 2015; Nwankwegu and Onwosi, 2017). Biodegradation may be referred to as the complete mineralization of organic contaminants into carbon dioxide, water, inorganic compounds, and cell proteins or the transformation of complex organic contaminants into other simpler organic compounds by biological agents such as microorganisms (Gan et al., 2009). Many indigenous microorganisms in water and soil are capable of degrading hydrocarbon contaminants (Marican and Dura´n-Lara, 2018). Bioagents such as bacteria, fungi, and archaea are the main microbes that can decontaminate a site. The microbes that have a resistant gene for a pesticide pollutant can survive in that contaminated site. They can then utilize the pesticide as a food source, break it down as a source of carbon, and increase the biomass. They do not store or collect the pesticide, so this helps in decontaminating the site. Bioremediation is a biotechnological method that can ensure the maintenance of environmental balance and stability.

3.1

Pesticide degradation by bacteria

The microbial degradation of pesticides leads to either complete mineralization of the compound and the release of CO2 and H2O due to the oxidation of the pesticide by degrading enzymes or conversion to less-toxic byproducts. Bacteria use the released energy to increase their number and biomass. Most of the bacterial species that are equipped with pesticidedegrading mechanisms are from Pseudomonas, Arthrobacter, Azotobacter, and Burkholderia. The details of pesticide-degrading bacteria are given in Table 2. According to Abo-Amer (2012) Pseudomonas is the most efficient bacterial genus that can degrade a variety of pesticides. Its degradation capacity is up to 90%–99%. Anwar et al. (2009) isolated a bacterial strain, C2A1 Bacillus pumilus, from soil that was found to be highly effective in degrading chlorpyrifos and its first hydrolysis metabolite 3,5,6trichloro-2-pyridinol (TCP). The bacterial strain showed 90% degradation of TCP (300 mg/L) within 8 days of incubation. Akbar and Sultan (2016) found two bacterial strains, Achromobacter xylosoxidans (JCp4) and Ochrobactrum sp. (FCp1). These isolates were

Chapter 3 Exploring microbes as bioremediation tool

57

Table 2 Pesticide-degrading bacteria (Prabha et al., 2017).

Bacteria

Pesticides

Acinetobacter calcoaceticus Acinetobacter johnsonii (MA-19) strain Acinetobacter Acidomonas sp. Azospirillum and Pseudomonas A. xylosoxidans JCp4 and Ochrobactrum sp. FCp1 Bacillus thuringiensis Burkholderia cepacia strain CH-9 Bacillus sp. and Chryseobacterium joostei Enterobacter aerogenes Escherichia coli Ochrobactrum Photosynthetic bacterium (GJ-22) Paracoccus sp. strain Pseudomonas Pseudomonas putida Pseudomonas mendocina Pseudomonas and Alcaligenes sp. Rhodococcus bacteria Rhodobacter sphaeroides Sphingobium japonicum Stenotrophomonas maltophilia Sphingomonas Sphingobacterium sp. Sphingobium sp. JQL4-5 Sphingomonas yanoikuyae Vibrio and Shewanella

Bifenthrin Organophosphate pesticides Esbiothrin Allethrin Ethion Chlorpyrifos Malathion Imidacloprid and metribuzin Lindane, methyl parathion, and carbofuran Bifenthrin, cypermethrin BHC, DDT, endosulfan, HCH isomers, and 2,4-D Triazophos Cypermethrin (CMP) Pyridine Endosulfan, atrazine Permethrin and cypermethrin Pesticides Herbicide 2,4-D, endosulfan, lindane, chlorpyrifos Para-nitrophenol Chlorinated pesticides, herbicides, and fungicides Hexachlorocyclohexane DDT and endosulfan DDT DDT Fenpropathrin Carbamate and pyrethrin Methyl parathion

able to degrade 84.4% and 78.6% of the initial concentration of chlorpyrifos in 8–10 days. Abraham and Gajendiran (2019) studied the degradation of fipronil and its metabolite fipronil sulfone using Actinomycetes.

3.2

Cyanobacteria for pesticide biodegradation

Cyanobacteria have photoautotrophic and nitrogen fixation capability, which gives them an advantage over other microbes. Due to these photoautotrophic and nitrogen fixation capabilities, they are self-sufficient for growing and can efficiently grow and

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Table 3 Pesticide-degrading cyanobacteria (Prabha et al., 2017).

Pesticide/herbicide/ insecticide

Cyanobacteria Anabaena PD-1 Aulosira fertilissima ARM 68 and Nostoc muscorum ARM 221 Anabaena sp. strain PCC 7120 and Nostoc ellipsosporum N. muscorum and S. platensis Anabaena sp., Arthrospira fusiformis, Leptolyngbya boryana, Microcystis aeruginosa, Nostoc punctiforme, Spirulina platensis Microcystis novacekii Nostoc sp. MM1, Nostoc sp. MM2, Nostoc sp. MM3, Nostoc muscorum and Anabaena sp. Phormidium valderianum BDU 20041 Synechococcus elongatus, Anacystis nidulans, and Microcystis aeruginosa Spirulina sp. Synechocystis sp.

Polychlorinated biphenyls (PCBs) Monocrotophos and malathion Lindane (g-hexachlorocyclohexane) Malathion Glyphosate Methyl parathion Fenamiphos Chlorpyrifos Organophosphorus and organochlorine insecticides Glyphosate Anilofos

survive at the contaminated site. Also, many of them are capable of pesticide degradation. Some of the pesticide-degrading cyanobacteria are listed in Table 3. Ibrahim et al. (2014) used three strains of filamentous Cyanobacteria to study their growth and the utilization of the organophosphorus pesticide malathion. A sharp decrease in the growth of the algal strains was observed by increasing the concentration of malathion. Among them, Nostoc muscorum tolerated different concentrations and was recorded as the most efficient strain for biodegradation (91%) of this compound. This study clarified that N. muscorum with its capability to utilize malathion as a sole phosphorous source is considered an inexpensive and efficient biotechnology for the remediation of organophosphorus pesticides from contaminated wastewater.

3.3

Fungi for pesticide biodegradation

Fungal species are very successful in degrading pesticides, as they secrete extracellular enzymes for pesticide degradation. Different fungal species from nature are identified for the biodegradation of different pesticides. The studies suggest that fungal

Chapter 3 Exploring microbes as bioremediation tool

Table 4 Fungi capable of degrading pesticide (Prabha et al., 2017).

Fungi

Pesticide/herbicide/insecticide

Aspergillus Fusarium oxysporum, Lentinula edodes, Penicillium brevicompactum, and Lecanicillium saksenae Fusarium verticillioides Mortierella sp. strains W8 and Cm1-45 Trichoderma viride and T. harzianum Rot fungi White-rot fungi

Endosulfan Terbuthylazine, difenoconazole, and pendimethalin

Lindane Endosulfan Pirimicarb Methomyl and diazinon Aldrin, aldicarb, alachlor, atrazine, chlordane, diuron, DDT, dieldrin, gamma-hexachlorocyclohexane (g-HCH), heptachlor, lindane, mirex, metalaxyl, terbuthylazine

strains Fusarium oxysporum, Lentinula edodes, Penicillium brevicompactum, and Lecanicillium saksenae are efficient in degrading pesticides (Hai et al., 2012). A list of pesticidedegrading fungi is given in Table 4.

3.4

Mechanism of biodegradation

The mechanism of biodegradation involves the metabolic capability of the microbes to remove the pesticide from the contamination site. It is a very cost-effective technique. The microbes can transform the pesticide into simpler compounds. The microbial transformation is driven by the requirement of energy. As much as diverse and complex pesticides are present, there may be more diverse microbes that are capable of degrading the pesticides. The microbe interacts both chemically and physically with the pesticide and transforms or degrades the target pesticide. Microbes, especially fungi and bacteria, produce extracellular enzymes to degrade pesticides. For the degradation of pesticides, there are some enzyme systems responsible such as hydrolases, esterases (also hydrolases), and the mixed-function oxidases (MFO). Some of the enzymes and respective bacteria are listed in Table 5. The metabolism process of pesticides may involve three phases. In the first phase, the parent compound goes through oxidation, reduction, or hydrolysis and is transformed into a more water-soluble and less-toxic product. In the second phase, the transformed product is conjugated with an amino

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Table 5 Enzymes involved in the degradation of pesticides.

Enzyme

Organism

Pesticide

References

Oxidoreductases (Gox)

Pseudomonas sp. LBr Agrobacterium strain T10

Glyphosate

Jacob et al. (1988)

Monooxygenases: ESd Ese

Mycobacterium sp. Arthrobacter sp.

Sutherland et al. (2002) Weir et al. (2006)

P450

Pseudomonas putida

Dioxygenases (TOD) Phosphotriesterases: OPH/ OpdA

Pseudomonas putida Agrobacterium radiobacter Pseudomonas diminuta Flavobacterium sp. Sphingobium sp. Sphingomonas sp. Pseudomonas sp. ADP Nocardioides sp. Sphingobium sp. Shingomonas sp. Ralstonia eutropha

Endosulfan and endosulphato Endosulfan, aldrin, malathion, DDDT, and endosulphato Hexachlorobenzene and pentachlorobenzene Herbicides trifluralin Insecticides phosphotriester

Hexachlorocyclohexane (b and d isomers) Herbicides chloro-s-triazine Herbicides chloro-s-triazine Hexachlorocyclohexane (g isomers) 2,4-Dichlorophenoxyacetic acid and pyridyl-oxyacetic Dicamba

Nagata et al. (2005) Sharma et al. (2006) de Souza et al. (1996) Mulbry et al. (2002) Raina and Hall (2008)

Haloalkane dehalogenases: LinB AtzA TrzN LinA TfdA DMO

Pseudomonas maltophilia

Chen et al. (2002) Yeh et al. (1977) Harcourt et al. (2002) Serdar et al. (1985) Mulbry et al. (1986)

Streber et al. (1987) Herman et al. (2005)

group or sugar group to form a more water-soluble and less-toxic product. In the third phase, the bacteria and fungi produce extracellular hydrolytic enzymes for the further break down of the compound (Van Eerd et al., 2003)

3.5

Factors affecting microbial bioremediation

Bioremediation involves immobilizing, detoxifying, removing, degrading, or altering the pollutant pesticide from the soil, water bodies, and effluents by the action of microbes. Microbes utilize their enzymatic pathways as a catalyst for pesticide degradation. Many factors are responsible for the efficient degradation of

Chapter 3 Exploring microbes as bioremediation tool

pesticides. The chemical nature, toxicity, and quantity of the pesticide are also responsible for their degradation. Microbes are biological entities and they need optimal physiochemical conditions, so biotic and abiotic factors such as pH, salinity, temperature, light, nutrition, and the presence of oxygen and metal ions play a crucial role in the efficiency of pesticide degradation.

4 4.1

Microbial bioremediation strategies for pesticide degradation In situ bioremediation

The application of bioremediation methods to the contamination site to check the pesticide contamination and to employ microbes at optimal conditions to decontaminate the site is called in situ bioremediation. Some in situ bioremediation methods are listed below: Bioaugmentation: Cultured microbes or genetically modified microbes are added to the contamination site to degrade the pesticides. This process is adopted by many industrial wastewater treatment plants and municipal wastewater plants. Bioventing: It is the mixing and regulating of air or oxygen to the soil. It stimulates the natural in situ bioremediation of pesticides by providing oxygen and air through wells. Biostimulation: This strategy involves the stimulation of indigenous microbes or naturally present bacteria on the site by adding specific nutrients or by supplying fertilizers. Supply of environmental stress factors such as pH, temperature, and oxygen also help to stimulate the microbial operon systems which can help in the production of required enzymes for the degradation of a targeted substance (Abatenh et al., 2017).

4.2

Ex situ bioremediation

Ex situ bioremediation involves methods in which the samples are composted and the microbes break down the pesticide. The pollutant material is transported to the site of degradation. It is a time-consuming and labor-intensive process.

4.3

Microbial consortia for bioremediation

Pesticide degradation in the environment can be achieved by a consortium of microbes. This is more efficient than a single microbe because the products that are produced as a byproduct

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of one bacterium can be utilized by other bacteria. These consortia work synergistically to give maximum efficiency.

4.4

Genetic manipulation of microbes for efficient bioremediation

Microbes can decontaminate pesticide contamination because their metabolic activities allow them to degrade the pesticides. Sometimes microbes are introduced to a new pollutant or toxic pesticide that they have never interacted with before. In that case, the microbes that are resistant toward that pesticide will survive and slowly change their metabolic pathways to survive in the stress condition. If the necessary genes for resistance and the production of vital enzymes are supplied from outside, the efficiency will increase. Successful gene manipulation will provide the microbe with functional genes for efficient degradation. This can be achieved by the insertion of a genomic gene or plasmid in the microbe that will help the microbe in adapting to the changing environment. Many essential genes are identified on the plasmid and genome of various microbes.

4.5

Characteristics of microorganisms suitable for remediation

In microbes, bacteria have the best suitable characteristics for pesticide degradation, as many bacteria possess a catabolic gene that provides resistance against that particular pesticide as well as the enzymes that can help in degrading that pesticide. There are some microbes listed in Table 6 with their catabolic gene. It is well studied that when a fraction of a soil biota is in direct contact with a pesticide for a long period of time, it will develop the ability to degrade that pesticide. Different biological systems adapt differently in a variable environment. Microbes use this pesticide as an electron donor and carbon source.

5 Pros and cons of using microbes in bioremediation Microbes are omnipresent in the biosphere. They can be found in diverse habitats such as deep thermal vents or an organism’s gut. They can resist and survive in any adverse condition and this ability makes them a suitable candidate for bioremediation. The microbes utilize a pesticide as their main source of carbon. For this, they have various enzymes to break down pesticides and

Chapter 3 Exploring microbes as bioremediation tool

Table 6 Microbes with the ability to degrade bacteria with their catabolic genes (Singh and Walker, 2006).

Gene

Organism

Opd opaA opdA adpB pepA hocA pehA Phn ophB ophC2 OpdB Imh Mpd Oph Mph MpdB opdE

Pseudomonas diminuta Alteromonas spp. A. radiobacter Nocardia sp. Escherichia coli Pseudomonas monteilii Burkholderia caryophylli Bacillus cereus Burkholderia sp. JBA3 Stenotrophomonas sp. SMSP-1 Lactobacillus brevis Arthrobacter sp. scl-2 Ochrobactrum sp. Yw28, Rhizobium radiobacter Arthrobacter sp. Arthrobacter sp. L1 (2006) Burkholderia cepacia Enterobacter sp. Fungi Aspergillus niger Penicillium lilacinum

A-opd P-opd

the products of that are then processed by various metabolic pathways. When microbes take up the pollutant as their food source, they increase their number. Their biomass also increases rapidly on the site of application and this leads to the decontamination of the application site. It eliminates the need for moving polluted material and it can be applied on site. The methods to use microbes as tools for degradation are cost-effective in comparison to conventional and physiochemical methods. Through this method, the complete mineralization of pesticides can be achieved, as this is a highly specific method and the chances of the formation of toxic byproducts are very low. However, despite having many advantages, the methods of bioremediation have some disadvantages also. They can be applied to only biodegradable substances, and not all pollutants can be easily biodegradable. The results that are achieved in a pilot-scale study cannot be replicated in the field. Sometimes, the biodegradation product is more toxic and stable than the

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parent compound. The microbes are sometimes specific to pesticides they degrade, so whenever there is a mixture of pesticide pollutants present, a single microbe is not efficient to degrade it. Therefore, a consortium has to be deployed and finding a suitable consortium is a difficult task. Another problem is of the abiotic and biotic factors on which bioremediation depends: its efficiency is optimal only when these conditions are optimal. After pesticide degradation, the bacterial biomass degrades and may cause a serious issue of biofouling. When a genetically modified microbe is used for biodegradation, its stability is also a factor; it can also be a concern for environment safety.

6 Conclusion and future prospects The pesticide contamination of soil and water poses a great concern for the whole world. The solution to this problem is the various remediation processes, but bioremediation is a better choice because of its many advantages over other remediation processes. The microbes can play a much larger role in curbing soil and water contamination by pesticides. There is a lot of potential for the search of novel microbial species or consortia that can easily and efficiently degrade the pesticides present in the soil and water bodies. Their metabolic and enzymatic procedures for the breakdown of pesticides can be studied in detail and their ability to degrade pesticides can be harnessed on an industrial level. The use of microbes to check the contamination from pesticides and other harmful chemicals is a promising approach. However, various genetic manipulation techniques can be explored to optimize the metabolic pathways, the enzyme production, and the growth of the microbe, which can increase the efficiency. The genetic approach can be used to study the genes that are responsible for biodegradation as well as their locations in the microbial genome. That will help us enhance the capacity of local microbes present in the polluted site and they can locally degrade the pesticides, which will be more ecofriendly.

References Abatenh, E., Gizaw, B., Tsegaya, Z., Wassie, M., 2017. Application of microorganisms in bioremediation—review. J. Environ. Microbiol. 1, 2–9. Abo-Amer, A.E., 2012. Characterization of a strain of Pseudomonas putida isolated from agricultural soil that degrades cadusafos (an organophosphorus pesticide). World J. Microbiol. Biotechnol. 28, 805–814. Abraham, J., Gajendiran, A., 2019. Biodegradation of fipronil and its metabolite fipronil sulfone by Streptomyces rochei strain AJAG7 and its use in bioremediation of contaminated soil. Pestic. Biochem. Physiol. 155, 90–100.

Chapter 3 Exploring microbes as bioremediation tool

Akbar, S., Sultan, S., 2016. Soil bacteria showing a potential of chlorpyrifos degradation and plant growth enhancement. Braz. J. Microbiol. 47, 563–570. Anwar, S., Liaquat, F., Khan, Q.M., Khalid, Z.M., Iqbal, S., 2009. Biodegradation of chlorpyrifos and its hydrolysis product 3, 5, 6-trichloro-2-pyridinol by Bacillus pumilus strain C2A1. J. Hazard. Mater. 168, 400–405. Badii, M., Landeros, J., 2007. Pesticides that affect human health and sustainability. CULCyT 19, 21–34. Chen, X., Christopher, A., Jones, J.P., Bell, S.G., Guo, Q., Xu, F., Roa, Z., Wong, L.L., 2002. Crystal structure of the F87W/Y96F/V247L mutant of cytochrome P-450cam with 1,3,5-trichlorobenzene bound and further protein engineering for the oxidation of pentachlorobenzene and hexachlorobenezene. J. Biol. Chem. 277, 37519–37526. Chiarle, S., Ratto, M., Rovatti, M., 2000. Mercury removal from water by ion exchange resins adsorption. Water Res. 34, 2971–2978. de Souza, M.L., Sadowsky, M.J., Wackett, L.P., 1996. Atrazine chlorohydrolase from Pseudomonas sp. strain ADP: gene sequence, enzyme purification, and protein characterization. J. Bacteriol. 178, 4894–4900. Gan, S., Lau, E.V., Ng, H.K., 2009. Remediation of soils contaminated with polycyclic aromatic hydrocarbons (PAHs). J. Hazard. Mater. 172, 532–549. George Tyler Miller, 2004. Sustaining the Earth: An Integrated Approach. Thomson/ Brooks/Cole, Belmont, CA, pp. 211–216, 1 January. Gupta, P., 2004. Pesticide exposure—Indian scene. Toxicology 198, 83–90. Hai, F.I., Modin, O., Yamamoto, K., Fukushi, K., Nakajima, F., Nghiem, L.D., 2012. Pesticide removal by a mixed culture of bacteria and white-rot fungi. J. Taiwan Inst. Chem. Eng. 43, 459–462. Hamby, D.M., 1996. Site remediation techniques supporting environmental restoration activities. Sci. Total Environ. 191, 203–224. Harcourt, R.L., Horne, I., Sutherland, T.D., Hammock, B.D., Russell, R.J., Oakeshott, J.G., 2002. Development of a simple and sensitive fluorimetric method for isolation of coumaphos-hydrolysing bacteria. Lett. Appl. Microbiol. 34, 263–268. Helbling, D.E., 2015. Bioremediation of pesticide-contaminated water resources: the challenge of low concentrations. Curr. Opin. Biotechnol. 33, 142–148. Herman, P.L., Behrens, M., Chakraborty, S., Chrastil, B.M., Barycki, J., Weeks, D.P., 2005. A three-component dicamba O-demethylase from Pseudomonas maltophilia, strain DI-6. J. Biol. Chem. 26, 24759–24767. Ibrahim, W.M., Karam, M.A., El-Shahat, R.M., Adway, A.A., 2014. Biodegradation and utilization of organophosphorus pesticide malathion by cyanobacteria. Biomed. Res. Int. 2014, 1–6. Jacob, G.S., Garbow, J.R., Hallas, L.E., et al., 1988. Metabolism of glyophosate in Pseudomonas sp. strain LBr. Appl. Environ. Microbiol. 54, 2953–2958. Kaiser, J., 2000. Endocrine disrupters: panel cautiously confirms low-dose effects. Science 290, 695–697. Kandpal, V., 2014. Biopesticides. Int. J. Environ. Res. Dev. 4, 191–196. Marican, A., Dura´n-Lara, E.F., 2018. A review on pesticide removal through different processes. Environ. Sci. Pollut. Res. 25, 2051–2064. Mohan, D., Gupta, V.K., Srivastava, S.K., Chander, S., 2001. Kinetics of mercury adsorption from wastewater using activated carbon derived from fertilizer waste. Colloids Surf. A Physicochem. Eng. Asp. 177, 169–181. Mulbry, W.W., Karns, J.S., Kearney, P.C., Nelson, J.O., McDaniel, C.S., Wild, J.R., 1986. Identification of a plasmid-borne parathion hydrolase gene from Flavobacterium sp. by southern hybridization with opd from Pseudomonas diminuta. Appl. Environ. Microbiol. 51, 926–930. Mulbry, W.W., Zhu, H., Nour, S.M., Topp, E., 2002. The triazine hydrolase gene trzN from Nocardioides sp. strain C190: cloning and construction of gene-specific primers. FEMS Microbiol. Lett. 206, 75–79.

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Nagata, Y., Prokop, Z., Sato, Y., Jerabek, P., Kumar, A., Ohtsubo, Y., Tsuda, M., Damborsky, J., 2005. Degradation of β-hexachlorcyclohexane by haloalkane dehalogenase LinB from Sphingomonas paucimobilis UT26. Appl. Environ. Microbiol. 71, 2183–2185. Nwankwegu, A.S., Onwosi, C.O., 2017. Bioremediation of gasoline contaminated agricultural soil by bioaugmentation. Environ. Technol. Innov. 7, 1–11. Prabha, R., Singh, D.P., Verma, M.K., 2017. Microbial interactions and perspectives for bioremediation of pesticides in the soils. In: Singh, D., Singh, H., Prabha, R. (Eds.), Plant-Microbe Interactions in Agro-Ecological Perspectives. Springer, Singapore, pp. 649–671. Raina, R., Hall, P., 2008. Comparison of gas chromatography-mass spectrometry and gas chromatography-tandem mass spectrometry with electron ionization and negative-ion chemical ionization for analyses of pesticides at trace levels in atmospheric samples. Anal. Chem. Insights 2008, 111–125. Rajendran, S., 2003. Environment and health aspects of pesticides use in Indian agriculture. Int. J. Recent Sci. Res. 4, 415–424. Rani, S., Sud, D., 2015. Effect of temperature on adsorption-desorption behaviour of triazophos in Indian soils. Plant Soil Environ. 61, 36–42. Serdar, C.M., Gibson, D.T., Munnecke, D.M., Lancaster, J.H., 1985. Enzymatic hydrolysis of organophosphates: cloning and expression of a parathion hydrolase gene from Pseudomonas diminuta. Biotechnology 3, 367–371. Sharma, P., Raina, V., Kumari, R., Shweta, M., Dogra, C., Kumari, H., Kohler, H.P.E., Holliger, C., Lal, R., 2006. Haloalkane dehalogenase LinB is responsible for βand δ-hexachlorocyclohexane transformation in Sphingobium indicum B90A. Appl. Environ. Microbiol. 72, 5720–5727. Singh, B.K., Walker, A., 2006. Microbial degradation of organophosphorus compounds. FEMS Microbiol. Rev. 30, 428–471. Sinha, R.K., Valani, D., Sinha, S., et al., 2011. Bioremediation of contaminated sites: a low-cost nature’s biotechnology for environmental clean up by versatile microbes, plants and earthworms. In: Solid Waste Management and Environmental Remediation. Nova Science Publishers, Inc., UK, pp. 1–72. Streber, W.R., Timmis, K.N., Zenk, M.H., 1987. Analysis, cloning, and high-level expression of 2,4-dichlorophenoxyacetate monooxygenase gene tfdA of Alcaligenes eutrophus JMP134. J. Bacteriol. 169, 2950–2955. Sutherland, T.D., Horne, I., Harcourt, R.L., Russell, R.J., Oakeshott, J.G., 2002. Isolation and characterization of a Mycobacterium strain that metabolizes the insecticide endosulfan. J. Appl. Microbiol. 93, 380–389. US EPA, 2007. Treatment Technologies for Mercury in Soil, Waste, and Water. Office of Superfund Remediation and Technology Innovation, Washington, DC, p. 20460. Van Eerd, L.L., Hoagland, R.E., Zablotowicz, R.M., Hall, J.C., 2003. Pesticide metabolism in plants and microorganisms. Weed Sci. 51 (4), 472–495. Wang, J., Chen, C., 2009. Biosorbents for heavy metals removal and their future. Biotechnol. Adv. 27, 195–226. Weir, K.M., Sutherland, T.D., Horne, I., Russell, R.J., Oakeshott, J.G., 2006. A single monooxygenase, ese, is involved in the metabolism of the organochlorides endosulfan and endosulfate in an Arthrobacter sp. Appl. Environ. Microbiol. 72, 3524–3530. Yeh, W.K., Gibson, D.T., Liu, T.N., 1977. Toluene dioxygenase: a multicomponent enzyme system. Biochem. Biophys. Res. Commun. 78, 401–410.

Chapter 3 Exploring microbes as bioremediation tool

Further reading Repetto, R., Baliga, S., 1996. Trends and patterns of pesticide use. In: Pesticides and the Immune System; Public Health Risks. World Resources Institute, Washington, DC, pp. 3–8.

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4 Solar photocatalysis and its application for emerging contaminant removal from wastewater V. Prashanth, Pyla Jayasree, Parth Rajput, Neelancherry Remya Indian Institute of Technology Bhubaneswar, Argul, Odisha, India

1

Introduction

With the evolution of new technologies with higher precision and sensitivity, many emerging contaminants (ECs) are detected in natural and wastewater bodies. These ECs such as endocrinedisrupting compounds (EDCs), pharmaceuticals, and personal care products (PPCPs) as well as pesticides are of major concern due to their long-term toxicity, high resistance to biodegradation, and ease of accumulation in living species, which could lead to potential threats to the environment and human health (Ahmed et al., 2011; Karmakar and Kulshrestha, 2009; Xie et al., 2019). Concern about the continuous rise of the concentrations of these compounds shows that conventional wastewater treatments are not adequate for the complete removal of many ECs (Ahmed et al., 2011). Most of these ECs are thus expected to remain in the environment for a long time, unless eliminated by an effective treatment process (Almomani et al., 2018). In recent years, advanced oxidation processes (AOPs) such as photocatalysis, photo-Fenton, and ozonation have been effectively employed to remove these contaminants. AOPs are based on the production of highly oxidative radicals that facilitate the rapid degradation and mineralization of a range of organic and inorganic pollutants from wastewater (Almomani et al., 2018). Among these AOPs, solar photocatalysis draws major attention in treating ECs. Owing to its enormous availability, sustainability, and cleanliness, the concept of harvesting solar energy for various applications is booming all over the world. On the basis of energy, Advanced Oxidation Processes for Effluent Treatment Plants. https://doi.org/10.1016/B978-0-12-821011-6.00004-9 # 2021 Elsevier Inc. All rights reserved.

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the solar spectrum constitutes about 3%–5% UV (λ < 400 nm) and about 47% visible light (400 > λ < 700 nm) (Bora and Mewada, 2017). Solar photocatalysis makes use of the abundant solar energy to treat wastewater containing biorecalcitrant compounds with low cost and ease of operation (Almomani et al., 2018). This chapter summarizes solar photocatalysis, the types of modifications done to make the catalyst excite in the visible light spectrum, and their application in the removal of ECs such as pharmaceuticals, pesticides, and endocrine-disrupting compounds.

2 Solar photocatalysis A photon means light and catalysis is the process of increasing the chemical reaction with the help of a catalyst, which remains unchanged during reaction. Photocatalysis is the process of acceleration of a chemical reaction with the help of a catalyst by the absorption of the light source. Solar photocatalysis utilizes photon energy from sunlight by the semiconductor catalyst for its activation. Upon irradiation with light (photon energy) greater than or equal to its bandgap energy, the semiconductor photocatalyst, electron (e
) from the valence band (VB), gets excited to the conduction band (CB), thus generating an electron (e
) and hole (h+) pair (EHP) (Eq. 1). These photogenerated EHPs would promote different redox reactions, facilitating pollutant degradation. The hole (h+) in the valence band produces hydroxyl radicals (%OH), whereas the electron (e
) in the conduction band reacts with the dissolved oxygen species (O2) and forms superoxide radicals (O2%
). These radicals facilitate the degradation of pollutants through oxidation or reduction, as shown in Fig. 1. Catalyst

hγ photocatalyst + hþ + e
ðphotonÞ

(1)

There are two ways in which photooxidation of the pollutant can take place: (i) Direct photocatalysis. (ii) Mediated photocatalysis.

2.1

Direct photocatalysis

In this process, the dissolved organic pollutants (DOP) present in the solution will be transferred from the liquid or gaseous phase to the solid phase and adsorbed on the surface of the catalyst.

Chapter 4 Solar photocatalysis for emerging contaminant removal

Fig. 1 Mechanism of solar photocatalysis.

Then it is directly oxidized by the hole (h+) formed in the valence band of the catalyst (Eq. 2). ðoxidationÞ

hþ + Dissolved organic pollutant ƒƒƒƒƒ! CO2 + H2 O

2.2

(2)

Mediated photocatalysis

In this process, the hole (h+) formed in the valence band of the catalyst initially reacts with the water molecule (H2O) to form hydroxyl radicals (%OH). Then, the %OH oxidizes the dissolved organic pollutant present in the solution (with/without any adsorption on the surface of the semiconductor catalyst) and mineralizes the pollutant to carbon dioxide (CO2) and water (H2O) (Eqs. 3, 4). h + + H2 O ! HO ðoxidationÞ

HO + Dissolved organic pollutant ƒƒƒƒƒ! CO2 + H2 O

3

(3) (4)

Modification of photocatalyst

The semiconductors used in the photocatalysis process include TiO2, ZnO, WO3, Cds, Fe2O3, Cu2O, CdSe, etc. Among these, TiO2 catalysts have been the most widely used because of their unique properties such as strong redox ability, low cost, nontoxicity, and chemical stability. However, they also have some drawbacks such as a higher bandgap energy of 3.2 eV that needs UV light excitation, which makes it costly. So, to excite TiO2 with

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visible light, the bandgap of TiO2 needs to be reduced. However, this bandgap reduction facilitates EHP recombination, leading to heat generation while the photocatalytic efficiency will also be reduced. On the other hand, TiO2 has a photocatalytic efficiency of only 10% (i.e., only 10% of the generated electrons are utilized for photocatalytic degradation) and the remaining 90% of the electrons are again recombined with the holes in picoseconds with the generation of heat, a process known as electronhole pair (EHP) recombination. If EHP recombination takes place on the surface of the catalyst, then it is known as surface recombination; if it happens in the bulk it is known as volume recombination. This EHP recombination reduces the photocatalytic efficiency of the catalyst. To reduce the bandgap energy for the excitation of the catalyst in the visible spectrum and to increase photocatalytic efficiency, modifications such as metal/nonmetal doping, coupling, structural modification, and noble metal deposition on the surface have to be done. The subsequent sections elaborate the TiO2 phototocatalyst modifications to improve the solar photocatalytic degradation efficiency.

4 Types of photocatalyst modification 4.1

Doping

Various doping techniques adopted include (i) doping of TiO2 with metal dopants, (ii) doping of TiO2 with nonmetal dopants, (iii) codoping TiO2 with different elements, and (iv) coupling TiO2 with a semiconductor having a narrow bandgap or a lower bandgap energy. The addition of metal or nonmetal dopants induces new doping (energy) levels inside the bandgap of TiO2. Hence, TiO2 can be excited by a relatively lower energy, that is, by light having higher wavelengths. Therefore, instead of UV light, doped TiO2 can be excited by visible light.

4.1.1

Metal doping

Metal doping means substituting suitable metal ions for Ti4+. Generally, metal dopants are added near the conduction band to form new energy levels below the conduction band (Fig. 2). In addition, metal doping helps to reduce the EHP recombination and to increase the availability of free electrons to take part in the degradation process. Different types of metal dopants used for modifying the properties of TiO2 include: (i) noble metals, (ii) rare earth metals, and (iii) transition metals.

Chapter 4 Solar photocatalysis for emerging contaminant removal

e–

CB

e–

e–

hn hn1

3.2 eV

h+

h+ VB

Fig. 2 Energy level of metal-doped TiO2.

4.1.2

Noble metal doping

Noble metals such as palladium (Pd), platinum (Pt), silver (Ag), ruthenium (Ru), and iridium (Ir) are used to reduce the bandgap energy of TiO2 and to excite/activate the catalyst by visible light. The deposition of noble metals on the surface of the TiO2 near the conduction band facilitates the absorption of the photogenerated electron (e
), thus reducing EHP recombination and making it available for pollutant degradation. Basically, the noble metals are known to form intermediators to trap and transfer the photogenerated electron (e
) to the electron (e
) acceptor.

4.1.3

Rare metal doping

Rare earth metals are comprised of 17 elements such as scandium (Sc), yttrium (Y), and 15 lanthanides. Doping TiO2 with rare earth metals increases the adsorption of organic pollutants on the catalyst surface, thereby increasing the photocatalytic efficiency of TiO2 (Shayegan et al., 2018). In addition, rare earth metal doping improves the light absorption capacity, the surface area phase structure, and the surface morphology of TiO2. Further, research indicates that doping with rare earth metals such as Ce-TiO2 enhances the thermal stability of TiO2 (Korologos et al., 2012).

4.1.4

Transition metal doping

The incorporation of transition metals such as manganese (Mn), iron (Fe), copper (Cu), vanadium (V), and nickel (Ni) helps in reducing the bandgap and also reducing the EHP recombination. The main factors affecting photocatalytic oxidation are the type of transition metal and the amount of transition metal used. The dopant will act as a charge carrier bridge and thus increase

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the separation of e
and h+. But a higher amount of dopant than the optimum value will reduce the crystallinity of the catalyst. Therefore, the dopant will act as an EHP recombination center, thereby reducing the photocatalytic efficiency. Therefore, finding the optimal type and concentration of dopant is of paramount importance in this process. 4.1.4.1 Iron The incorporation of Fe with TiO2 reduces the EHP recombination and also narrows the bandgap. Hence, with visible light TiO2 can be excited. On the other hand, iron is cheaper compared to noble metals such as Pd, Pt, Ag, and Au, which makes the Fe more attractive as a metal dopant (Yang et al., 2015). Moreover, iron will act as an electron acceptor, get reduced from Fe3+ to Fe2+, and increase the separation of EHP (Fig. 3). The photoexcited electron (e
) will be taken up and transferred by iron to the O2 molecule, and superoxide radicals will be formed. Fe3+ has a lesser radius ˚ ) than that of Ti4+ (0.68 A ˚ ), which makes the incorporation (0.64 A 3+ of Fe into the TiO2 crystal structure easier (Christoforidis et al., 2012; Hung et al., 2008). The electron-hole trap enabled by iron in the iron-doped TiO2 is illustrated in the following equations (Eqs. 5–7) (Dong et al., 2010; Yang et al., 2015). Fe3 + + e
! Fe2 +

(5)

Fe2 + + O2 ! Fe3 + + O
2

(6)

Fe3 + + h + ! Fe4 +

(7)

e

–

CB

e–

e

Fe3+

–

e– ----------------------------------------------hn

Fe2+

hn1

3.2 eV

h+

h+ VB

Fe3+ Fe4+

Fig. 3 Energy level of iron-doped TiO2.

Chapter 4 Solar photocatalysis for emerging contaminant removal

4.1.4.2

Nickel

Like iron, the incorporation of nickel will also reduce the EHP recombination by acting as an electron acceptor. The efficient separation of EHP and the photocatalytic efficiency of Ni-TiO2 are dependent on the dopant concentration. In low concentrations of dopant, Ni2+-doped TiO2 will separate the charge carriers by trapping e
and enabling h+ to take part in the oxidation of pollutants. However, in a high dopant concentration, Ni-TiO2 becomes the recombination center for EHP. The charge separation in Ni-TiO2 follows different routes, as shown in (Eqs. 8–11). Ni+ could transfer electrons to the adsorbed O2 on the surface of TiO2 (Eq. 9) and react with Ti4+ (Eq. 10) to create an interfacial electron transfer (Tseng et al., 2009).

4.1.5

Ni2 + + e
! Ni +

(8)

Ni + + O2ðadsÞ ! Ni2 + + O
2

(9)

Ni + Ti4 + ! Ni2 + + Ti3 +

(10)

Ti3 + + O2ðadsÞ ! Ti4 + + O
2

(11)

Nonmetal doping

Apart from the metal dopant, the bandgap of TiO2 can be reduced by incorporating anionic nonmetal dopants such as nitrogen (N), carbon (C), sulfur (S), boron (B), and fluorine (F). Generally, the nonmetal dopants are incorporated near the valence band, but they don’t take part in the separation of the charge carrier (Fig. 4). Studies have shown that nonmetal anionic dopants are more effective in expanding the TiO2 light absorption range into the visible light region than cationic metal dopants. It has been seen that N- and C-doped TiO2 nanomaterials have superior photocatalytic activity under visible light excitation than other anionic nonmetal dopants (Chen et al., 2007). 4.1.5.1

N-doped TiO2

Among all the nonmetal dopants, nitrogen is most widely used for doping. In addition to the reduction in bandgap and visible light excitation of the modified TiO2, N can easily enter the lattice of the TiO2 due to the smaller radius of N compared to O2 (Pandian et al., 2016). Researchers have seen that TiO2 nanoparticles incorporating with any anion having lesser electronegativity than O2 results in a shifting of 2P orbitals of the valence band upward, which further results in a reduction in the bandgap energy of

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e–

CB

e–

hn

hn1

3.2 eV

h

+

h

h+

+

VB

Fig. 4 Energy levels of nonmetal-doped TiO2.

TiO2 (Fig. 5A) and makes TiO2 excitation possible with visible light (Albrbar et al., 2016). The problem with N-doped TiO2 is that if the concentration of dopant (N) is low, then there will be a considerable shift in the energy in the valence band. This indicates that appropriate techniques should be followed for N-TiO2 preparation for extension of the light absorption range and improvement of the photocatalytic activity (Daghrir et al., 2013). 4.1.5.2

C-doped TiO2

Doping with carbon improves the stability and conductivity of TiO2 and also enhances the removal of organic pollutants by adsorption on the catalyst surface. Further, carbon doping improves the visible light absorption of TiO2. Inorganic carbon or residual carbon containing organic species can also be used CB e

–

CB e–

–

e

CB e– e– e–

hn1

hn 3.2 eV

h

+

h

hn

hn2

+

h+

hn 3.2 eV

h+

VB

VB

(A)

hn1

3.2 eV

(B)

(C)

Fig. 5 Energy level of (A) N-doped TiO2, (B) C-doped TiO2, and (C) S-doped TiO2.

hn1

hn2 h+

VB

Chapter 4 Solar photocatalysis for emerging contaminant removal

for doping. Doping with C causes a shifting of the midband up, and enhances the oxidative potential for photogenerated holes (Fig. 5B) (Kavitha and Devi, 2014). 4.1.5.3 S-doped TiO2 Doping with sulfur is also extensively explored. Incorporation with sulfur results in an increase in the valence band width and also decreases the bandgap. Sulfur doping can be achieved in two ways: incorporating sulfur into the TiO2 lattice as (i) anion S2
or (ii) cation (S4+/S6+) (Fig. 5C). As an anion, dopant sulfur will replace O2– in the TiO2 lattice and as a cation, sulfur will replace ˚ ) than that of Ti4+ Ti4+. The S6+ ion has a lesser atomic radius (0.29 A 6+ 4+ ˚ (0.64 A) and hence S can easily replace Ti . However, the atomic ˚ ) is higher than that of O2– (1.22 A ˚ ). Thus, the radius of S2
(1.7 A energy requirement to form a bond between TidO is lower than that of TidS. Therefore, replacing Ti4+ with S6+ is easier compared to replacement of O2– with S2
(Ma et al., 2014).

4.2

Co-doping

Codoping means instead of doping TiO2 with a single material, it is doped with two or more materials. It eliminates the limitations of the single-doped TiO2. Codoping can be done in the following ways: (i) metal codoping, (ii) nonmetal codoping, and (iii) metal and nonmetal codoping.

4.2.1

Metal codoping

In metal codoping, two or more metals are doped with photocatalysts such as TiO2. For example, silver/vanadium (Ag/V) codoped TiO2 deposited on polyurethane showed significant photodegradation of hexane and butyl acetate (Pham and Lee, 2017). It is difficult to replace Ti4+ with Ag+ in single doping as ˚ ) is less than that of Ag+ the atomic radius of the former (0.74 A 4+ ˚ (1.26 A). On the other hand, Ti can be easily replaced by V4+ hav˚ , thus resulting in the formation of a ing an atomic radius of 0.72 A 4+ TidOdV bond. Also, V and Ag+ together as dopants can easily enter the lattice of TiO2, resulting in the conversion of Ti4+ to Ti3+. Codoping exhibited a higher ratio of Ti3+/Ti4+ than of single Ag- or V-doped TiO2. This reflects the increased concentration of dopants in the codoped TiO2 as compared to that in the single Agor V-doped TiO2. The formation of Ti3+ indicates the vacancy of oxygen, which means that codoping with Ag or V increases the oxygen vacancies inside the TiO2 lattice. Also, Ag/V codoping into TiO2 causes the uniform distribution of Ag, Ag2O, and V2O5 on the

77

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Chapter 4 Solar photocatalysis for emerging contaminant removal

CB

e–

hn1 h+

e–

e–

hn hn2

3.2 eV

h+

h+ VB

Fig. 6 Energy level of metal and nonmetal doped TiO2.

TiO2 surface and thus reduces their aggregation (Nguyen et al., 2019a).

4.2.2

Metal and nonmetal codoping

Codoping with metal and nonmetal dopants results in the formation of new energy levels within the bandgap of TiO2 and thus causes an increase in the visible light absorption capacity. The replacement of Ti4+ inside the TiO2 by a metal dopant creates a new doping level just below the conduction band (Fig. 6). Conversely, the nonmetal dopant will create a new doping level just above the valence band (Fig. 6). Also, the metal and nonmetal dopants increase EHP separation, thereby increasing the photocatalytic efficiency of TiO2 under visible light. Previous studies showed improvement in the photocatalytic activity of N- or S-doped TiO2 when it is codoped with Fe3+. The photodegradation efficiency of gaseous toluene was observed as 97% with Fe/N codoped TiO2 (Dong et al., 2010). Moreover, the suitability of N-doped TiO2 will be increased due to the addition of Fe3+ because of the surface modification of the catalyst.

4.3

Coupling of semiconductors

Coupling is one of the techniques used to reduce the bandgap energy of a semiconductor by coupling it with another semiconductor. A TiO2 semiconductor is coupled with another semiconductor such as WO3, SiO2, SnO2, MnCO3, CdS, Ag3VO4, or ZnO and a porous material with a large surface area (e.g., activated carbon). The coupling improves the visible light photodegradation of pollutants compared to pure TiO2. For example, the visible light excitation of CdS-coupled TiO2 proceeds as follows: (i) visible light

Chapter 4 Solar photocatalysis for emerging contaminant removal

CB

e–

CB

CdS TiO 2

VB

e–

hn 2.4 eV

hn 3.2 eV

VB

h+

h+

Fig. 7 Energy level of TiO2 coupled with CdS.

excitation of the e
from the valence band (VB) to the conduction band (CB) of CdS (bandgap energy 2.25 eV), (ii) transfer of e
from the CB of CdS to the CB of TiO2 (Fig. 7), (iii) visible light excitation of e
from the VB of TiO2 to the VB of CdS by the energy of visible light, and (iv) the formation of a hole in the VB of TiO2. As a result, EHP is formed in the TiO2 semiconductor with visible light excitation (Eqs. 12, 13) (Liu et al., 2012). One of the key points to be remembered in coupling is that the valence band and conduction band of the coupling semiconductor should not lie in between the valence band and conduction band of the semiconductor to be coupled. Visible light

5

CdS
TiO2 ƒƒƒƒƒ! CdS ðe
+ h + Þ
TiO2

(12)

CdS ðe
+ h + Þ
TiO2 ! CdS ðh + Þ
TiO2 ðe
Þ

(13)

Application of solar photocatalysis for the removal of emerging contaminants

Emerging contaminants (ECs) include pharmaceuticals, personal care products (PCPs), pesticides, industrial products, and many other types of anthropogenic chemicals. ECs occur in natural and built environments at low concentrations (a few ng/L to tens of μg/L), thus they are also known as micropollutants. As current wastewater treatment plants are not designed for EC removal, these pollutants are released into aquatic environments. Pharmaceuticals such as sulfamethoxazole, acetaminophen, metformin,

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and diclofenac substances were found in different fish species. Pharmaceutical active compounds (PhACs) are metabolized and adsorbed by organisms at low levels. This results in exposure to the residues of these compounds such as their original molecular form or their transformed product when they enter aquatic life. Although the environmental impacts of ECs are not fully understood, several studies have indicated their deleterious effects and chronic damage to the ecosystem. Therefore, advanced treatment methods for the removal of ECs from water and wastewater have gained attention these days. Several studies have employed solar photocatalysis for EC removal from aqueous systems. Table 1 summarizes various modifications of the semiconductor photocatalyst demonstrated with enhanced degradation efficiencies of pharmaceuticals, one of the major ECs of concern. Acetaminophen is one of the most widely used analgesics and antipyretic drugs throughout the world. It transforms to N-acetyl-p-benzoquinone-imine upon oxidation, which can further be hydrolyzed to 1,4-benzoquinone, also a toxicant of major concern. Concentrations of acetaminophen varying from 3 to 20 mg/L were investigated for solar photocatalytic degradation using different modified catalysts (Table 1). The results indicated 100% degradation efficiency at lower concentrations within 1 h of s et al., 2019; Nasr et al., 2019; Shaban and treatment (Go´mez-Avile Fallata, 2019; Tobajas et al., 2017). Bisphenol-A (BPA) is a starting material for the synthesis of plastics. In 2015, an estimated 4 million tonnes of BPA chemicals were produced globally for manufacturing polycarbonate plastic, making it one of the highest volumes of chemicals produced worldwide. BPA is also classified as one of the EDCs. Codoped TiO2 catalysts such as ZnFe2O4-TiO2 and Co/N TiO2 indicated >97% degradation efficiency under optimum conditions (Garg et al., 2019; Nguyen et al., 2019a). Complete removal of various pharmaceuticals such as carbamazepine, sulfamethazine, ibuprofen, propranolol, and diclofenac as well as pesticides such as dimethoate and malathion was observed in solar photocatalytic degradation studies using various doped/codoped catalysts (Bo et al., 2019; He et al., 2016; Khedr et al., 2019; Vela et al., 2019; Yu et al., 2019). In addition, significant removal of ECs such as ofloxacin, diazinon, 1,4-dioxane, ciprofloxacin, and amoxicillin was reported using the solar photocatalytic treatment with various catalysts with/without modification (Nguyen et al., 2019b; Pattnaik et al., 2019; Phuong et al., 2019; Shabani et al., 2019; Xu et al., 2019).

Table 1 Application of solar photocatalysis for the removal of emerging contaminants.

Catalyst

Bandgap (eV)

Catalyst dosage (g/L); treatment time (h); pH

Removal efficiency (%)

Glycine (NH2CH2COOH)

C-TiO2

1.78

2; 1.5; 7

100

Acetaminophen (5 mg/L)

Lignin

C-TiO2

2.95

0.25; 1

100

Acetaminophen (10 mg/L) Acetaminophen (20 mg/L) Bisphenol-A (10 mg/L)

Zinc acetate dihydrate, clay HPtCl6

TiO2-0.5ZnO/ clay Pt/TiO2

3.2

0.5; 6 h

>90

NA

0.4; 2; 5.6–6.3

73.65

Zn(NO3)2  6H2O, Fe(NO3)3 9H2O

ZnFe2O4-TiO2

2.56

1; 0.5; 9

98.70

Cobalt (II) nitrate hexahydrate [Co (NO3)2.6H2O], urea Indium nitrate, thioacetamide, zinc nitrate Cu(NO3)2

Co/N TiO2

2.88

0.14; 2.5; 3

97

ZnIn2S4/TiO2

1.9

0.075; 4

100

Bo et al. (2019)

TiO2/CuxO

2.2

0.5; 4.2; 3

100

BiOClBi24O31Cl10/ rGO ZnO/Na2S2O8

2.48

1; 2

84.80

Yu et al. (2019) Shabani et al. (2019)

3.3

0.2; 4.2; 7.2

100

Pollutant (concentration)

Dopant

1

Acetaminophen (3 mg/L)

2

3

S. No.

4 5

6

Bisphenol-A (30 mg/L)

7

Carbamazepine (100 mg/L)

8

Sulfamethazine (2.5 mg/L) Ofloxacin (25 mg/L)

9

10

Dimethoate, malathion (0.3 mg/L)

Reduced graphene oxide (rGO) Na2S2O8

Reference Shaban and Fallata (2019) Go´mezAviles et al. (2019) Tobajas et al. (2017) Nasr et al. (2019) Nguyen et al. (2019a) Garg et al. (2019)

Vela et al. (2019) Continued

Table 1 Application of solar photocatalysis for the removal of emerging contaminants—cont’d

S. No.

Pollutant (concentration)

11

Diazinon (25 mg/L)

12

Ibuprofena (20 mg/L)

13

Propranolol, Diclofenaca (5 mg/L) 1,4-dioxanea (50 mg/L) Ciprofloxacina (20 mg/L) Amoxicillina (0.365 mg/L)

14 15 16

a

No doping/co-doping.

Bandgap (eV)

Catalyst dosage (g/L); treatment time (h); pH

Removal efficiency (%)

Dopant

Catalyst

Fe(NO3)39H2O, bentonite clay (Bent) Glycine/NaOH

Fe-TiO2/BentFe

3.17

0.5; 6; 4.5

80.60

Phuong et al. (2019)

Heterojunction TiO2 TiO2

3.17

0.5; 5

100

3.2

25; 48 h

100

WO3/ng-Al2O3

2.8

0.7; 4

85

g-C3N4

2.8

1; 1; 6.5

78

WO3

2.8

0.3; 3; 4

99.39

Khedr et al. (2019) He et al. (2016) Xu et al. (2019) Pattnaik et al. (2019) Nguyen et al. (2019b)

Quartz sand as a support g-Al2O3 as a support NA NA

Reference

Chapter 4 Solar photocatalysis for emerging contaminant removal

References Ahmed, S., Rasul, M.G., Brown, R., Hashib, M.A., 2011. Influence of parameters on the heterogeneous photocatalytic degradation of pesticides and phenolic contaminants in wastewater: a short review. J. Environ. Manag. https://doi.org/ 10.1016/j.jenvman.2010.08.028.  c, J., Mitric, M., Janackovic, D., Albrbar, A.J., Djokic, V., Bjelajac, A., Kovac, J., Cirkovi Petrovi c, R., 2016. Visible-light active mesoporous, nanocrystalline N,S-doped and co-doped titania photocatalysts synthesized by non-hydrolytic sol-gel route. Ceram. Int. https://doi.org/10.1016/j.ceramint.2016.07.144. Almomani, F., Bhosale, R., Kumar, A., Khraisheh, M., 2018. Potential use of solar photocatalytic oxidation in removing emerging pharmaceuticals from wastewater: a pilot plant study. Sol. Energy. https://doi.org/10.1016/ j.solener.2018.07.041. Bo, L., Liu, H., Han, H., 2019. Photocatalytic degradation of trace carbamazepine in river water under solar irradiation. J. Environ. Manag. https://doi.org/10.1016/ j.jenvman.2019.03.132. Bora, L.V., Mewada, R.K., 2017. Visible/solar light active photocatalysts for organic effluent treatment: fundamentals, mechanisms and parametric review. Renew. Sust. Energ. Rev. https://doi.org/10.1016/j.rser.2017.01.130. Chen, D., Jiang, Z., Geng, J., Wang, Q., Yang, D., 2007. Carbon and nitrogen co-doped TiO2 with enhanced visible-light photocatalytic activity. Ind. Eng. Chem. Res. https://doi.org/10.1021/ie061491k. Christoforidis, K.C., Figueroa, S.J.A., Ferna´ndez-Garcı´a, M., 2012. Iron-sulfur codoped TiO2 anatase nano-materials: UV and sunlight activity for toluene degradation. Appl. Catal. B Environ. https://doi.org/10.1016/j.apcatb. 2012.01.029. Daghrir, R., Drogui, P., Robert, D., 2013. Modified TiO2 for environmental photocatalytic applications: a review. Ind. Eng. Chem. Res. https://doi.org/10.1021/ ie303468t. Dong, F., Wang, H., Wu, Z., Qiu, J., 2010. Marked enhancement of photocatalytic activity and photochemical stability of N-doped TiO2 nanocrystals by Fe3+/ Fe2+ surface modification. J. Colloid Interface Sci. https://doi.org/10.1016/ j.jcis.2009.11.012. Garg, A., Singhania, T., Singh, A., Sharma, S., Rani, S., Neogy, A., Yadav, S.R., Sangal, V.K., Garg, N., 2019. Photocatalytic degradation of Bisphenol-A using N, Co codoped TiO2 catalyst under solar light. Sci. Rep. https://doi.org/ 10.1038/s41598-018-38358-w. s, A., Pen ˜ as-Garzo´n, M., Bedia, J., Rodriguez, J.J., Belver, C., 2019. CGo´mez-Avile modified TiO2 using lignin as carbon precursor for the solar photocatalytic degradation of acetaminophen. Chem. Eng. J. https://doi.org/10.1016/ j.cej.2018.10.154. He, Y., Sutton, N.B., Rijnaarts, H.H.H., Langenhoff, A.A.M., 2016. Degradation of pharmaceuticals in wastewater using immobilized TiO2 photocatalysis under simulated solar irradiation. Appl. Catal. B Environ. https://doi.org/10.1016/ j.apcatb.2015.09.015. Hung, W.C., Chen, Y.C., Chu, H., Tseng, T.K., 2008. Synthesis and characterization of TiO2 and Fe/TiO2 nanoparticles and their performance for photocatalytic degradation of 1,2-dichloroethane. Appl. Surf. Sci. https://doi.org/10.1016/ j.apsusc.2008.07.079. Karmakar, R., Kulshrestha, G., 2009. Persistence, metabolism and safety evaluation of thiamethoxam in tomato crop. Pest Manag. Sci. https://doi.org/10.1002/ ps.1776.

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Kavitha, R., Devi, L.G., 2014. Synergistic effect between carbon dopant in titania lattice and surface carbonaceous species for enhancing the visible light photocatalysis. J. Environ. Chem. Eng. https://doi.org/10.1016/j.jece.2014.02.016. Khedr, T.M., El-Sheikh, S.M., Ismail, A.A., Bahnemann, D.W., 2019. Highly efficient solar light-assisted TiO2 nanocrystalline for photodegradation of ibuprofen drug. Opt. Mater. (Amst.). https://doi.org/10.1016/j.optmat.2018.11.027. Korologos, C.A., Nikolaki, M.D., Zerva, C.N., Philippopoulos, C.J., Poulopoulos, S.G., 2012. Photocatalytic oxidation of benzene, toluene, ethylbenzene and m-xylene in the gas-phase over TiO2-based catalysts. J. Photochem. Photobiol. A Chem. https://doi.org/10.1016/j.jphotochem. 2012.06.016. Liu, Z., Fang, P., Wang, S., Gao, Y., Chen, F., Zheng, F., Liu, Y., Dai, Y., 2012. Photocatalytic degradation of gaseous benzene with CdS-sensitized TiO2 film coated on fiberglass cloth. J. Mol. Catal. A Chem. https://doi.org/10.1016/j.molcata.2012.06.004. Ma, D., Xin, Y., Gao, M., Wu, J., 2014. Fabrication and photocatalytic properties of cationic and anionic S-doped TiO2 nanofibers by electrospinning. Appl. Catal. B Environ. https://doi.org/10.1016/j.apcatb.2013.08.004. Nasr, O., Mohamed, O., Al-Shirbini, A.S., Abdel-Wahab, A.M., 2019. Photocatalytic degradation of acetaminophen over Ag, Au and Pt loaded TiO2 using solar light. J. Photochem. Photobiol. A Chem. https://doi.org/ 10.1016/j.jphotochem.2019.01.032. Nguyen, T.B., Huang, C.P., Doong, R., 2019a. Photocatalytic degradation of bisphenol A over a ZnFe2O4/TiO2 nanocomposite under visible light. Sci. Total Environ. https://doi.org/10.1016/j.scitotenv.2018.07.352. Nguyen, T.T., Nam, S.N., Son, J., Oh, J., 2019b. Tungsten trioxide (WO3)-assisted photocatalytic degradation of amoxicillin by simulated solar irradiation. Sci. Rep. https://doi.org/10.1038/s41598-019-45644-8. Pandian, R., Natarajan, G., Dhaipule, N.G.K., Prasad, A.K., Kamruddin, M., Tyagi, A.K., 2016. Types of nitrogen incorporation in reactively sputtered titania thin films: influence on UV–visible, photocatalytic and photoconduction properties. Thin Solid Films. https://doi.org/10.1016/j.tsf.2016.08.067. Pattnaik, S.P., Behera, A., Martha, S., Acharya, R., Parida, K., 2019. Facile synthesis of exfoliated graphitic carbon nitride for photocatalytic degradation of ciprofloxacin under solar irradiation. J. Mater. Sci. https://doi.org/10.1007/s10853-01803266-x. Pham, T.D., Lee, B.K., 2017. Selective removal of polar VOCs by novel photocatalytic activity of metals co-doped TiO2/PU under visible light. Chem. Eng. J. https:// doi.org/10.1016/j.cej.2016.08.068. Phuong, N.M., Chu, N.C., Van Thuan, D., Ha, M.N., Hanh, N.T., Viet, H.D.T., Minh Thu, N.T., Van Quan, P., Thanh Truc, N.T., Sharma, A.K., 2019. Novel removal of diazinon pesticide by adsorption and photocatalytic degradation of visible lightdriven Fe-TiO2/bent-Fe photocatalyst. J. Chem. https://doi.org/10.1155/ 2019/2678927. Shaban, Y.A., Fallata, H.M., 2019. Sunlight-induced photocatalytic degradation of acetaminophen over efficient carbon doped TiO2 (CTiO2) nanoparticles. Res. Chem. Intermed. https://doi.org/10.1007/s11164-019-03750-2. Shabani, M., Haghighi, M., Kahforoushan, D., Haghighi, A., 2019. Sonosolvothermal hybrid fabrication of BiOCl-Bi24O31Cl10/rGO nano-heterostructure photocatalyst with efficient solar-light-driven performance in degradation of fluoroquinolone antibiotics. Sol. Energy Mater. Sol. Cells. https://doi.org/10.1016/j.solmat.2019.01.034.

Chapter 4 Solar photocatalysis for emerging contaminant removal

Shayegan, Z., Lee, C.S., Haghighat, F., 2018. TiO2 photocatalyst for removal of volatile organic compounds in gas phase—a review. Chem. Eng. J. https://doi.org/ 10.1016/j.cej.2017.09.153. Tobajas, M., Belver, C., Rodriguez, J.J., 2017. Degradation of emerging pollutants in water under solar irradiation using novel TiO2-ZnO/clay nanoarchitectures. Chem. Eng. J. https://doi.org/10.1016/j.cej.2016.10.002. Tseng, H.H., Wei, M.C., Hsiung, S.F., Chiou, C.W., 2009. Degradation of xylene vapor over Ni-doped TiO2 photocatalysts prepared by polyol-mediated synthesis. Chem. Eng. J. https://doi.org/10.1016/j.cej.2008.12.015. rez-Lucas, G., ˜ ez-Gasco´n, M.J., el Aatik, A., Garrido, I., Pe Vela, N., Calı´n, M., Ya´n Fenoll, J., Navarro, S., 2019. Removal of pesticides with endocrine disruptor activity in wastewater effluent by solar heterogeneous photocatalysis using ZnO/Na2S2O8. Water Air Soil Pollut. https://doi.org/10.1007/s11270-0194185-y. Xie, H., Wang, X., Chen, J., Li, X., Jia, G., Zou, Y., Zhang, Y., Cui, Y., 2019. Occurrence, distribution and ecological risks of antibiotics and pesticides in coastal waters around Liaodong Peninsula, China. Sci. Total Environ. https://doi.org/10.1016/ j.scitotenv.2018.11.449. Xu, X., Liu, S., Cui, Y., Wang, X., Smith, K., Wang, Y., 2019. Solar-driven removal of 1,4-dioxane using WO3/nγ-Al2O3 nano-catalyst in water. Catalysts. https://doi. org/10.3390/catal9040389. Yang, S.B., Chun, H.H., Tayade, R.J., Jo, W.K., 2015. Iron-functionalized titanium dioxide on flexible glass fibers for photocatalysis of benzene, toluene, ethylbenzene, and o-xylene (BTEX) under visible- or ultraviolet-light irradiation. J. Air Waste Manag. Assoc. https://doi.org/10.1080/10962247.2014.995838. Yu, J., Kiwi, J., Wang, T., Pulgarin, C., Rtimi, S., 2019. Evidence for a dual mechanism in the TiO2/Cu x O photocatalyst during the degradation of sulfamethazine under solar or visible light: critical issues. J. Photochem. Photobiol. A Chem. https://doi.org/10.1016/j.jphotochem.2019.02.033.

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5 Advanced oxidative processes: An overview of their role in treating various wastewaters Komal Agrawal, Pradeep Verma Bioprocess and Bioenergy Laboratory, Department of Microbiology, Central University of Rajasthan, Ajmer, Rajasthan, India

1

Introduction

The present era has seen the depletion of natural resources and clean water as a result of globalization and industrialization. Artificially generated pollutants are contaminating water bodies, land, and air at a very fast pace. Thus, an immediate and effective step is required at the individual and governmental levels for the preservation and conservation of the environment. Human negligence in achieving a cleaner environment can be documented for many decades. The goal for economic development via industrialization is visible globally. However, the question to be raised is, “Is the development justified at the sake of the environment and our existence?” (Swaminathan et al., 2013). However, a step has to be taken that would balance the disturbance to attain a sustainable green environment. The water bodies contaminated by various pollutants and pathogens have to be treated to enable the water to be used for irrigation and various other purposes. In the past decade, the conventional purification techniques have seen a divergence toward newer, effective, and sustainable treatment methods called advanced oxidation processes (AOPs). The AOPs are cheap, sustainable, and effective treatment techniques. The AOPs treat the pollutants by generating hydroxyl radicals, which are strong oxidants that degrade or mineralize the pollutants nonselectively into harmless products (Swaminathan et al., 2013). Due to their broad range of action, AOPs can act upon both organic and inorganic pollutants present in the wastewater, along Advanced Oxidation Processes for Effluent Treatment Plants. https://doi.org/10.1016/B978-0-12-821011-6.00005-0 # 2021 Elsevier Inc. All rights reserved.

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with deactivating the pathogen and its indicators (Cho et al., 2005; Ikai et al., 2010). AOPs are usually not applied for disinfection as the radicals have too short a half-life (microseconds), as a result of which the time required for proper disinfection cannot be attained due to low radical concentration (Tchobanoglous et al., 2003). AOPs have been used for wastewater treatment as they are powerful oxidizing agents that destroy the pollutant and transform it into a nontoxic end product, thereby enabling its utilization (Huang et al., 1993). This chapter deals with various kinds of AOPs and their roles in wastewater treatment, along with their limitations and future prospects.

2 Types of AOPs AOPs have gained interest and various types of AOPs have been studied over time, along with the mechanisms of action of pollutants. Thus, various types of AOPs (Fig. 1) are discussed elaborately in the paper by Deng and Zhao (2015), and they are as follows:

2.1

Hydroxyl radical-based AOPs (HR-AOPs)

The HR-AOP, as compared to the saturated calomel electrode (the reference electrode), has an oxidation potential between 2.8 V (pH 0) and 1.95 V (pH 14) (Tchobanoglous et al., 2003). As mentioned in the previous section, due to the hydroxyl radicals (OH
), AOPs can act upon a broad range of pollutants. The OH


Fig. 1 Schematic representation of various types of AOPs.

Chapter 5 Advanced oxidative processes

radical has a rate constant on the order of 108  1010 M1 s1. The four mechanisms by which OH
radical can act upon the organic pollutant are: (1) radical addition, (2) hydrogen abstraction, (3) electron transfer, and (4) radical combination (Deng and Zhao, 2015). The HR-AOP acts on the organic compound and produces carbon-centered radicals (R
or R
dOH), which in the presence of oxygen can be transformed into organic peroxyl radicals (ROO
). The radicals then form hydrogen peroxide and super oxide, which act upon the pollutant for mineralization (Huang et al., 1993).

2.2

Ozone-based AOPs (Oz-AOPs)

Ozone as compared to the saturated calomel electrode has an oxidation potential of 2.07 V. The O3 oxidation is a selective reaction and has a rate constant of 1.0  100 to 103 M1 s18 as a result of which it reacts with organic compounds which are either ionized or dissociated. Under an indirect mechanism, ozone can also produce OH
that effectively acts upon the broad range of pollutants, which in the absence of OH
would not be possible (Gottschalk et al., 2009).

2.3

UV-based AOPs (UV-AOPs)

The generation of OH
radicals is very crucial to AOP implementation in wastewater treatment. They can be generated by photons in the presence of catalysts or oxidants such as titanium dioxide and RO-type semiconductors. When hydrogen peroxide and ozone are used as oxidants in the presence of UV, additional OH
can be produced. Similarly, by incorporation of the wavelength of 90%) within 24 h using Cu (II)/glucaric acid/H2O2 system. In the work done by Shah et al. (2003), various dyes (Remazol Brilliant Blue R (RBBR), Reactive Blue, Poly B-411, Chicago Sky Blue, Evans Blue, and Methyl Orange) were effectively degraded (>85%) by Cu (II), succinic acids, and an H2O2 system. On the other hand, using cobalt (II)/ascorbic acid/H2O2 > 90% decolorization was obtained with all the dyes except Remazol Brilliant Blue R (75%).

3

Compatible wastewater for AOP

Wastewater eliminated by industries have unique characteristic properties and the method of treatment varies from effluent to effluent. The common trend includes the substitution of biological treatment methods for physical chemical treatment methods; however, this is not always effective and thus other alternatives are required such as AOP. In certain cases, the removal of odorous compounds in water might be of prime importance as compared to the removal of total soluble organic. Thus, in such scenarios fast treatment is beneficial for the removal of odorous organics, which would remove the complaints associated with the odor prior to its release. The biologically treated pollutant can be treated with AOP to remove the nonbiodegradable compounds present in the polluted water effluents. The AOP can also be used for the removal of pollutants with high TOC/COD but less BOD (Feitz, 2005).

4

AOP in combination with other treatment technologies

Various treatment technologies such as physical, chemical, and biological treatments have been developed for the bioremediation of polluted wastewater. Thus, these treatments are either used singly or in combination for the effective bioremediation of the contaminated effluents. Thus AOPs, due to their easy

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integration capability in any process, can be combined with other treatments for the better removal of pollutants from wastewater (Augugliaro et al., 2006). AOP can be inserted into physiochemical and biological treatment steps and it has been stated (Oller et al., 2011) that whether AOP is the initial step or placed after the biological treatment, the cost would at the end be the same. However, when considering the combination of systems, the characteristic feature of each treatment has to be considered along with its limitation, for better understanding and implementation of the process at the treatment plants. It also has to be considered that when combinations of treatment technologies are applied, each step has to be individually optimized for the efficient implementation of the process. After the optimization of the entire treatment process, additional steps may have to be included for the better removal of pollutants (Malato et al., 2009).

5 Integration of AOP for efficient pollutant removal Factors such as the composition and load of pollutants influence the application of AOP in any treatment processes. Higher concentration of pollutants requires a harsher treatment method for pollutant removal. It also becomes necessary to integrate more than one treatment technique for pollutant removal. Thus, various combinations of treatment methods include various AOPs for the simultaneous removal of organic pollutants (e.g., UV/ ozone, UV/TiO2/H2O2, UV/H2O2, UV/Fenton’s reagent, ozone/ H2O2, etc.) via the generation of reactive species. The other methods would involve the sequential integration of AOPs for the effective removal of industrial wastes, which have different oxidizability to the different reactive species generated; an example would be a combination of Fenton oxidation followed by boron-doped diamond electrochemical treatment for the effective mineralization of olive mill effluents. Lastly, the solid phase of the pollutant should be separated via coagulation, sedimentation, etc., from the liquid phase for the efficient treatment of the pollutant by AOP (Comninellis et al., 2008). AOP (pretreatment) can be easily integrated with biological treatment (posttreatment) as well and has gained attention from researchers globally as a biological treatment is economically feasible and cost effective. If an AOP is used singly for treatment, it would generate huge costs for the complete removal of the

Chapter 5 Advanced oxidative processes

pollutant. Thus, the pollutants first treated by AOPs are broken down into short carboxylic acids, which are easily degraded by biological treatment methods (Khare et al., 2007)

6

Treatment of various wastewaters by AOP in combination with other treatment techniques

Industrial wastewater have unique characteristic properties from each industry that releases it along with varied concentrations, as a result of which the treatment process is unavoidably compromised at certain stages of treatment, particularly due to lack of information. However, Scott and Ollis (1995) identified four types of wastewater treatment: 1. Wastewater containing biorecalcitrant compounds (e.g., soluble polymers). 2. Industrial wastewater containing high amounts of biodegradable organic compounds and low concentrations of recalcitrant compounds. 3. Wastewater containing inhibiting compounds. 4. Wastewater containing inert intermediates (e.g., metabolites). Thus, over the years AOPs and in combination with other treatment processes have been effectively utilized for the treatment of various pollutants, which are discussed as follows:

6.1

Treatment of pesticides and/or herbicides

The pollution of water bodies by pesticides is a serious concern as they are highly soluble in water, which pollutes the groundwater due to easy propagation. Pesticides are carcinogenic and neurotoxic and have adverse effects on the reproductive system and cell development (Gray et al., 1994). The major sources of contamination occur from agricultural discharge along with nonagricultural activities. Thus, the removal of pesticides is of utmost importance in order to reutilize the water. The various conventional oxidation techniques used for pesticide removal include photochemical degradation (Kitsiou et al., 2009) and chemical oxidation processes (Benitez et al., 2002). On the other hand, the normal criteria involve biologically treating pesticides containing organic compounds. However, one major disadvantage includes the susceptibility of the microorganism to the toxic compounds, which reduces the effectiveness of the treatment process.

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Thus, an effective alternative would be to first treat the pollutant by oxidative treatment technologies such as AOPs. The AOPtreated waste then produces intermediates that are less toxic as compared to the parent compound and are more easily biodegraded by microorganisms without reducing the treatment efficiency (Oller et al., 2011).

6.2

Treatment of emerging pollutants and pharmaceutical wastes

As the name suggests, emerging pollutants are unregulated compounds for which future regulations need to be developed. Their impact on health and occurrence has to be studied for better understanding and development of efficient treatment technologies (Petrovi
c et al., 2003). Pharmaceutical waste is regarded as an emerging pollutant as it has endocrine-disrupting properties and can be released by both domestic and household waste (Kuster et al., 2005). Pharmaceutical waste and its concentrations vary depending upon its production, utility, and its end fate. Therefore, its treatment becomes mandatory prior to its discharge. The most common removal techniques for wastewater include adsorption, aerobic and anaerobic degradation, chemical degradation (Belgiorno et al., 2007), and granular activated carbon (GAC) adsorption. However, due to various limitations such as high cost, toxic end products have restricted their utility. Biological treatment technologies have also been applied, but the pollutants remain unaltered and are discharged from the treatment plants. Therefore, an alternative treatment, that is, AOPs, should be integrated for the effective removal of the pollutant. These alternatives include photocatalysis ozonation and ultrasound oxidation (Xekoukoulotakis et al., 2010; Dantas et al., 2008; Naddeo et al., 2010). AOPs can also be used in combination with other biological treatment processes for the removal of both emerging and pharmaceutical pollutants. Ozonation and preozonation followed by biological treatment of penicillin reduced the nonbiodegradable COD fraction from the effluent (Arslan-Alaton and Gurses, 2004). The other AOP that has been used for the treatment of pharmaceutical waste is solar photo-Fenton. Thus, it is observed that the AOP in combination with a biological system can be effectively used for pollutant removal, thereby enhancing its biodegradability (Oller et al., 2011).

Chapter 5 Advanced oxidative processes

6.3

95

Treatment of textile wastewaters

The textile industry utilizes tremendous quantities of water for dyeing purposes and then discharges it as effluent. These effluents consist of various types of dyes and chemicals, thereby polluting water bodies (Bisschops and Spanjers, 2003). Fully remediating these pollutants is a necessity. The dye industries also use anionic and ionic detergents prior to dyeing to remove the impurities from the material. The most commonly used azo dye is toxic and causes foam formation on the surface of water bodies, followed by eutophization. Physical, chemical, and biological wastewater treatment methods have been applied. However, AOPs exhibit more promising results, as in the prior treatment where various treatment techniques had to be used in combination to achieve the desired nontoxic end product (Table 1). The recent treatment trends involve the use of AOPs for the partial preoxidation of dyes. Various AOPs have been used in combination with aerobic

Table 1 The various AOPs used for the treatment of textile industrial wastewaters.

Sl. no.

AOP type

Type of effluent

1

Pulse light/H2O2

> 95

2

Fenton-like process (H2O2/Fe3+) UV/H2O2

Decolorization of azo dyes; Methyl orange Degradation of AY-17 Treatment of wastewaters containing Reactive Green 19

63

Removal of dye Reactive Red 198 from textile industry wastewater

100

Degradation of two textile dyes, Acid Blue 113 (AB 113) and Acid Red 88 (AR 88) Dye-contaminated river water and industrial wastewater Methylene Blue

AB113—98.7 and AR—8899.6 99

Treatment of 3,30 dimethoxybenzidine in sludge

98.24

3

4.

5.

6.

7.

Photo-Fenton activated carbon/CoFe2O4 nanocomposite UV-C/TiO2 suspension system MnO2-coated Fe3O4 nanocomposite (Fe3O4-MnO2) and adsorptive bubble separation Combined ultrasound-Mn (VII) treatment

Percentage removal (%)

83

References Martı´nez-Lo´pez et al. (2019) Khan et al. (2019) Zuorro and Lavecchia (2014) Heidari et al. (2019) Mortazavian et al. (2019) Kang et al. (2019)

Liang et al. (2019) Continued

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Table 1 The various AOPs used for the treatment of textile industrial wastewaters—cont’d

Sl. no.

AOP type

Type of effluent

8.

Photo-Fenton reagent based on hematite nanocrystals

9.

O3 =S2 O8 2

Textile wastewater streams loaded with Methylene Blue dye Removal of Reactive Red 120

10.

Combined biological-AOP treatment

11.

Sequential photocatalytic and biological treatment

Treatment of textile wastewater dye Reactive Red 180 (RR-180), Reactive Black 5 (RB 5), Remazol Red (RR) Removal of model textile dye, Methyl Red from wastewater

Percentage removal (%) 70 100 100

70

References Mansour et al. (2019) Azner et al. (2019) Thanavel et al. (2019) Waghmode et al. (2019)

biological remediation of three different types of wastewater from textile and paper industries as well as household detergents. It was observed that the OH
radicals produced without UV irradiation exhibited better results for all types of wastewater used. It also has to be noted that the preoxidation step did not enhance the € rsch et al., 2003). The AOPs (H2O2/ biodegradation of dye (Ho UV, TiO2/UV, and photo-Fenton) have also been used in combination with biofilm reactors for the treatment (99% removal) of azo dyes (Garcı´a-Montan˜o et al., 2008).

6.4

Treatment of paper industry wastewaters

Paper industry utilizes huge quantities of freshwater and simultaneously releases tremendous quantities of wastewater. The recycling of water allows the removal of fibers from paper industries. However, effluent released in the environment causes changes in the properties of water, enhancing slime growth and also disturbing the local environmental flora and fauna harshly affecting the ecosystem. The major stages where large quantities of water are utilized include wood preparation, pulp washing, screening, washing followed by bleaching, paper machine, and paper coating. The paper industries release chlorine, resins, juvaniones, unsaturated fatty acids, etc., during the process. These effluents released in the environment increase the biochemical oxygen demand, chemical oxygen demand, suspended solids, toxicity, and color of the natural water after its release from the

Chapter 5 Advanced oxidative processes

97

industries (Pokhrel and Viraraghavan, 2004). The various treatment techniques that have been used are sedimentation (removes suspended solids), coagulation (removes turbidity and color), and adsorption (removes color, COD, and AOX). The AOPs such as ozonation, Fenton’s reagent, adsorption, and membrane technology are effective, but they are expensive (de Pinho et al., 2000). Though various treatments have been applied, the desired COD level of