Wastewater Management and Technologies (Water and Wastewater Management) 3031362977, 9783031362972

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Water and Wastewater Management Global Problems and Measures

Eyüp Debik · Müfit Bahadir · Andreas Haarstrick   Editors

Wastewater Management and Technologies

Water and Wastewater Management Global Problems and Measures

Series Editors Müfit Bahadir, Institut für Ökologische und Nachhaltige Chemie, Technische Universität Braunschweig, Braunschweig, Germany Andreas Haarstrick, Leichtweiss-Institut für Wasserbau, Exceed, Technische Universität Braunschweig, Braunschweig, Germany

Water and wastewater management are among the greatest challenges of our century and the challenges posed by climate change will become even greater. Unfortunately, however, most efforts, especially in developing countries but also in the so-called developed countries, have been less than optimal or not optimal at all. In particular, there are still too many people who have to live without clean water and decent sanitation. Today, 2.2 billion people lack access to safely managed drinking water and wastewater, and 4.2 billion people lack safely managed sanitation services. The question, why this is the case - especially in developing countries - as well as other urgent water and wastewater management issues, are discussed in this book-series. Contributions therein present in more detail critical reviews, discussions, and analysis of the water and wastewater situation and management aspects in different regions and countries worldwide.

Eyüp Debik • Müfit Bahadir Andreas Haarstrick

•

Editors

Wastewater Management and Technologies

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Editors Eyüp Debik Environmental Engineering Davutpasa Campus Yıldız Technical University Istanbul, Türkiye

Müfit Bahadir Institute for Ecological and Sustainable Chemistry TU Braunschweig Braunschweig, Niedersachsen, Germany

Andreas Haarstrick Waste and Resource Management Leichtweiß Institute for Hydroengineeri Braunschweig, Germany

ISSN 2731-3166 ISSN 2731-3174 (electronic) Water and Wastewater Management Global Problems and Measures ISBN 978-3-031-36297-2 ISBN 978-3-031-36298-9 (eBook) https://doi.org/10.1007/978-3-031-36298-9 © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature Switzerland AG 2023 This work is subject to copyright. All rights are solely and exclusively licensed by the Publisher, whether the whole or part of the material is concerned, specifically the rights of translation, reprinting, reuse of illustrations, recitation, broadcasting, reproduction on microfilms or in any other physical way, and transmission or information storage and retrieval, electronic adaptation, computer software, or by similar or dissimilar methodology now known or hereafter developed. The use of general descriptive names, registered names, trademarks, service marks, etc. in this publication does not imply, even in the absence of a specific statement, that such names are exempt from the relevant protective laws and regulations and therefore free for general use. The publisher, the authors, and the editors are safe to assume that the advice and information in this book are believed to be true and accurate at the date of publication. Neither the publisher nor the authors or the editors give a warranty, expressed or implied, with respect to the material contained herein or for any errors or omissions that may have been made. The publisher remains neutral with regard to jurisdictional claims in published maps and institutional affiliations. This Springer imprint is published by the registered company Springer Nature Switzerland AG The registered company address is: Gewerbestrasse 11, 6330 Cham, Switzerland

Foreword

Wastewater Management and Technologies is a very timely and welcome addition to the growing body of evidence to inform wide audiences of the availability and actual, real-world examples of the various options to tackle wastewater. In an era of growing water scarcity, more frequent and more severe droughts and floods, visible, less visible, and invisible water and soil quality challenges, wastewater has long been synonymous with both neglect and oversight. It is a long-standing neglect which is marked by the widespread practice of releasing large quantities of wastewater into the environment without any tangible, if any, treatment. And this applies only if it is collected at all. The global estimate of untreated wastewater release into the environment is about 80% [1] in many developing countries. The neglect, coupled with the increasing production and generally continuous nature of wastewater, build-up of pollutants in the receiving bodies as well as the introduction of new, emerging pollutants, and a lack of ability to implement globally accepted “user pays” and “polluter pays” principles, has resulted into problems of gigantic proportions. Marine, as well as terrestrial and aquatic ecosystems, are hard hit by the impacts of the wastewater pollution, leading to losses in livelihoods, fisheries, and food chains. Contaminated drinking water, together with substandard handwashing and insufficient sanitation facilities, has become a major health hazard leading to hundreds of thousands of deaths year after year, nearing, perhaps exceeding one million globally, a figure that we are uncertain of due to the lack of reliable data and information. As a matter of fact, less than one-third of countries in the world have reliable data on the production, treatment, and use of wastewater. And you guessed it right: majority of these are upper middle income and high-income countries. Wastewater is an oversight when not neglected because, to most, it is a burden whose disposal is costly and difficult and whose potential for reuse and by-product recovery is under-recognised. This picture has started to change for some time, and we have been seeing innovations using both lowand high-technology solutions and sound, adaptive management approaches that support societal objectives of well-being and health; water, food, and energy security; sustainable development; and climate benefits. The four primary action options involved, i.e. addressing pollution at its source, contaminant removal from wastewater flows, reuse of the reclaimed water v

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and recovery of by-products, offer a wide array of combinations that can be mixed and matched to produce custom-made solutions to the biophysical, social, cultural, and economic specificities and in line with the available financial, institutional, and human capabilities. Wastewater Management and Technologies brings together a wealth of expertise through the authors who ably present the large spectrum of options available, from nature-based solutions to varying levels of technology, and diverse experiences using case studies from around the world, with special emphasis on developing country circumstances. Worthy of mention are the coverage of pharmaceuticals and microplastics; resource recovery; and a welcome inclusion of the lesser understood potential of wastewater management in climate change mitigation. Wastewater Management and Technologies is poised to be a standard reference material for professionals, practitioners, and decision makers in public sector, industry, and agriculture around the world.

Dr. Olcay Ünver Ph.D. Professor of Practice Arizona State University Member, Water Policy Group Adjunct Industry Fellow Australian Rivers Institute Arizona, USA

Reference 1. Bijekar, S., Padariya, H. D., Yadav, V. K., Gacem, A., Hasan, M. A., Awwad, N. S., Yadav, K. K., Islam, S., Park, S., & Jeon, B.-H. (2022). The state of the art and emerging trends in the wastewater treatment in developing nations, Water 14(16), 2537.

Preface

The situation of water resources is deteriorating day by day due to rapid growth of the world population, changes in land use, expansion of productive activities (e.g. agriculture, industry, and tourism), and urban development. It is now common knowledge that water demand has increased exponentially in recent decades. From 1980 to today, water consumption worldwide has increased by an average of 1% per year (UN World Water Report, 2019). Projections until 2050 indicate an increase in freshwater consumption of 20– 30% [4]. Given these projections, an urgent and legitimate question is to what extent treated wastewater can be reused. Global examples show that more than 75% of fresh water is used in agriculture. This is not an option for the future in view of climate change. It therefore makes sense to bring wastewater treatment up to a quality level as quickly as possible, so that it is possible to trickle wastewater onto the fields. Of course, this must be monitored, because it makes no sense to distribute heavy metal ions, micropollutants or microplastics evenly with the wastewater and thus ultimately contaminate the soil and groundwater. However, there are positive approaches and studies. Numerous studies have demonstrated the suitability of treated wastewater for agriculture. Among the advantages of using wastewater for irrigation include increased reliability availability, crop stability, recovery of nutrients useful to the soil, reduction in the use of fertilisers, and a minimisation of pollution from discharges [1, 2]. However, despite all the positive approaches, there is a great need for knowledge to enable the development and implementation of new wastewater management systems for agriculture. This is an investment in the future for people, because not only can the efficiency and useful life of this valuable natural resource be extended but food production and supply can also be secured on a global scale. If we look at available and used wastewater treatment technologies, the effectiveness of wastewater treatment has steadily improved. This is an important signal, as the main objective of the plants is to remove the various impurities and contaminants present in the wastewater such as suspended solids, organic carbon, nutrients, inorganic salts, heavy metals, pathogens. Various physical and chemical treatment methods have been applied such as biodegradation, ion exchange, chemical precipitation, adsorption, chemical precipitation, reverse osmosis, coagulation, flocculation. However, the

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quality of treated wastewater continues to be challenged by the presence of a diverse array of pathogens and anthropogenic chemicals that enter urban and rural waters is a challenge. Wastewater from municipal and industrial treatment plants is considered one of the major contributors to the pollution of water bodies around the world. In many developing countries, still most of treated domestic and industrial wastewater is discharged into surface water systems without reuse. Moreover, in many developing countries, the majority of domestic and industrial wastewater is discharged into surface waterways without any treatment. In addition to the focus of this book, which is to look at efficient wastewater treatment technology, this introduction will also look at the implementation of appropriate Wastewater Management and Technologies. An initially simple observation is that functioning and sustainable wastewater management starts at the household level and depends largely on the human component. Only when perception of the need and perhaps the expectation for a wastewater reuse system at the user level is internalised at the user level, further planning and implementation can be successfully carried out. Therefore, the selection of a wastewater treatment technology must be accompanied in advance by a detailed examination of self-sufficiency and technological capacity. To achieve satisfactory performance, qualified operation and maintenance are essential. This is particularly important when a high degree of reduction of pathogens, heavy metals, and micropollutants and microplastics is required, so that the wastewater can be utilised for irrigation and organic soil improvement finally. Recent developments in wastewater treatment technology include novel physicochemical, biological, and advanced oxidation processes, which are modified processes or combined with nanomaterials and/or newly developed products to improve the performance of current treatment processes. In addition, membrane processes include recent research on the removal of toxic chemicals through various membrane bioreactors and reverse osmosis processes, as well as new techniques to assess membrane integrity and the propensity for membrane fouling. Further innovations are being made in novel flotation technologies suitable for the removal of micropollutants and microplastics. Another method of wastewater treatment is based on nature-based solutions. This method is also suitable for decentralised use in combination with other (conventional) wastewater treatment technologies (hybrid systems). Among the current treatment technologies used for reuse of municipal wastewater for irrigation, constructed wetlands are one of the most suitable technologies. One major advantage is the low maintenance costs and low energy demand. The behaviour and efficiency of constructed wetlands in terms of wastewater treatment depend mainly on macrophyte composition, substrate, hydrology, surface loading, type of inflow, availability of microorganisms, interconnection of various wetland plants, and microbial organisms in the rhizosphere supporting the treatment of wastewater. According to this, constructed wetlands are very effective in removing organic matter and suspended solids, while nitrogen removal is relatively

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low, but—as mentioned—the combination of different types of constructed wetlands or in combination with conventional systems, limits can be reliably met. Braunschweig, Germany

Andreas Haarstrick Müfit Bahadir

References 1. Bixio, D., Thoeye, C., De Koning, J., Joksimovic, D., Savic, D., Wintgens, T., & Melin, T. (2006). Wastewater reuse in Europe. Desalination 187, 89–101. 2. Oteng-Peprah, M., De Vries, N., & Acheampong, M. (2018). Greywater characterization and generation rates in a peri urban municipality of a developing country. Journal of Environmental Management 206, 498–506. 3. The United Nations World Water Development Report 2019—Leaving No One Behind. Available, online: https://Unesdoc.unesco.org/in/rest/annotationSVC/Download WatermarkedAttachment/attach_import_77a13b04-19c4-4368-b0d0-8f9c6bf1349f?= 367306eng.pdf&to=201&from=1 (Accessed on 9 March 2020). 4. WWAP (United Nations World Water Assessment Programme)/UN-Water. (2018). The United Nations World Water Development Report 2018: Nature-Based Solutions for Water; UNESCO. Available online: http://repo.floodalliance.net/jspui/handle/ 44111/2726.

About This Book

This special edition on Wastewater Management and Technologies brings together a wealth of expertise by the authors, who exemplify the wide range of options available—from nature-based solutions to different levels of technology—and the different experiences through case studies from around the world, with a particular focus on conditions in developing countries. This book is part of a book series (special editions) based on the publication of the book “Water and Wastewater Management”, published by Springer in 2022 (ISBN 978-3-030-9528-7). The part about “Wastewater Management (Technologies)” edited in this book will be deepened with this first special edition in terms of technological topics.

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Contents

Low-Cost Wastewater Treatment Constructed Wetlands and Resource Protection . . . . . . . . . . . . . . . Elina Domscheit Horizontal Subsurface Flow Constructed Wetlands in Arid and Semi-Arid Areas—A Review . . . . . . . . . . . . . . . . . . . . . . . . . . . Maria Benbouzid, Naif Al-Jadabi, Souad El Hajjaji, Najoua Labjar, Driss Dhiba, and Abdelmalek Dahchour Anaerobic Dynamic Membrane Bioreactors for the Domestic Wastewater Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Eyüp Debik, Neslihan Manav-Demir, Esra Celik, and Aslican Ihtiyaroglu Agro-Industrial Wastewater Management—Case Studies of Wastewaters from the Olive Industry and Pig Farming—Advantages of Anaerobic Co-Digestion for Small Units in Remote Areas . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Georgios Pilidis and Ioannis Zarkadas

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Advanced and Novel Treatment Technologies Electrode Materials and Their Effects on Electricity Generation and Wastewater Treatment in a Microbial Fuel Cell . . . . . . . . . . . Andika Wahyu Afrianto and Sandhya Babel

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Microbial Electrolysis Cells for Biohydrogen Generation and Wastewater Treatment—A Short Review and Current Trends . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Victor Alcaraz-Gonzalez, René Alejandro Flores-Estrella, Marcelo Nolasco, Vitor Cano, and Victor González-Alvarez

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Novel Membrane Technologies in the Treatment and Recovery of Wastewaters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Mehmet Emin Pasaoglu, Recep Kaya, and Ismail Koyuncu

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Application of Supported Fenton and Fenton-Like Catalysts in the Degradation of Pharmaceuticals in Wastewater—A Review of New Technologies in the Last Decade . . . . . . . . . . . . . . 107 Sanaa Rashid, Dominic Bale, and Katherine Huddersman Combined Ferrite Treatment of Multi-component Wastewaters Under the Elevated Temperature . . . . . . . . . . . . . . . . . . . . . . . . . . . 125 Gheorghe Duca, Victor Covaliov, Olga Covaliova, and Lidia Romanciuc Magnetically Separable and Reusable Mag-Mg/Al Layered Double Hydroxides for the Adsorption of Disperse Dyes of Navy Blue and Yellow F3G . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 137 Sri Juari Santosa, Lutfia Isna Ardhayanti, Desy Permatasari, and Narsito Heavy Metal Removal from Wastewater Using Different Cheap Adsorbents: Olive Cake, Moringa, Eucalyptus, and Pine Cone . . . 153 Mohammed Matouq, Zaid Al-Anbar, Mohammed Al-Anber, and Omar Al-Ayed Treatment Technologies in Developing Countries Leachate Treatment in Brazil and Potential Technologies: A General Approach . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 167 Mariana Islongo Canabarro, Siara Silvestri, Victor Alcaraz Gonzalez, and Elvis Carissimi Sanitation Context and Technological Challenges to Municipal Wastewater Management in Africa . . . . . . . . . . . . . . . . . . . . . . . . . 183 Patrick Ogola Onyango Management of Water and Wastewater in Morocco and Arab Countries . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 193 Abdelmalek Dahchour and Souad El Hajjaji New Concerns on Treatment Technology Paradigm Shift in Domestic Wastewater Treatment: Toward Energy Minimization, Greenhouse Gas Emission Reduction, and Resources Recovery . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 211 Bilge Alpaslan Kocamemi, Sümeyye Çelik, Abdullah Bugra Senol, Halil Kurt, and Esra Erken Microplastics Removal Performance Through Advanced Treatment Technologies: A Mini Review . . . . . . . . . . . . . . . . . . . . . 239 Hanife Sari Erkan and Guleda Onkal Engin

Contents

About the Editors

Eyüp Debik graduated from Department of Environmental Engineering at Istanbul Technical University in 1992 and received his doctorate degree from Yildiz Technical University in 1999 in the field of physicochemical treatability of domestic wastewater in large-scale applications. He had research activities at Iowa State University in the USA (2002–2004) in the field of anaerobic treatment processes and wastewater treatment using fungus. He received the title of Professor in 2015 at Yildiz Technical University in Environmental Engineering and has been working in the same department ever since. His teaching and research cover the application and designing the biological and chemical treatment technologies of wastewaters, anaerobic and aerobic digestion processes, capacitive membrane deionization processes, and nanoparticle applications in degradation of persistent organic pollutants. He has several on-going national and international projects. He had several management experiences including the Director of Technical Vocational School of Higher Education (2014–2020), Manager of Technology Transfer Office (2014–2021) at Yildiz Technical University, and Advisor of General Secretary in Istanbul Metropolitan Municipality.

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About the Editors

Müfit Bahadir studied chemistry at the Free University Berlin and Bonn University in Germany, received his Ph.D. 1975 from Bonn University and his Assoc. Prof. 1988 from Munich Technical University. He became a full professor of Environmental and Sustainable Chemistry at the Technische Braunschweig in 1989. Since 1997, he is an honorary doctor and visiting professor at the Selcuk University Konya in Turkey. His research fields cover environmental chemistry and analyses, environmental pollution through industrial processes and products, pesticide chemistry and metabolism in soil and water, ecotoxicology, sustainable chemistry, renewable feed stocks, and biodiesel and bio-lubricants. During the last twenty years, his research focused on R&D of sustainable water and wastewater management in developing countries. He retired in 2016. Andreas Haarstrick studied Chemistry at the Technische Universität Braunschweig (1983– 1989) and received his doctorate in 1992 in bioengineering. Since 2006, he is Professor for Bioprocess Engineering at the TU Braunschweig. His teaching and research cover modelling biological and chemical processes in heterogeneous systems, development of models predicting pollutant reduction in and emission behaviour of landfills, growth kinetics at low substrate concentrations under changing environmental conditions, advanced oxidation processes (AOP), and groundwater management. Since 2012, he is the managing director of the DAAD exceedSwindon Project dealing with sustainable water management in developing countries.

Low-Cost Wastewater Treatment

Constructed Wetlands and Resource Protection Elina Domscheit

the soil pore volume and thus hydraulic conductivity. This phenomenon can limit wastewater infiltration sufficiently but may also lead to decreasing treatment performance. Quantification of water exchange is not trivial. Thus, in this review, exemplary studies are also presented, where exchange between surface water and groundwater took place.

Abstract

Constructed wetlands can be utilized for treatment of a wide range of water types. They convince among other reasons due to their low-maintenance requirements compared to conventional wastewater treatment plants. On the one hand, CWs can help to clean up and to protect water resources. On the other hand, they pose a potential risk of groundwater pollution due to infiltration of contaminated water. Furthermore, water exchange between surface water in the CW and groundwater can influence removal efficiencies. To detect water exchange, the CW water budget can be used. In literature, a wide range from 3 to 47% of seepage losses can be found. To prevent water exchange and potential groundwater pollution, CW cells should be sealed. Sealing can be done either with artificial or with compacted in-situ soils. A good monitoring system can help to ensure groundwater security in the long term. Clogging reduces

E. Domscheit (&) Lower Saxony Water Management, Coastal and Nature Protection Agency (Niedersächsischer Landesbetrieb für Wasserwirtschaft, Küsten- und Naturschutz), Bürgermeister-Münchmeyer-Str. 6, 27283 Verden (Aller), Germany e-mail: [email protected]

Keywords







Clogging Constructed wetlands Groundwater Sealing Water budget

1







Introduction

Constructed wetlands (CW) developed into a secure non-conventional wastewater treatment technology that convinces due to the wide range of possible utilization, such as treatment of domestic, agricultural, and industrial wastewater as well as for runoff, mine, and leachate waters [1–3]. CWs can provide low-cost, low-energy, and low-maintenance water treatment in decentralized settings [4] and show sufficient removal rates for standard parameters as well as for emerging contaminants. They persuade due to their strong adaptability [5], low sludge generation [1], and good self-purification capacity [6]. Furthermore, CWs provide ecosystem services, improve ecosystem health [4], and have positive

© The Author(s), under exclusive license to Springer Nature Switzerland AG 2023 E. Debik et al. (eds.), Wastewater Management and Technologies, Water and Wastewater Management, https://doi.org/10.1007/978-3-031-36298-9_1

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effects on local climate parameters such as precipitation, temperature, and humidity [7]. The typically named disadvantage of CWs is the requirement of comparatively high land area. Additionally, there should also be a focus on groundwater aspects. Driven by among others climate change-induced water scarcities, the understanding of the importance of groundwater resources for human consumption, agricultural and industrial uses, and ecosystem health increased and highlighted the need of groundwater quality evaluation [8]. On the one hand, CWs help to treat polluted water and thereby help to protect resources such as surface waters and groundwater [9]. Some CWs are even used to remediate groundwater [3]. Next to that, Jahangir et al. [10] showed a supplemental environmental benefit of CWs next to the regular water purification. They presented that integrated constructed wetlands (ICW) can contribute to further decrease of nitrate–nitrogen in the underlying groundwater with the help of enhanced denitrification and dissimilatory nitrate reduction to ammonium (DNRA). On the other hand, there is a potential risk of groundwater contamination due to infiltration of pollutants and nutrients [11]. Furthermore, groundwater flow could transport nutrientand contaminant-rich infiltration water to surrounding and nearby surface waters [12]. The quality of (shallow) groundwater beneath CWs concerning, e.g., the carbon and nitrogen species is highly unknown [8, 13]. In general, groundwater should be considered in relation to hydrological CW design parameters and when determining removal efficiencies [3]. This is of importance because interaction between surface water and groundwater may occur [14]. Next to that, groundwater exchange in wetlands is a driving factor in biogeochemical processes [15]. On the one hand, groundwater can have influences on the removal efficiencies within the CW [14]. On the other hand, nutrient and contaminant losses due to infiltration must be considered when calculating removal rates [14, 16]. By now, most CW research focused on nutrient cycling only in surface waters, while groundwater seepage into and out of the CW stayed mostly unattended [16].

E. Domscheit

Deliberately applying contaminated water to a previously unpolluted area requires careful consideration. Possible risks must be determined and limited. This paper will set the basis of the discussion by summarizing the water budget of CWs. It will then be determined, which factors may influence the groundwater pollution. An overview of different methods of water exchange detection will be given before the focus will be on the probability of a groundwater contamination beneath CWs. Finally, there will be a look on how to prevent or rather to minimize groundwater pollution.

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Water Budget of CWs

The installation of a CW system can change the flow dynamics and chemistry of the underlying groundwater [16]. To derive statements about the water exchange between CWs and groundwater, the overall water budget of CWs has to be considered [17]. Generally, the water balance of a CW is set up by the inflow, storage, and outflow of water [18]; (Fig. 1). Hydraulic aspects include the climate and weather as well as the hydraulic residence time, hydraulic loading rate, groundwater exchange, and evapotranspiration [18]. Usually, precipitation and evapotranspiration account only for a small part of the water budget [19]. Kynkäänniemi et al. [20] used the water balance to detect the water exchange between wetland cells and groundwater. In their study, the outflow was set in relation to the inflow and was found to be only 88%. Regarding the water table measurements in the monitoring wells, it could be found that water from the wetland cells percolated downward and reached the underlying groundwater [20]. In comparison, Larson et al. [21] found total seepage volumes for two wetlands of 27% and 47%, respectively, of the total inlet flow. Ackerman et al. [16] used simulations with MODFLOW to quantify the water loss of a CW to the groundwater below. They concluded that the groundwater seepage accounted between 3 and 11% of the surface water flux through the wetland. The authors state that these values might be on the lower end of seepage rates when

Constructed Wetlands and Resource Protection

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Fig. 1 General water budget of a CW system (own drawing based on information from [18])

comparing the values with other studies. Choi and Harvey [22] suggested in their study that 31% of the water percolates in a CW to the groundwater. In contrast, Kadlec and Wallace [3] assumed that groundwater interactions account only for a minor flux within the wetland water budget. Generally, it can be stated that the seepage flow term is difficult to determine. In the study of Favero et al. [23], the seepage flow term of the investigated freshwater surface CW had the highest uncertainty of the water budget parameters since it is the most difficult to estimate. The uncertainty of this parameter could be minimized by adjustment of the groundwater level measurement intervals. Increasing the measurement frequency from 18 to 5 days in the study led to a reduction of the water budget error of 10% [23].

of a CW with the so-called treatment technology “groundwater percolation” as a polishing step [24]. The most prominent example of treatment wetlands using natural liners is integrated constructed wetlands (ICWs) [11, 25]. Typically, the ICW concept uses in-situ soils to construct and to line their cells by compaction of the material to decrease permeability [11, 18, 25]. According to McCook [26], the structure and resultant permeability of a clay liner are mainly affected by the following three factors:

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Depending on the soil parameters and the compound, pollutants might be hold back according to the retention and degradation capacity of the soil. For instance, ammonium can be fixed in soil clay until the point of saturation [10]. Further influences on the hydraulic conductivity might be the season, since Larson et al. [21] speculates that hydraulic conditions may increase in the summer due to flow paths that are created by channels from roots and burrowing insects. An additional sealing effect of natural liners can be provided through soil clogging [25]. With this phenomenon, hydraulic conductivity decreases by reduction of the pore volume [27] through physical, chemical, or biological

Factors Influencing Groundwater Pollution Through CWs

3.1 Sealing and Clogging Quantities of water exchange between a CW and groundwater depend on the site characteristics and the CW type as well as on CW construction and operation. However, the biggest influence on water exchange and possible groundwater pollution has the type of sealing. It can be assumed that artificial liners made of plastic or concrete prevent an exchange between the CW water and groundwater. However, few systems exist that do not use artificial liners. Some of the CWs without sealing are constructed to combine the advantages

• soil type used for the liner; • density and water content to which the clay is compacted; • construction procedures used for spreading, processing, wetting, compacting, and bonding between lifts.

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mechanisms [3, 25]. Physical clogging occurs, when suspended solids and organic particulates are deposited, filtrated, and accumulated [3, 25]. In this case, especially the pores near the surface get blocked [3]. Chemical sealing results from salt-induced dispersion of clay [25] or chemical precipitation in the pores [3]. Biologically induced clogging occurs through the growth of polysaccharides, other microbial by-products, or biofilms that seal the interstices in the hydrated clay liner [3, 25, 28]. According to Dzakpasu et al. [25], several studies show that the selfsealing effect can sufficiently limit wastewater infiltration to rates that would not significantly contaminate shallow groundwater. Generally, the following parameters affect clogging of CWs: substrate porosity, plants, hydraulic load, oxygen supply conditions, organic loading, and water depth, with organic loading and substrate porosity having the most significant impact [29]. The development of a biomat within the wetland soils requires organic matter and humic substances [3, 27]. Mustafa et al. [27] could show the relationship in their study, where cells with the highest organic matter content had the lowest infiltration rates. On the one hand, next to the sealing aspect, the formation of a biomat supports the reduction of pollutant concentrations [27]. On the other hand, clogging is a serious concern in the development of CWs [29] since it leads often to a decreasing treatment performance [30]. Since the major treatment principle of CWs is that wastewater flows through the porous medium planted with wetland vegetation to achieve removal of pollutants [31], sealed pores cannot provide this condition anymore. Furthermore, the hydraulic design parameters may not match anymore due to increasing surface flow and evaporative losses [32]. Further information about soil clogging and how to prevent clogging in CWs can be found at Wang et al. [29].

3.2 Other Factors Next to CW sealing, there are few other factors that influence potential groundwater pollution.

E. Domscheit

An important aspect is the depth to the water table since it has an impact on potential further treatment of the infiltrating water [33]. Furthermore, the type of water treated plays an important role. On the one hand, the concentration ranges of pollutants in the untreated water must be considered [25, 33]. On the other hand, the type of parameters and their environmental properties play a role when it comes to pollutant transfer. Soluble compounds can be transported with the percolating water, while particulate-bound contaminants need other flow paths through the soil.

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Determination of Water Exchange Between CWs and Groundwater

Quantifying the amount of water that is exchanged between the wetland cells and the groundwater beneath is not trivial. Sometimes, it is already difficult to certainly state that there is a water exchange taking place. To come to a secured statement, Dzakpasu et al. [25] used zero-tension pan lysimeters of 900 mm diameter to provide a sampling possibility of the infiltrating water. The lysimeters were constructed and placed in the ground below the first wetland cell during the ICW construction. More commonly, conclusions are drawn from measured quality parameters. In many studies, groundwater wells for sampling are installed. Results are either compared with surface water [34] or groundwater quality parameters after the CW get compared with the ones before the CW [8]. Jahangir et al. [13] presented a third possibility by comparison of parameters in groundwater beneath the CW and in disconnected groundwater. In the following, few exemplary studies will be presented. Ouyang [8] compared quality parameters of shallow groundwater from monitoring wells before and behind the wetland system. The studied CW was constructed to treat storm water. Even though the average concentration of phosphorus, arsenic, chromium, nickel, and zinc decreased, an increase could be determined for

Constructed Wetlands and Resource Protection

total Kjeldahl nitrogen, ammonium, chloride, and sulfate. Based on that, Ouyang [8] concluded that the treatment system had a noticeable effect on some of the analyzed parameters in the shallow groundwater below the CW. Lazareva and Pichler [35] examined the sodium and chloride fluxes of a wetland, which was constructed with the aim to treat wastewater and surface waters. To determine the parameter fluxes, six monitoring wells were installed along the flow path. Samples were taken once a month. First of all, the analyzed water samples indicated that there was little to no leakage from surface water bodies in the north and south of the wetland into the wetland treating system. Furthermore, it could be shown that the composition of the water within the wetland system was significantly different than the one in the groundwater. Instead, the chemical composition of the water samples taken out of the monitoring wells was much closer to the surface water and groundwater from the Surficial Aquifer System [35]. In the study of Ackerman et al. [16], the CW treated municipal wastewater effluent. Indication that water from the CW infiltrated into the groundwater could be seen from a temporal point of view, since groundwater concentrations of chloride were like non-impacted groundwater before the CW start-up but increased rapidly in the month after the treatment system started. Besides, effluent-groundwater mixing calculations using chloride as a conservative tracer indicated that most of the sampled groundwater was dominated by the received municipal wastewater effluent (64–100%) [16].

5

Probability of Groundwater Pollution Underneath CWs

Typically, CWs for domestic wastewater, agricultural wastewater, and mine drainage are sealed to prevent infiltration and percolation of contaminated water through the soil and to the water table [18]. In contrast, there are situations, in which sealing is not needed or not feasible [3]. The use of liners is, e.g., often dispensed with in large wetlands treating non-toxic contaminants [3]. Some

7

storm water wetlands are even constructed to receive groundwater to ensure sufficient base flow [18]. Besides, ICWs are generally not sealed with artificial material. Depending on the hydraulic conditions, unsealed wetland systems are likely to receive groundwater, when the water table is high, and may discharge to groundwater, when the water table is low [18]. Furthermore, damage on the wetland liners, either plastic membranes or clay layers, e.g., due to plant roots that puncture the liners, might possibly allow wastewaters to infiltrate to the groundwater [3]. Even though the infiltrating water may receive some additional treatment by microbial degradation or adsorption during the passage through the soil, not all potential pollutants are subjected to soil and aquifer treatment [3]. Multiple examples exist, where CW storages were partly or fully connected with the groundwater aquifer, and the significance of the interaction between surface water and groundwater on hydrological processes and contaminant transport could be shown [14, 17, 19]. In the following, studies will be presented in which water exchange and/or groundwater pollution could be determined. In Dzakpasu et al. [11, 25], the studied fullscale ICW system treating domestic wastewater was characterized by a series of interconnected free water surface flow constructed wetland cells, which were lined with 500 mm thick local subsoil material. The subsoil material contained a mixture of alluvium, organic soils, tills, and gravel with a calculated hydraulic conductivity of 9  10–11 m/s. Even though the findings of the study showed a connection between ICW cells and groundwater, it was concluded that the ICW system had only a minor influence on the groundwater quality [11]. Nevertheless, this might only count under conditions, when the level of the groundwater table is above the basis of the wetland cell. Under conditions with the level of the groundwater table being below the basis of the wetland cell, the water will most likely reach the groundwater [11]. In Dzakpasu et al. [25], the amount of infiltration below the ICW cells increased from the proximal cells to the distal one. In the first two cells, less than

8

0.5% of the influent loading of most of the considered contaminants was lost through infiltration. In comparison, approx. 2% was lost in the third cell [25]. The ICW site in SE Ireland, studied by Jahangir et al. [13], was constructed with earthen liners at the base of the cell, which may allow different parameters to migrate across the liner into the groundwater. The results of the study showed that geochemical conditions in groundwater below the ICW liner corresponded with very low DO, Eh, and SO42−. Likewise, the concentrations of reactive carbon and nitrogen in this connected groundwater were significantly higher than those of the disconnected upgradient groundwater. By this, the direct link of the ICW with the underlying groundwater geochemical properties and processes could be presented. The DOC, DIC and dissolved CO2, and CH4 concentrations as well as TN concentration in the underlying groundwater were significantly higher than in the control [13]. Comparing parameter trends in groundwater wells before and after the CW, Ouyang [8] could show a decrease in the average values of water levels, redox potential, and conductivity. Increasing organic nitrogen concentrations were explained with leaching from decomposed organic matter, which has been accumulated in the CW. Penetration of contaminants from surface water into shallow groundwater was also shown by Wu et al. [34]. The Xiantao CW was constructed with natural resources and lacked an anti-seepage layer. The long-term monitoring data clearly showed a positive correlation of pollutant concentrations in surface water and shallow groundwater for COD, NH3-N, TN, and TP indicating the relation of the surface water to the groundwater quality [34]. Several studies focused on nitrogen fluxes leaving the wetland cells by infiltration and seepage. Larson et al. [21] found 4–10% of the total inlet NO3-N load in this context. Kadlec and Wallace [3] estimated annual nitrogen leakage fluxes of 24.7%. Ackerman et al. [16] found that the flux of nitrate lost during seepage to groundwater was 3% for a 2-day retention time, while a 7-day retention time led to a nitrate loss

E. Domscheit

flux of 10%. Summing up all nitrate removal processes, up to 30%, has occurred along the groundwater pathway [16]. A special case is presented by Kill et al. [36] focusing on the influence of groundwater on an in-stream free surface flow CW. During the construction of the CW, no specific sealing materials such as geomembrane or clay were used. Thus, surface water and groundwater could exchange. The study showed that groundwater seepage to the wetland affected the CW operation and nutrient removal efficiency due to increased flow rates in the wetland, which in turn might decrease water retention time and increase nutrient concentration in case of nutrient rich groundwater. Concentrations of TIC, TN, NO3, and NO2-N in the CW outlet increased by 9.7%, 27.7%, 31.6%, and 15.1%, respectively. Kill et al. [36] concluded that these increases resulted from groundwater inflow. Another perspective on the surface water groundwater exchange aspect in CWs is presented by Jahangir et al. [10]. They could demonstrate a positive effect of the studied ICW system on the underlying groundwater by further attenuation of nitrate–nitrogen through enhanced denitrification and dissimilatory nitrate reduction to ammonium (DNRA). The ICW system was used to treat municipal wastewater and storm water from a village of 500 persons’ equivalent. It was set up in a series of interconnected ponds. The wetland cells were lined with local soil material, which consisted of 26% sand, 47% silt, and 27% clay. The saturated hydraulic conductivity could be reduced to 1  10–8 m/s by compaction of the soil material to a 0.5 m layer. The determined mean vertical flow rate of water from CW cells to groundwater was 74 mm/year.

6

Prevention and Minimization of Groundwater Pollution Through CWs

When selecting a site for construction of a treatment wetland, several aspects should be considered to prevent groundwater pollution.

Constructed Wetlands and Resource Protection Table 1 Different recommendations of soil properties for natural sealing of CWs

9

Permeability coefficient (of compacted soil) < 10–7 m/s  10–8 m/s < 10–6 cm/s –6

10 –10

–7

cm/s

Sealing plays the key role, as it separates the wetland cell from the groundwater and avoids water exchange in both directions [18]. If applicable and in conformity of regulatory requirements, the use of locally available material for natural liners is usually preferred to reduce costs [3]. However, deciding which liner to install depends also on the project goals [3]. Artificial liners are not feasible for systems of more than few hectares, while for smaller CWs, plastic liners may be [3]. Recommendations for sealing with in-situ soils vary according to the source. Table 1 gives an overview of the different suggestions. According to Davis [18], soils that contain more than 15% clay usually show the necessary properties for sufficient compaction. McCook [26] points out the importance of efficient construction processing of the compacted liner and further puts emphasis on the uniformity of the process. Furthermore, McCook [26] suggests inspecting an excavated part of the compacted liner, while Davis [18] advices a laboratory analysis of the construction material already before choosing a sealing method. Further guiding information for compaction can be found in McCook [26]. To prevent penetration of the natural liner by plant roots, it should be covered by a layer with rooting medium of 30–40 cm [3] (approx. 7– 10 cm according to Davis [18]). If no adequate soils are available on site, it is recommended to use artificial liners like asphalt, synthetic butyl rubber, and plastic membranes [18]. Generally, the liner material should be strong, thick, acid resistant, alkali proof, frost, UV resistant, root and rodent resistant, non-toxic, easy to carry and move, and preferably made of recyclable materials [18, 24]. Different artificial liner materials are available with 0.76 mm

Thickness of sealing

Source

> 30 cm

[37]

> 60 cm

[38]

Not specified

[18]

30 cm

[3]

polyvinyl chloride (PVC) and 1.0 mm highdensity polyethylene (HDPE) being the most common ones [3]. Due to its characteristics, PVC liners are usually used for small projects with CWs of less than 0.1 ha, while HDPE liners are generally delivered in rolls applicable for larger projects [3]. Further advantages and disadvantages of the different artificial liner material can be found in Kadlec and Wallace [3]. In case that the construction site contains angular stones, sand bedding or geotextile cushions should be placed below the liner to prevent punctures [18]. If CW applications involve hazardous substances, e.g., landfill leachates, double liners may be required [3]. Especially, when the CW is designed to treat mine drainage, the reaction of the natural or artificial liner should be tested before usage, since some soils and synthetics are affected by acids that can be found in mine drainages [18]. In case CW systems are neither sealed with natural nor with artificial liners, a specific vertical travel distance to the groundwater table should exist (typically about 1 m of unsaturated soil) [3]. Next to the sealing aspect, management and monitoring are important measures to prevent or rather to minimize water exchange between CWs and groundwater beneath [34] and to have the possibility to react in cases, where liners functions are decreasing. To ensure that the CW does not contaminate groundwater, groundwater should be monitored at least once or twice a year [18].

7

Conclusion

Constructed wetlands are multifunctional naturebased treatment systems that help to clean up different water sources and thereby to protect crucial resources. However, while designing and

10

constructing a CW, groundwater should never be neglected. Depending on the site characteristics and inflow water specifications, sealing of the wetland cells is highly recommended to prevent water exchanges between surface water and groundwater and to minimize potential groundwater pollution. In case of utilization of natural liners, careful planning and design as well as a good compaction practice should not be underestimated. A good monitoring system can help to ensure groundwater security in the long term. Cleaning of water resources should not only focus on surface waters but should always consider groundwater as well. Purification of surface water at the expense of groundwater should be avoided, having the water cycle in mind.

References 1. Gökalp, Z., Karaman, S., Taş, I., & Kirnak, H. (2016). Constructed wetland technology to prevent water resources pollution. Current Trends in Natural Sciences, 5(9), 125–132. 2. Vymazal, J. (2010). Constructed wetlands for wastewater treatment. Water, 2, 530–549. 3. Kadlec, R. H., & Wallace, S. D. (2009). Treatment wetlands (Vol. 2). Taylor & Francis Group. 4. Avellan, C. T., Ardakanian, R., & Gremillion, P. (2017). The role of constructed wetlands for biomass production within the water-soil-waste nexus. Water Science & Technology, 75(10), 2237–2245. 5. Yang, Q., Wu, Z., Liu, L., Zhang, F., & Liang, S. (2016). Treatment of oil wastewater and electricity generation by integrating constructed wetland with microbial fuel cell. Materials, 9. 6. Liu, S., Song, H., Li, X., & Yang, F. (2013). Power generation enhancement by utilizing plant photosynthate in microbial fuel cell coupled constructed wetland system. International Journal of Photoenergy. 7. Gokalp, Z., & Karaman, S. (2017). Critical design parameters for constructed wetlands natural wastewater treatment systems. Current Trends in Natural Sciences, 6(12), 156–164. 8. Ouyang, Y. (2013). Effects of a constructed wetland and pond system upon shallow groundwater quality. Environmental monitoring and assessment, 185(5), 4245–4259. 9. Ibekwe, A. M., Grieve, C. M., & Lyon, S. R. (2003). Characterization of microbial communities and composition in constructed dairy wetland wastewater effluent. Applied and Environmental Microbiology, 69(9), 5060–5069.

E. Domscheit 10. Jahangir, M. M. R., Fenton, O., Müller, C., Harrington, R., Johnston, P., & Richards, K. G. (2017). In situ denitrification and DNRA rates in groundwater beneath an integrated constructed wetland. Water Research, 111, 254–264. 11. Dzakpasu, M., Scholz, M., Harrington, R., McCarthy, V., & Jordan, S. (2014). Groundwater quality impacts from a full-scale integrated constructed wetland. Groundwater Monitoring & Remediation, 34(3), 51–64. 12. Hathaway, J. M., Cook, M. J., & Evans, R. O. (2010). Nutrient removal capability of a constructed wetland receiving groundwater contaminated by swine lagoon seepage. Transactions of the ASABE, 53(3), 741–749. 13. Jahangir, M. M. R., Fenton, O., McAleer, E., Johnston, P., Harrington, R., Müller, C., & Richards, K. G. (2019). Reactive carbon and nitrogen concentrations and dynamics in groundwater beneath an earthen-lined integrated constructed wetland. Ecological Engineering, 126, 55–63. 14. Kazezyılmaz-Alhan, C. M., Medina, M. A., Jr., & Richardson, C. J. (2007). A wetland hydrology and water quality model incorporating surface water/groundwater interactions. Water Resources Research, 43(4). 15. Hunt, R. J., Walker, J. F., & Krabbenhoft, D. P. (1999). Characterizing hydrology and the importance of ground-water discharge in natural and constructed wetlands. Wetlands, 19(2), 458–472. 16. Ackerman, J. R., Peterson, E. W., Hoven, S. V. D., & Perry W. L. (2015). Quantifying nutrient removal from groundwater seepage out of constructed wetlands receiving treated wastewater effluent. Environmental Earth Sciences, 74(2), 1633–1645. 17. Ludwig, A. L., & Hession, W. C. (2015). Groundwater influence on water budget of a small constructed floodplain wetland in the ridge and valley of Virginia, USA. Journal of Hydrology: Regional Studies, 4(B), 699–712. 18. Davis, L. (1995). A handbook of constructed wetlands—a guide to creating wetlands for: Agricultural wastewater, domestic wastewater, coal mine drainage, stormwater in the Mid-Atlantic Region. Government Printing Office. 19. Huang, G., & Yeh, G.-T. (2012). Integrated modeling of groundwater and surface water interactions in a manmade wetland. TAO—Terrestrial, Atmospheric and Oceanic Sciences, 23(5), 501–511. 20. Kynkäänniemi, P., Ulén, B., Torstensson, G., & Tonderski, K. S. (2013). Phosphorus retention in a newly constructed wetland receiving agricultural tile drainage water. Journal of Environmental Quality, 42(2), 596–605. 21. Larson, A. C., Gentry, L. E., David, M. B., Cooke, R. A., & Kovacic, D. A. (2000). The role of seepage in constructed wetlands receiving agricultural tile drainage. Ecological Engineering, 15(1–2), 91–104.

Constructed Wetlands and Resource Protection 22. Choi, J., & Harvey, J. W. (2000). Quantifying timevarying ground-water discharge and recharge in wetlands of the northern florida Everglades. Wetlands, 20(3), 500–511. 23. Favero, L., Mattiuzzo, E., & Franco, D. (2007). Practical results of a water budget estimation for a constructed wetland. Wetlands, 27(2), 230–239. 24. Wendland, C., & Chiarawatchai, N. (unknown). Constructed wetlands for wastewater treatment, Lesson B4, Institute of Wastewater Management, Hamburg University of Technology. 25. Dzakpasu, M., Scholz, M., Harrington, R., Jordan, S. N., & McCarthy, V. (2012). Characterising infiltration and contaminant migration beneath earthen-lined integrated constructed wetlands. Ecological Engineering, 41, 41–51. 26. McCook, D. (1991). Measurement and estimation of permeability of soils for animal waste storage facility design. In Technical note—engineering. Texas: South National Technical Center. 27. Mustafa, A., Scholz, M., Harrington, R., & Carroll, P. (2009). Long-term performance of a representative integrated constructed wetland treating farmyard runoff. Ecological Engineering, 35(5), 779–790. 28. Doody, D., Harrington, R., Johnson, M., Hofmann, O., & McEntee, D. (2009). Sewerage treatment in an integrated constructed wetland. Proceedings of the Institution of Civil Engineers—Municipal Engineer, 162(4), 199–205. 29. Wang, H., Sheng, L., & Xu, J. (2021). Clogging mechanisms of constructed wetlands: A critical review. Journal of Cleaner Production, 295 (126455). 30. Al-Isawi, R., Scholz, M., Wang, Y., & Sani, A. (2015). Clogging of vertical-flow constructed wetlands treating urban wastewater contaminated with a diesel spill. Environmental Science and Pollution Research, 22(17), 12779–12803. 31. Vymazal, J. (2018). Does clogging affect long-term removal of organics and suspended solids in gravel-

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based horizontal subsurface flow constructed wetlands? Chemical Engineering Journal, 331, 663– 674. Ellman, E. D. (1997). Groundwater quality effects and operational considerations of an unlined constructed wetland treating raw sewage. The University of Arizona. Ham, J. M. (2002). Seepage losses from animal waste lagoons: A summary of a four-year investigation in Kansas. Transactions of the American Society of Agricultural Engineers (ASAE), 45(4), 983–992. Wu, H., Gao, X., Wu, M., Zhu, Y., Xiong, R., & Ye, S. (2020). The efficiency and risk to groundwater of constructed wetland system for domestic sewage treatment—A case study in Xiantao, China. Journal of Cleaner Production, 277(123384). Lazareva, O., & Pichler, T. (2010). Long-term performance of a constructed wetland/filter basin system treating wastewater, Central Florida. Chemical Geology, 269, 137–152. Kill, K., Pärn, J., Lust, R., Mander, Ü., & Kasak, K. (2018). Treatment efficiency of diffuse agricultural pollution in a constructed wetland impacted by groundwater seepage. Water, 10(11). Bahlo, K., Conte, G., Ebeling, B., Masi, F., MittererReichmann, G., Platzer, C., Regelsberger, B., Urtane, L., & Wach, G. (2005). Guidelines for sustainable water management in tourism facilities. In M. Regelsberger (Ed.), SWAMP-sustainable water management and wastewater purification in tourism facilities. DWA. (2017). Grundsätze für Bemessung, Bau und Betrieb von Kläranlagen mit bepflanzten und unbepflanzten Filtern zur Reinigung häuslichen und kommunalen Abwassers (Principles for the design, construction and operation of sewage treatment plants with planted and unplanted filters for treatment of domestic and municipal wastewater). Arbeitsblatt DWA-A 262, Deutsche Vereinigung für Wasserwirtschaft, Abwasser und Abfall e.V. (DWA): Hennef.

Horizontal Subsurface Flow Constructed Wetlands in Arid and Semi-Arid Areas—A Review Maria Benbouzid, Naif Al-Jadabi, Souad El Hajjaji, Najoua Labjar, Driss Dhiba, and Abdelmalek Dahchour

Abstract

Horizontal subsurface flow constructed wetlands (HSSFCWs) were first developed by Seidel in the early 1960s and upgraded by Reinhold Kickuth as the Root Zone Method in the late 1960s and early 1970s. These constructed wetlands (CWs) are used in wastewater treatment and replicate the natural processes of pollution removal. HSSFCWs employ substrate bed planted with wetland plants, and the wastewater maintained below

M. Benbouzid
N. Al-Jadabi
S. E. Hajjaji (&) Laboratory of Spectroscopy, Molecular Modeling, Materials, Nanomaterials, Water and Environment, (LS3MN2E-CERNE2D), Chemistry Department, Faculty of Sciences, Mohammed V University in Rabat, Rabat, Morocco e-mail: [email protected] N. Labjar Laboratory of Spectroscopy, Molecular Modeling, Materials, Nanomaterials, Water and Environment, (LS3MN2E-CERNE2D), ENSAM, Mohammed V University in Rabat, Rabat, Morocco D. Dhiba Water Research Institute IWRI, University Mohammed VI Polytechnic (UM6P), Ben Guerir, Morocco

the bed surface flows horizontally from the inlet of the system to its outlet. They are proposed as a sustainable solution instead of other systems such as membrane bioreactors or sequencing batch reactors mainly because of their minimal energy consumption and relatively low construction and operational costs. Indeed, they are very practical for wastewater treatment in decentralized, rural, and remote areas, where conventional treatment is not possible and limited. Nevertheless, they are very effective in removing organic and microbial pollutants. However, for the proper functioning of these filters, some parameters may affect their processing efficiency such as type of vegetations, composition of the substrate, retention time, or temperature. This review focuses on the influence of arid and semi-arid climatic conditions on the effectiveness of the removal of different contaminants in HSSFCWs. Various studies of wastewater treatment using HSSFCWs under these specific climatic conditions, particularly in Morocco, Algeria, Tunisia, and Egypt, will be discussed. Keywords




Horizontal subsurface flow Constructed wetlands Arid Semi-arid climate Wastewater treatment










A. Dahchour Agronomy and Veterinary Institute, Rabat, Morocco © The Author(s), under exclusive license to Springer Nature Switzerland AG 2023 E. Debik et al. (eds.), Wastewater Management and Technologies, Water and Wastewater Management, https://doi.org/10.1007/978-3-031-36298-9_2

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1

M. Benbouzid et al.

Introduction

Water resources are becoming more and more limited and many of them are contaminated and polluted by industrial, agricultural, or household wastewaters. Indeed, to eliminate these various contaminants, such as organic carbon, inorganic salts, suspended solids, nutrients, and heavy metals, the treatment of these wastewaters becomes essential before discharged. In addition, the treatment of wastewater for the purpose of reuse becomes more and more necessary, especially due to climate change and the more and more decrease of water resources [1]. Moreover, by 2030, water scarcity will affect 40% of the world [2]. Traditional processes of wastewater treatment such as activated sludge systems, membrane bioreactors, filters, coagulation, and flocculation using chemicals are not only expensive, energy intensive, and sometimes polluting, but in turn produce by-products such as concentrated waste, sand, and sludge. Therefore, the constructed wetlands’ (CWs) technology provides important advantages such as cost efficiency, relatively uncomplicated construction, operation and maintenance, relatively low energy consumption, tolerance for fluctuations of feed amount and quality, process stability, and high buffer capacity [3]. They surely can be constructed in decentralized, rural, and remote areas, where conventional wastewater treatment is not achievable [4]. It is also a great alternative for the treated wastewater production that can be reused for irrigation purposes and thus limiting the freshwater consumption [5]. The treatment of 1 m3 of domestic wastewater using CW systems costs between $186 and $357 (including investments and operational costs) as opposed to $571– 715 using conventional treatment techniques [4]. Planted filters or so-called CWs are systems designed and constructed to facilitate the treatment of wastewater, while using natural processes that involve wetland vegetation, soils, and their microbial assemblage [6]. The use of plants in CWs for the wastewater treatment was first experienced in Germany by Käthe Seidel in the

early 1950s. The first large-scale systems were put into operation service in the late 1960s, and since then, they have spread all over the world [7]. These CWs have been mainly used for the treatment of domestic and municipal wastewater [8], but they are now used for many other types of wastewaters, like it is the case for agricultural wastewater [9], industrial wastewater [10], and for the treatment of landfill leachates [11]. Horizontal subsurface flow constructed wetlands (HSSFCWs) are probably the most widely used type of CWs in the world [8]; (Fig. 1). In HSSFCWs, the influent is distributed over the width and height of the filter and flows slowly through the substrate in horizontal way, where wastewater meets aerobic, anoxic, and anaerobic areas (Fig. 1), and then it reaches the system outlet [7]. Indeed, aerobic zones are around the roots and rhizomes that release oxygen into the substratum [7]. In fact, the organic compounds are aerobically and anaerobically degraded by bacteria attached to the roots, the rhizomes, and to the middle surface.

2

Horizontal Subsurface Flow of Constructed Wetlands

The design of these systems has evolved around two main elements: the hydraulic load and the removal of pollutants. The main parameters evaluated are the surface area of the system, the hydraulic loading rate (HLR), and the hydraulic retention time (HRT).

2.1 The Surface Area According to the literature, there are various options for the design of HSSFCWs. The design is mainly based on minimizing the pollution load to produce effluents with the highest quality. Several mathematical models based on the kinetics of major pollutant elimination have been developed using semi-empirical coefficients based on existing stations data.

Horizontal Subsurface Flow Constructed Wetlands in Arid …

15

Fig. 1 Horizontal subsurface flow constructed wetlands [12]

Following the approach of [13], also called the k − C* model, which is based on the firstorder irreversible degradation kinetics and piston-type hydraulic flow, the area of the constructed wetland can be calculated according to the following formula (1):   Q Ci  C  A ¼  ln K Ce  C 

ð1Þ

A¼

Ci Ce K

Area of the CW (m2). Average flow rate (m3/d). Minimum concentration of the noted pollutant (mg/L). Concentration of the noted pollutant for water at the inlet of CW (mg/L). Concentration of the noted pollutant for water at the outlet of the CW (mg/L). Kinetic constant (m/d).

A Q Ci Ce KT y q KT

CW area (m2). Average flow rate (m3/d). Concentration of the noted pollutant for water at the inlet of the CW (mg/L). Concentration of the noted pollutant for water at the outlet of the CW (mg/L). Temperature-dependent first-order reaction rate constant (d−1). Average depth of the saturated zone of water (m). Porosity of the material used (%). Value is estimated using the following Eq. (3) [14]:

ðT20Þ

Another approach realized by [14] can be also followed, and the surface area is calculated from the formula below (2):

ð2Þ

with

with A Q C*

Q Ci  lnð Þ KT  y  q Ce

KT ¼ K20  h20

ð3Þ

16

M. Benbouzid et al.

2.2 The Hydraulic Loading Rate (HLR)

with T h20

K20 and h20

Operational temperature of the system (°C) Temperature coefficient for rate constant and K20: Rate constant at 20 °C. Values depend on the pollutants type met in surface and subsurface flow systems

HLR leads to evaluate wastewater level in the wetland. It avoids wastewater overflow risks that can contaminate in its turn nearby zones. Indeed, after the calculation of the surface area, HLR is calculated using the following formula (5): HLR ¼ V=A

ð5Þ

with Values for common pollutants are listed in Table 1, [14]: However, the simplest design model for domestic wastewater treatment is used in the UK. In this design, the surface area Ah is calculated using the following formula (4) [15]. Ah ¼ ðQd ðlnC0  lnCe ÞÞ=kBOD

ð4Þ

with Ah Qd C0 Ce kBOD

Surface area (m2). Average daily flow rate of wastewater (m3 day–1). Average concentration of BOD in the influent (mg/L). Average design BOD of the effluent (mg/L). Reaction rate constant, m/day.

kBOD = 0.06 and 0.31 for systems used for secondary wastewater treatment and tertiary treatment, respectively. Ah = 2.5 to 5.0 m2 per person, for secondary treatment when BOD concentration is ranging from 150 to 300 mg/L while 0.5 to 1 m2 per person for tertiary treatment [15]. Table 1 Temperature coefficients and rate constants in subsurface flow systems [14] Pollutant

h20

K20

Biochemical Oxygen Demand (BOD)

1.060

1.104

Nitrification

1.048

0.411

Denitrification

1.150

1.000

Pathogen removal

1.190

2.600

HLR V A

Hydraulic loading rate (cm/d). Volume of water per day (m3/d). Surface Area (m2).

However, the Water Pollution Control Federation (WPCF) (1990) suggests an HLR of 2 cm/d–20 cm/d for HSSFCWs, whereas authors [13] estimate an HLR between 8 cm/d and 30 cm/d.

2.3 The Hydraulic Retention Time (HRT) HRT is the time given to the wastewater to be percolated in the wetland substrate. It can be calculated with the following formula (6): HRT ¼ ðln Ce =Co Þ=KT

ð6Þ

with Co Ce

Concentration of the water at the inlet of the CW (mg/L). Concentration of water at the outlet of the CW (mg/L).

ITRC (2003) (Technical and Regulatory Guidance Document for Constructed Treatment Wetlands) suggests a retention time between 4 and 15 days, while Kadlec and Knight [13] propose a retention time between 2 and 4 days for HSSFCWs. However, too short retention time does not offer enough time for pollutant degradation, whereas prolonged retention time can cause stagnation [7].

Horizontal Subsurface Flow Constructed Wetlands in Arid …

2.4 The Depth of the Substrate To make sure to have a good water circulation, the bed depth should be equal to the plant roots’ maximum depth, which are generally 60 cm. However, since HSSFCWs are frequently small, there are no significant errors in the determination of the bed or water depth, and therefore, the ratio Vbed/Vnominal is near unity [3].

17

adsorption than coarse-grained soils do, but a higher rate of removal can be attributed to the higher cation exchange capacity in the soil. However, due to their low hydraulic conductivity, fine grains are no longer used for subsurface horizontal systems, and as a result, the adsorption capacity of commonly used media (gravel, gravillon) is severely constrained.

3.2 Phosphorus Reduction

3

Organic Matter and Nutrient Reduction Mechanisms in HSSFCWs

HSSFCWs meet the treatment objectives of reducing germs, total nitrogen, and phosphorus.

3.1 Reduction of Total Nitrogen Depending on its chemical form, nitrogen is one of the principal wastewater contaminants that can be destructive to aquatic life. Indeed, both organic and inorganic forms of nitrogen can be found in wastewater. Organic nitrogen may be present in amino acids, urea, uric acids, purines, and pyrimidines [13]. However, ammonium (NH4+), nitrite (NO2−), nitrate (NO3−), nitrous oxide (N2O), and dissolved elemental nitrogen or nitrogen gas (N2) are the several forms of inorganic nitrogen. Furthermore, nitrogen (N2), nitrous oxide (N2O), nitric oxide (NO2), and free ammonia are all components of gaseous nitrogen. In HSSFCWs, ammonification proceeds more quickly in the upper basin zone, where the environment is aerobic, and more slowly in the lower basin zone, where the environment is anaerobic [4]. The main mechanisms for the removal of nitrogen in HSSFCWs are nitrification/ denitrification reactions. However, it has been demonstrated that poor oxygenation of the rhizosphere in CW systems is the primary reason for the restricted nitrogen elimination. Plants absorptions play a much less significant role in the removal of nitrogen [16]. Limited volatilization occurs by the fact that these horizontal filters have no activity on the surface of the water [8]. Although fine-grained soils always remove nitrogen more effectively by

Phosphorus removal in HSSFCWs is limited by the fact that the media used for these systems (gravel, crushed stone) generally do not contain large amounts of Fe, Al, or Ca to facilitate precipitation and/or phosphorus sorption. In order to improve this efficiency, phosphorus removal materials that have a high capacity for sorption, such as steel slag or low weight clay aggregates, have been used recently [8]. Sani et al. [17] reported that there was no clear seasonal pattern for phosphorus removal and that removal of total phosphorus appeared to be temperature independent if contact time and loading rate varied, as the physical and non-dominant removal processes were primarily responsible for the removal of total phosphorus. In addition, vegetation can promote the right conditions for microbial activity by expanding the surface area in the water, oxygenating the area around the roots, and promoting resting conditions to facilitate filtration and sedimentation. Thus, vegetation role in phosphorus removal from wetlands is greater than the only proportion of physical removal [4]. In order to increase the processing capacity of a system and to optimize it, it is possible to make a system combining several constructed wetlands (combined constructed wetlands or hybrid constructed wetlands) [3].

3.3 Reduction of BOD5, Chemical Oxygen Demand (COD), and Total Suspended Solids (TSS) Constructed wetlands use biological (biomass attached to small media), chemical (adsorption,

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complexing, etc.), and physical (filtering) processes, which are all like natural processes occurring in natural wetlands, to reduce BOD5, COD, and TSS. These processes can occur at varying speeds depending on several factors like organic surface loading rate, water depth, etc. [18]. In fact, in HSSFCW, the filtration bed allows the penetration of plant roots in the entire bed and affords the whole bed oxygenation. Indeed, plants’ roots and rhizomes are hollow and have air-filled channels related to the atmosphere to bring oxygen in the root system. These aerobic conditions lead to the degradation of soluble organic matter due to the aerobic heterotrophic bacteria [19] resulting in the reduction of BOD5 and COD. According to the studies made by [20, 21], it can be noticed that BOD5 removal depends on the temperature and the values’ decrease in low temperatures’ period [12]. Same statement was done by [22] approving that the BOD5 removal efficiency was higher in summer. Concerning TSS removal in wetlands, it is mainly due to physical processes such as sedimentation and filtration, which is not related to temperature variations [21].

4

HSSCWs in Arid and Semi-Arid Climates

The aridity index (AI) is the ratio of mean annual precipitation to average annual evapotranspiration. When the value of AI in an area is less than 1, the latter is designated as an arid zone [23]. In such areas, proper management of water resources is highly necessary, and a proper wastewater treatment can generate an additional source of water. CWs in arid regions require specific knowledge because of the climate conditions [23]. Studies have recently been done under these climate conditions to treat domestic wastewater [23–25], municipal wastewater [4, 26], gray wastewater [18], industrial wastewater [27], and landfill leachates [11].

5

Parameters for HSSFCWs in Arid and Semi-Arid Climates

Table 2 summarizes the performances of study cases in arid and semi-arid areas, where HSSFCWs were used to treat different types of wastewaters. The HSSFCWs’ performances differ according to the different parameters that represent for the studies in a wetland.

5.1 Vegetation Types The choice of plant species is a parameter that has an influence on water loss and can be studied before considering the construction of the wetland in these areas. They truly play a role on the stabilization of the surface bed of wetlands, increase the porosity, absorb, and adsorb nutrients in wastewater. A lot of studies have shown the contribution of vegetation on wastewater treatment. Indeed, vegetation roots offer a massive surface area, where microbial growing occurs [4]. In fact, plant does simplify aerobic degradation especially when oxygen is released to the rhizosphere and the pollutants’ removal rate depends on the quantity of oxygen released [31]. Plants that have been used for wastewater treatment in arid and semi-arid climates include Echinochloa pyramidalis [27], Vetiveria zizanioides [11], Typha latifolia [18, 32], Phragmite australis (reeds) [4, 22, 23, 27], Kenaf (Hibiscus cannabinus L) [26], Canna and Cyperus papyrus [4]. The common reed (P. Australis) is still the most used plant for the treatment of wastewater in arid and semi-arid zones. Due to its physiological and morphological characteristics (high biomass production, high evapotranspiration capacity, resistance to high levels of salinity and pollution, high organic matter content, resistance to dry periods, and rapid growth of roots and rhizomes), the common reed is an excellent candidate for this type of treatment [33]. Moreover, when the HSSFCWs are composed of several types of vegetation (mixed culture), the performances are

Horizontal Subsurface Flow Constructed Wetlands in Arid …

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Table 2 Horizontal subsurface flow constructed wetland system cases in arid and semi-arid regions References

Country

Climate

Type of wastewater

Vegetation

HRT

Performances

Mandi et al. [28]

Morocco

Arid

Domestic

Phragmites Australis

3.5 h

COD = 71% TSS = 71% TKN = 55% TP = 30%

El Hamouri et al. [29]

Morocco

Semiarid

Municipal

Arundo Donax

Kouki et al. [25]

Tunisia

Semiarid

Domestic

P. Australis + Quennouille

3.6 days

COD = 89% BOD5 = 93% TSS = 98% TKN = 38% TP = 72%

Abou-Elela et al. [4]

Egypt

Arid

Municipal

Multi culture: Canna + P. Australis + Cyprus Papirus

11 days

COD = 91.5% BOD5 = 92.8% TSS = 92.3%

Laaffat et al. [18]

Morocco

Semiarid

Graywater

Tipha latifolia

–

COD > 85% BOD5 > 92% TSS > 94% TKN > 45% TP = 40%

Zidan et al. [24]

Egypt

Arid

Domestic

P. Australis

–

COD = 49% BOD5 = 49.4% TSS = 60.7%

Bakhshoodeh et al. [11]

Iran

Arid

Composted lixiviat

Vetiveria Zizanioides

5 days

COD = 53.7% BOD5 = 74.5 TSS = 73% NH3-N = 73.5% TP = 30%

Saggaï et al. [30]

Algeria

Arid

Domestic

Multi culture

5 days

COD = 80% BOD5 = 95% TSS = 96% TP = 60%

COD = 82% BOD5 = 82% TSS = 79% TKN = 11% N-NH4+ = 8% TP = 15% P-PO43− = 33% FC/100 mL = 1 in log unit

TKN Total Kjeldahl nitrogen; TP Total phosphorus; FC Fecal Coliforms

better than the performance of filters composed of monocultures (only one type of plant), as it is the case for [4] (mixed culture system) that reached a reduction of 91.5%, 92.8%, and 92.2% for COD, BOD5, and TSS, respectively. On the contrary to the study done with monoculture for the same type of wastewater and in the same climate conditions [24], the reduction of COD, BOD5, and

TSS reached 49%, 49.4%, and 60.7%, respectively. The explanation for this is given by the presence of numerous species that allows a more efficient spread of roots and a more favorable habitat, which favors the development of a great variety of microbial communities. In addition, root diversity delays the passage of wastewater into the system, which increases the HRT and,

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consequently, the removal efficiency. These results coincide with those found by [34], who assessed the performances of HSSFCWs planted with an assortment of species and HSSFCWs planted with a single type of vegetation. The performance of the horizontal filters using the common reed in arid climates revealed results equivalent to the performance of filters planted under different climate conditions, particularly for the study of [35] in humid climates, which had BOD5 abatement of 93%, COD abatement of 91%, TSS abatement of 95%, TN abatement of 67%, and TP abatement of 62%.

5.2 The Hydraulic Retention Time (HRT) Studies have shown that HRT has an influence on the performance of HSSFCWs [23, 28]. Indeed, it has been observed that the efficiency of the filter in arid climatic conditions increases when the HRT is higher. The best values were recorded for an HRT of 11 days [4] with an abatement of 88%, 91%, 91%, 42%, and 63% for COD, BOD5, TSS, TKN, and TP, respectively.

5.3 Evapotranspiration (ET) As the surfaces of the HSSFCWs are large relative to the volume of water, evapotranspiration is an important factor. It is the total loss of water through transpiration of plants and evaporation of surface water. In an arid country, the rate of ET is influenced by temperature, design parameters, and plant species used. According to the findings, the ET rate fluctuated from 35% in winter to a maximum of 60% during the hottest part of summer, when the temperature reached 40–45 °C [4].

5.4 The Substrate In HSSFCWs, the substrate is a path, through which wastewater can move, and a surface on which microorganisms can live. When wastewater passes through the pores between the media

M. Benbouzid et al.

particles, the microorganisms present in wastewater consume organic waste and remove it from the water. The choice of an appropriate substrate is very important, since substrate type may affect the plant growth and the removal of organic materials from wastewater [10]. Several substrate types were used, and a study done by Collac and Roston [24] concluded that the use of shredded tires as a substrate for an HSSFCW for treating domestic wastewater worked successfully. Other studies have concluded that a substrate consisting of plastic chips was used in HSSFCW for treating primary treated municipal wastewater [24] and achieved better purification performance than a substrate consisting of gravel with a diameter of less than 20 mm. This has been proven by studying the porosity of each of the beds. The reduction of porosities is related to the development of plant roots and the growth of biofilm on the surfaces of the support, in addition to the periodic accumulation of suspended solids. It was after then concluded that plastic media has a lower clogging ability [24]. Also, in another study done in arid climate conditions [4], good purification efficiencies (91% reduction of BOD5) were observed using coarse gravel with a diameter between 40 and 80 mm as substrate for the draining layer and fine gravel with diameter of 20 mm as filter layer. To conclude, several parameters are responsible for CWs efficiency, and the suitable substrate should permit oxygen transfer to carry organic matter reduction. It should also contain sorption sites that permit the sorption of pollutants like phosphorus and have good porosity to avoid clogging.

6

Comparison of CW Performances Under Different Climates

CWs used in tropical areas are significantly more efficient than systems used in non-tropical areas and can show almost 10 times higher removal rates of organic constituents and nutrients. In such climates, plants develop more quickly and year-round, and nutrient absorption may considerably increase nutrient removal [4]. Nevertheless, if the vegetations are not harvested, the

Horizontal Subsurface Flow Constructed Wetlands in Arid …

nutrients incorporated will be released during biomass decomposition [36]. However, in cold climate, CWs performance concerning soluble organic matter may be reduced since biological activity, which is mainly responsible for the reduction of organic matter, is highly depending on the temperature [12].

7

Conclusion

Constructed wetlands are becoming a suitable and affordable option for the treatment of various types of wastewaters in developing countries, since they are recognized as a reliable wastewater treatment technique. This review summarizes the performance of contaminant removal by HSSFCWs in arid and semi-arid climates. The results of various studies have shown a good reduction of BOD5, COD, TSS, and nutrients, when the HRT in the CWs is relatively high. The choice of the vegetation is a very important criterion to have good purification performance and planting different types of vegetations in the same CW shows better results.

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8. Vymazal, J., & Kröpfelová, L. (2008). Wastewater treatment in constructed wetlands with horizontal sub-surface flow (Vol. 14). Springer science & business media. 9. Vymazal, J. (2009). The use constructed wetlands with horizontal sub-surface flow for various types of wastewater. Ecological Engineering, 35(1), 1–17. https://doi.org/10.1016/j.ecoleng.2008.08.016 10. Calheiros, C. S. C., Rangel, A. O. S. S., & Castro, P. M. L. (2014). Constructed wetlands for tannery wastewater treatment in Portugal: Ten years of experience. International Journal of Phytoremediation, 16 (9), 859–870. https://doi.org/10.1080/15226514.2013. 798622 11. Bakhshoodeh, R., Alavi, N., Majlesi, M., & Paydary, P. (2017). Compost leachate treatment by a pilotscale subsurface horizontal flow constructed wetland. Ecological Engineering, 105, 7–14. https://doi.org/ 10.1016/j.ecoleng.2017.04.058 12. Wang, M., Zhang, D. Q., Dong, J. W., & Tan, S. K. (2017). Constructed wetlands for wastewater treatment in cold climate—A review. Journal of Environmental Sciences, 57, 293–311. https://doi.org/10. 1016/j.jes.2016.12.019 13. Kadlec, R. H., & Knight, R. L. (1996). Treatment wetlands. Lewis Publisher CRC press, Florida. 14. Reed, S. C., Crites, R. W., & Middlebrooks, E. J. (1995). Natural systems for waste management and treatment (2nd ed.). McGraw Hill. 15. Sundaravadivel, M., & Vigneswaran, S. (2001). Constructed wetlands for wastewater treatment. Critical Reviews in Environmental Science and Technology, 31(4), 351–409. 16. Vymazal, J. (2007). Removal of nutrients in various types of constructed wetlands. Science of The Total Environment, 380(1–3), 48–65. https://doi.org/10. 1016/j.scitotenv.2006.09.014 17. Sani, A., Scholz, M., & Bouillon, L. (2013). Seasonal assessment of experimental vertical-flow constructed wetlands treating domestic wastewater. Bioresource Technology, 147, 585–596. https://doi.org/10.1016/j. biortech.2013.08.076 18. Laaffat, J., Ouazzani, N., & Mandi, L. (2015). The evaluation of potential purification of a horizontal subsurface flow constructed wetland treating greywater in semi-arid environment. Process Safety and Environmental Protection, 95, 86–92. 19. Vymazal, J. (2005). Horizontal sub-surface flow and hybrid constructed wetlands systems for wastewater treatment. Ecological Engineering, 25(5), 478–490. https://doi.org/10.1016/j.ecoleng.2005.07.010 20. Vanier, S. M., & Dahab, M. F. (2001). Start-Up performance of a subsurface-flow constructed wetland for domestic wastewater treatment. Environmental Technology, 22(5), 587–596. https://doi.org/ 10.1080/09593332208618260 21. Kadlec, R. H. (2003). Pond and wetland treatment. Water Science and Technology, 48(5), 1–8. https:// doi.org/10.2166/wst.2003.0266

22 22. Garfí, M., Pedescoll, A., Bécares, E., Hijosa-Valsero, M., Sidrach-Cardona, R., & García, J. (2012). Effect of climatic conditions, season and wastewater quality on contaminant removal efficiency of two experimental constructed wetlands in different regions of Spain. Science of The Total Environment, 437, 61– 67. https://doi.org/10.1016/j.scitotenv.2012.07.087 23. Vera, I., Verdejo, N., Chávez, W., Jorquera, C., & Olave, J. (2016). Influence of hydraulic retention time and plant species on performance of mesocosm subsurface constructed wetlands during municipal wastewater treatment in super-arid areas. Journal of Environmental Science and Health, Part A, 51(2), 105–113. https://doi.org/10.1080/10934529.2015. 1087732 24. Zidan, A. R. A., El-Gamal, M. M., Rashed, A. A., & El-Hady Eid, M. A. A. (2015). Wastewater treatment in horizontal subsurface flow constructed wetlands using different media (setup stage). Water Science, 29(1), 26–35. https://doi.org/10.1016/j.wsj.2015.02. 003 25. Kouki, S., M’hiri, F., Saidi, N., Belaïd, S., & Hassen, A. (2009). Performances of a constructed wetland treating domestic wastewaters during a macrophytes life cycle. Desalination, 246(1–3), 452–467.https:// doi.org/10.1016/j.desal.2008.03.067 26. Albalawneh, A., Chang, T. K., Chou, C. S., & Naoum, S. (2016). Efficiency of a horizontal subsurface flow constructed wetland treatment system in an arid area. Water, 8(2), 51. https://doi.org/10.3390/ w8020051 27. Fonkou, T., Fonteh, M. F., Djousse Kanouo, M., & Akoa, A. (2010). Performances des filtres plantes de Echinochloa pyramidalis dans l’épuration des eaux usées de distillerie en Afrique subsaharienne. Tropicultura, 69–76. 28. Mandi, L., Bouhoum, K., & Ouazzani, N. (1998). Application of constructed wetlands for domestic wastewater treatment in an arid climate. Water Science and Technology, 38(1), 379–387. https:// doi.org/10.1016/S0273-1223(98)80004-8 29. El Hamouri, B., Nazih, J., & Lahjouj, J. (2007). Subsurface-horizontal flow constructed wetland for

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Anaerobic Dynamic Membrane Bioreactors for the Domestic Wastewater Treatment Eyüp Debik, Neslihan Manav-Demir, Esra Celik, and Aslican Ihtiyaroglu

a layer made of using a low-cost and porous support material (e.g., mesh, woven, and nonwoven fabric) that serves as a primary filter of biological granules. Thus, DM technology offers a cost-effective operation that can be combined with various anaerobic processes such as up-flow anaerobic sludge bed (UASB) or conventional anaerobic digesters. In this review, important factors affecting the performance of anaerobic dynamic membrane bioreactor processes are examined with respect to applicability of this technology in domestic wastewater treatment and reuse.

Abstract

Global population growth, urbanization, and industrialization have increased water demand and the need of treatment and reuse of wastewater. Conventional processes such as activated sludge and trickling filters are used for the treatment of domestic wastewater. These processes can be combined with membrane technology to generate water for reuse. Membrane technologies that are implemented in biological treatment processes typically include membrane bioreactors (MBR) and anaerobic membrane bioreactor (AnMBR) systems. Although MBR/AnMBR systems can produce high-quality effluents, their practical use may be limited due to disadvantages such as the condensation of the cake layer and potential clogs in the MBR/AnMBR system. As a result, capital investment as well as operation and maintenance of MBR systems tends to be costly for municipalities worldwide. As an alternative, dynamic membrane (DM) technology has been developed to offer economical and operational advantages over MBR/AnMBR. Dynamic membrane (cake) is

E. Debik (&)
N. Manav-Demir
E. Celik
A. Ihtiyaroglu Environmental Engineering Department, Yildiz Technical University, Davutpasa Campus, Esenler, Istanbul, Turkey e-mail: [email protected]

Keywords




Domestic wastewater treatment Anaerobic treatment Dynamic membrane technology Anaerobic dynamic membrane bioreactor (AnDMBR)




1




Introduction

Domestic wastewater is the primary wastewater type that is produced and collected by municipalities in large quantities, whereas industrial wastewater is typically produced by single sources such as a factory. Both types of wastewaters must be treated to protect environmental resources such as existing surface water bodies. Recent studies focus on the recovery of domestic wastewater as a resource within the scope of sustainability [1–4]. A key suggestion

© The Author(s), under exclusive license to Springer Nature Switzerland AG 2023 E. Debik et al. (eds.), Wastewater Management and Technologies, Water and Wastewater Management, https://doi.org/10.1007/978-3-031-36298-9_3

23

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by these studies is to consider domestic wastewater as an alternative source of water, food, and energy. Likewise, sustainability studies highlight the need for energy-efficient wastewater treatment technologies that will reduce energy consumption and carbon emissions of wastewater treatment plants. Hence, these new perspectives underscore the need to use advanced treatment processes instead of (or in combination with) the conventional treatment systems. Conventional aerobic treatment systems such as activated sludge are widely used for the treatment of domestic wastewater. Common disadvantages of such systems include their high energy requirement and sludge formation in large amounts. The use of sustainable technologies such as anaerobic treatment and wetlands is increasing to reduce energy consumption and carbon emissions by wastewater treatment plants. In this regard, anaerobic treatment processes, which are considered as sustainable methods of domestic wastewater treatment, have attracted the attention of researchers due to their significant advantages such as low energy consumption, low sludge formation, and bioenergy recovery via methane production [5–7]. However, slow growth rates of anaerobic microorganisms and lengthy start-up times are among the key disadvantages that limit the application of anaerobic treatment processes for common use [8, 9]. Using conventional treatment technologies in combination with membrane filtration processes offers new benefits and eliminates some of the disadvantages of conventional systems. For instance, when anaerobic membrane bioreactor (AnMBR) is used in combinations with conventional anaerobic treatment systems, hydraulic retention time (HRT) and sludge age (SRT) can be controlled independently of each other in wastewater treatment [10–12]. By doing so, slow-growing anaerobic microorganisms remain in the bioreactor via membrane filtration, and thereby large biomass concentrations can be maintained in the system to ensure effectiveness of the treatment technology. In addition to those advantages obtained because of more flexible HRT and SRT in AnMBR systems, studies have

E. Debik et al.

been also conducted for domestic wastewater treatment at low temperatures and effective process performances observed [13]. In parallel to domestic wastewater treatment, AnMBR systems have been successfully applied to treat industrial wastewaters [14–18]. In AnMBR systems, the advantages of anaerobic treatment processes are combined with the advantages of membrane filtration technology to achieve high-quality effluents as well as high removal of organic matter. Simply, AnMBR provides favorable conditions for extended biodegradation of organic matter in the treatment of both domestic and industrial wastewaters. In contrast to these advantages, AnMBR systems have also several disadvantages that limit their applicability. A cake layer forms on the membrane surface because of accumulation and condensation of various components such as sludge particles, biopolymers (soluble microbial products (SMP) and extracellular polymeric substances (EPS)), and inorganic colloids in wastewater. This cake layer on the membrane surface/pore causes significant operational problems [19, 20]. The cake layer causes resistance to filtration and accounts for more than 80% of the total filtration resistance at AnMBR systems. Hence, the cake layer resistance is the main cause of membrane fouling [21–23] that eventually leads to membrane clogging. A clogged membrane is the key problem that adversely affects treatment efficiency of AnMBR systems [24]. Similarly, other adverse effects of membrane clogging include low membrane flux, high filtration resistance, high energy consumption for fouling control, and relatively high maintenance cost. Such problems constitute important challenges for widespread implementation of AnMBR systems [25–27]. The dynamic membrane (DM) technology has been developed as an alternative method to solve the problems mentioned above. Dynamic membrane is the cake layer that forms on the porous support material that covers the actual conventional membranes. Hence, dynamic membrane application turns the cake layer formation into an advantage in terms of filtration in AnMBR processes [28–30]. The dynamic membrane can be

Anaerobic Dynamic Membrane Bioreactors for the Domestic …

cleaned easily and reformed on the support material in a short time [29]. DM can be viewed as the secondary membrane layer due to its ability to hold various types of pollutants such as colloidal substances, microbial products, and inorganic substances and eliminates the need of conventional membrane systems such as ultrafiltration (UF) and microfiltration (MF) [31]. This is related to the fact that the dense and compact DM layer performs more filtration function than the underlying membrane [32]. For the formation of the DM layer, cheap and simple designed mesh, woven fabric, and nonwoven fabric filter materials can be used as support materials instead of expensive conventional membranes [28, 33]. Thus, thanks to the DM technology, high costs of filtration membranes can be avoided in AnMBR systems. By combining the DM application with anaerobic treatment processes, systems defined as anaerobic dynamic membrane bioreactor (AnDMBR) have been created. AnDMBR processes have been widely researched since 2010 and can be considered as emerging technologies. AnDMBR processes have been successfully used for the treatment of domestic wastewater [34–38], synthetic wastewater [29, 31, 39], textile wastewater [40, 41], leachate [42], and solid waste such as waste sludge and food waste after pre-treatment like homogenization [43–46]. In studies with AnDMBR systems, 60–99% chemical oxygen demand (COD) and 90–100% suspended solids (SS) removal can be achieved [34, 47–49]. However, the removal of nutrients such as total nitrogen (TN) and total phosphorus (TP) with AnDMBRs occurs just at low level [29]. As a result, low organic content but nutrientrich effluents are obtained by using AnDMBR

25

systems. AnDMBR effluents with high nutrient content can be reused in agricultural irrigation, or it can be given to receiving water bodies after treatment with a nutrient removal method that will warrant to meet the discharge standards. In this chapter, authors reviewed the dynamic membrane (DM) technology and the applicability of anaerobic dynamic membrane bioreactor processes (AnMBR) in domestic wastewater treatment.

2

Dynamic Membrane (DM) Technology

There are two types of DMs according to their formation: self-forming and pre-coated [28, 50]. In self-forming DMs, the membrane layer is formed during filtration of solid substances that are in wastewater. Such solids include microbiological granules, inorganic solids such as sand and clay, and organic solids such as proteins, fat, and grease, all of which constituting of colloidal particulate matter (CPM). The DM layer is formed by the accumulation of CPM on the porous support material (Fig. 1). In pre-coated DMs, the membrane layer is formed by substances that are added to wastewater externally. Pre-coated DMs can be obtained by coating the porous support material with one or more colloidal materials. The main disadvantage of pre-coated DMs is the need for external material such as powdered activated carbon (PAC), kaolin, MnO2, and TiO2 [28]. In AnDMBR systems, self-forming DMs are mostly applied by forming a cake layer of colloidal materials and sludge flocs on the support material [32, 38, 51]. The selection of a suitable

Fig. 1 Schematic demonstration of the dynamic membrane layer

26

support material that provides an effective DM layer formation is very important [52]. In AnDMBR applications for wastewater treatment, mesh (metal and/or polymeric), woven fabric, and nonwoven fabrics are generally used as support materials [30, 35, 53]. Mesh and woven fabric materials are more suitable than nonwoven fabric materials for DM formation and filtration performance. Moreover, the DM layer cleaning on the nonwoven fabric support material is difficult physically due to the structure and pore distribution [28, 54]. However, the selection of support material with an appropriate pore size is important [33, 52]. Pore size of the support material directly affects DM layer formation. Pore size of the support materials is larger than that of conventional membranes. Considering that the wastewater quality is low due to the colloidal and particulate matter passing through the support material pores at the beginning of the filtration, the support material should have a suitable pore size in order to form a fast DM layer. In this direction, Kiso et al. [55] investigated the treatment and filtration performance of synthetic wastewater with nylon mesh support material with different pore sizes (100, 200, 500 lm). They concluded that the nylon mesh with a pore size of 100 lm provided a more effective filtration than the other sizes. In AnDMBR studies, support materials with a pore size of 10–200 lm typically are employed for wastewater treatment [36, 39, 51, 52]. Since the pore size affects the formation period and filtration flux of the DM layer, the selection of an appropriate pore size depends on the particle size of the materials (sludge flocs, etc.) forming the DM layer to provide a good filtration performance [28]. Selection of the pore size of the support material properly provides an effective filtration performance, high filtration fluxes, and rapid DM layer formation. DM layer formation is a complex process involving microbiological and physicochemical effects initiated by the accumulation of sludge flocs, inorganics, and biopolymers (EPS and SMP) on the support material. It consists of two layers, a gel layer mostly formed by EPS, and a cake (DM) layer formed by sludge flocs [30, 50].

E. Debik et al.

Initially, a gel layer is formed by the accumulation of EPS on the support material. The gel layer provides the opportunity for sludge flocks to bind and accumulate, thus forming a porous and compacted DM (cake) layer based on the sludge. Zhang et al. [30] emphasized that the accumulation of biopolymers on the support material is important for the formation of an effective DM layer. Few minutes after the start of the filtration process, a DM layer starts to form on the support material. This early-stage DM layer cannot provide effective filtration, while the filtrate quality increases in time. In particular, several hours and/or days will be required to establish an effective DM layer. Ersahin et al. [29] reported that they reached a stable DM layer after 10–20 days in their study with the synthetic wastewater having high organic content. When the DM layer becomes excessively dense and thick, it reaches high filtration resistances and causes a decrease of flux efficiency [34, 42]. At this stage, the DM layer needs to be cleaned easily and practically by physical means (biogas spraying, backwashing, brushing, etc.). Since the cake layer constitutes most of the total filtration resistance, physical cleaning is sufficient for DM cleaning [50]. DM regenerates on the support material in much shorter time during new DM formations after a layer cleaning. This can be explained by the fact that the gel layer, consisting mostly of EPS, tightly bound to the support material surface, offers a suitable medium for the rapid accumulation of sludge flocs for the formation of the DM (cake) layer [30].

3

Anaerobic Dynamic Membrane Bioreactor (AnDMBR)

In recent years, various laboratory-scale studies have been carried out that integrated the DM technology and anaerobic membrane bioreactors. The studies typically focus on how to use DM technology in continuous-stirred tank reactors (CSTR) and up-flow anaerobic sludge blanket (UASB) reactors [54, 56]; (Fig. 2). In this direction, Fig. 2 shows basic membrane configurations as submerged and side-stream that are

Anaerobic Dynamic Membrane Bioreactors for the Domestic …

used in AnDMBR systems. Submerged configurations are two types, one of which is internal submerged membrane unit that is immersed in the bioreactor, while the external (side-stream) submerged membrane unit is immersed in a different compartment outside of the bioreactor [57]. The submerged configuration operates at lower flux with higher permeability and is often used in medium–large scale, such as municipal wastewater treatment plants due to relatively low membrane and operation costs per volume treated. The side-stream configurations require a circulation pump and are often used in industrial applications, where the cost per unit time is usually small [58, 59]. The submerged membrane configuration has been preferred by many studies due to reasons such as cost and simple operation. However, Ersahin et al. [60] compared the application of submerged versus side-stream configurations and concluded that both configurations provided high COD removal efficiencies along with high organic matter removal (over 99%) [60]. While the submerged membrane

27

configuration achieved more biogas production, the side-stream configuration achieved a lower total filtration resistance. The findings highlighted that the submerged AnDMBR process offered high permeate quality and higher methane production than side-stream configurations. Thus, this configuration was more advantageous than the AnDMBRs with side-stream configuration. There is a clear need to further analyze the performances between the typical configurations [56]. Studies are needed to identify a suitable membrane module for the support material. Currently, there are two types of membrane modules as flat sheet and tubular employed by AnDMBR processes. Flat sheet modules are widely used due to their simple configuration during production and operation [29, 38, 61]. The key issues that limit the widespread use of AnMBR processes include investment cost, membrane fouling, low filtration flux, and high energy required for fouling control, maintenance, and operational costs. Hence, dynamic

Fig. 2 Different configurations of AnDMBR: a CSTR internal submerged, b CSTR external submerged, c CSTR sidestream cross flow, d UASB internal submerged, e UASB external submerged

28

membrane (DM) technology has been developed as an alternative method to cope with these issues. It has been observed that higher filtration fluxes and lower filtration resistances are obtained in AnDMBR systems with the porous support material used for the formation of the DM layer [54, 56]. The support material used for the formation of the DM layer is more economical than conventional membranes, and it reduces the investment and operation costs. Considering the low initial investment cost of support material, the cost of membrane modules in AnDMBR systems is considerably lower than AnMBRs. In addition, membrane cleaning is easier in AnDMBR systems, and the maintenance cost are, therefore, much less. In summary, AnDMBRs have proven advantages over AnMBRs in terms of membrane module costs, membrane flux, membrane cleaning, and energy requirement. AnDMBR processes have been developed as an alternative method to solve the problems encountered in AnMBR processes and continue to be used because of its effective performance. Anaerobic baffled reactor (ABR), which is another type of anaerobic reactor, is a reactor type that eliminates various problems such as clogging and sludge bed expansion that limit the usage conditions compared to other anaerobic reactors (such as UASB) [62]. In the ABR, suspended baffles are installed inside the reactor to divide the reactor into compartments and to direct the wastewater flow up-and-down from one compartment to the next [63, 64]. High sludge concentrations are retained through substrate/biomass contact in the upstream region (active zone) of each chamber, resulting in high treatment performance and low sludge generation [65, 66]. Because ABR is simpler in design than other anaerobic process types, it offers key advantages such as being more resistant to hydraulic shock loads, operating at longer sludge ages, low sludge generation, high treatment efficiency, and low investment and operating costs [67]. The most obvious advantage of ABR is its ability to separate the acidogenic and methanogenic phases longitudinally within the reactor. Thus, it is possible for the bacterial communities

E. Debik et al.

in the two stages to develop under the most effective conditions [64]. ABR is considered a suitable wastewater treatment process for lowincome communities due to its low energy requirement and simple design [68]. It is known that ABR provides an effective treatment of domestic wastewater. Thus, this reactor type is preferred for treatment of wastewater with a range of organic loads (0.04–2.4 kgBOD/(m3.d) or 0.07–2.62 kgCOD/(m3.d) [69, 70]. Zwain et al. [67] emphasized that ABR is a suitable anaerobic process for wastewater treatment with low organic content. However, they stated that most of the organic pollution is removed in the first compartments of the reactor, thus reducing the number of compartments in the reactor. Anaerobic baffled reactor (ABR) was applied in combination with conventional membrane processes in numerous studies [71–74]. The researchers investigated using the dynamic membrane in the ABR system (anaerobic baffled dynamic membrane bioreactor-AnBDMBR). As a result, a new treatment method having both economic and easy operational features has been developed.

4

Anaerobic Dynamic Membrane Bioreactor Processes for Domestic Wastewater Treatment

Treatment and filtration performances of some studies, in which anaerobic dynamic membrane bioreactors (AnDMBR) were applied for the domestic wastewater treatment, and the values of general design and operation parameters are listed in Table 1. An et al. [75] investigated the municipal wastewater treatment using an AnMBR with an active volume of 12.9 L and average HRT of 2.6 h at room temperature and achieved 70.1% of COD removal with an AnDMBR system using a nonwoven fabric as a support material. This study is one of the first AnDMBR studies applied for the municipal wastewater treatment and has enlightened other studies. In another study [30], an AnDMBR system was operated at 65 L/m2.h (LMH) filtration flux at psychrophilic

Flat sheet

Flat sheet

Flat sheet

Flat sheet

UASB/side-stream

CSTR/submerged

UASB/submerged

UASB/submerged

Nylon mesh

Nylon mesh

Polyester nonwoven

Polyethylene terephthalate

Polyamide/nylon woven mesh

Dacron mesh

LMH L/m .h; TMP transmembrane pressure

2

Flat sheet

CSTR/side-stream

Dacron mesh

Flat sheet

Flat sheet

UASB/submerged

UASB/side-stream

Dacron mesh

Flat sheet

UASB/submerged

Polyethylene terephthalate

Tubular

UASB/submerged

Support material

Module Config.

Reactor type/membrane configuration

3.6

20–25

22–25

25

15 3.5

25–30

20–24

6.9

0.898

10–30 –

45

10–15

15–20

Temp. (oC)

–

45

12.9

Vol (L)

Table 1 AnDMBR applications for the domestic wastewater treatment

75

75

30

28–46

200

61

61

61

0.64

Pore Size (µm)

272

292– 5151027

361

375

900

413

298

302

259

Influent COD (mg/L)

1–8

8

22.5

35

30

– 25–10

–

0.02

0.02

0.04

0.005

0.004

–

63.4

–

60–77

74–93

90

64–71

90

81.6

57.3

– 0.081

70.1

COD removal (%) 0.98

Filtration Area (m2)

Up to 35

Up to 25

Up to 25

Up to 30

TMP (kPa)

40

1.4–28

60

65

65

5

Flux (LMH)

3.6

0.25–5.7

2.2

8

8

2.6

HRT (h)

[38]

[37]

[61]

[36]

[39]

[35]

[34]

[30]

[75]

Refs.

Anaerobic Dynamic Membrane Bioreactors for the Domestic … 29

30

temperatures (10–15 °C) for approximately 100 days using an UASB reactor with an effective volume of 45 L and an HRT of 8 h to treat the municipal wastewater. While the COD removal was 56.6 ± 8.26% until day 20, this value increased to 57.7 ± 4.6% after the formation of the dynamic membrane layer. However, dissolved COD could not be effectively removed, because the formation of the dynamic membrane layer was limited in the AnDMBR system. In the study, when transmembrane pressure (TMP) increased, the DM layer was physically cleaned to regain membrane permeability, and then the system restarted consecutively. In another study by Zhang et al. [34], municipal wastewater was treated by using an AnDMBR system for 330 days. During this period, temperature of the wastewater that entered the AnDMBR system ranged between 10 and 30 °C, representing seasonal variations. Since the system was operated at low temperatures during the first 100 days of operation, COD removal efficiency was lower than the overall removal efficiency achieved for the whole study. COD removal efficiency of the system was recorded as 63.4% on average throughout the operation. When the TMP reached 25 kPa, the DM layer was physically cleaned with tap water. Ma et al. [35] investigated an AnDMBR system with a short HRT of 2.2 h for a long operation period (300 days). Organic matter removal efficiency was recorded as 81.6% on average throughout the operation. In another study, a side-stream AnDMBR system combined with a CSTR was used to treat synthetic domestic wastewater [39]. The AnDMBR system was operated at different HRTs and filtration fluxes, applied in the range of 1.4–28 LMH during the 121 days of operation period. During the operation, the COD removal efficiency was approximately 90%. Quek et al. [36] investigated the applicability of the side-stream DM filtration technology with a UASB type reactor. They investigated the effects of different HRTs (3, 6 h), different fluxes (30, 60, 100 LMH), and the pore size of the support material (28 and 46 lm) on the

E. Debik et al.

operational performance of the AnDMBR system. The increase of filtration flux caused an increase in the membrane fouling rate. Similarly, it has been found that clogging rate occurred four times faster for the smaller pore size. COD removal efficiencies were observed as 71% and 64% at 3 and 6 h of HRTs, respectively. This study showed that the AnDMBR system operated at 6 h of HRT provided more effective organic matter removal than 3 h of HRT. In another study, Sun et al. [61] used an AnDMBR system operated at ambient temperatures (25 °C) for the treatment of synthetic domestic wastewater. They found that the COD removal efficiency was approximately 90%. When the flux fell below 10 LMH, the operation was stopped and the operation continued after washing DM. Hu et al. [37] operated two parallel AnDMBR systems, which had a sludge recirculation and non-recirculation at psychrophilic temperatures and determined that sludge recirculation affected the DM filtration performance adversely. However, a COD removal efficiency of 70–95% was achieved in both systems. In another recent study, the performance of an AnDMBR system was investigated for 93 days at different HRTs (1–8 h) by varying the filtration flux in four stages (22.5–180 LMH) [38]. At 8, 4, and 2 h HRTs, COD removal efficiencies were observed as 74.4%, 77.3%, and 70.6%, respectively. When HRT was reduced to one hour, COD removal efficiency was recorded as 60.4%, and a decrease in organic matter removal efficiency was evident. The key reason for this outcome was the inability of anaerobic microorganisms to perform their functions at very short HRTs such as 1 h. According to the studies on the domestic wastewater treatment indicated in Table 1, selection of various parameters such as temperature, HRT, flux, type of support material, and pore size is very important to optimize the operation of AnDMBR processes. Moreover, many advantages for DM technology were also proven for the domestic wastewater treatment using AnDMBR systems and obtained effective performances.

Anaerobic Dynamic Membrane Bioreactors for the Domestic …

5

Conclusion

Considering the many proven advantages of dynamic membrane technology, AnDMBR systems are applied for the treatment of domestic wastewater and show effective treatment performance. However, studies on AnDMBR systems are still needed with particular emphasis on how to use DM technology with different reactor configurations. Likewise, their performance for the domestic wastewater treatment should be evaluated carefully considering that protection of water resources and environmental resources are important issues. Further studies must be conducted using different types of reactor configurations. A key goal for these studies is to develop an economic, efficient, and sustainable treatment of domestic wastewater to protect water resources. Thus, these new treatment systems will find widespread use and application worldwide. Acknowledgements This study is originated from the project of FBA-2021-4556 funded by Yıldız Technical University Scientific Research Projects Coordination Unit.

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Agro-Industrial Wastewater Management—Case Studies of Wastewaters from the Olive Industry and Pig Farming— Advantages of Anaerobic Co-Digestion for Small Units in Remote Areas Georgios Pilidis and Ioannis Zarkadas Abstract

Agriculture is one of the major contributors to the resource security of most countries. Unfortunately, agriculture combined with food manufacturing and processing is generating enormous quantities of waste and wastewaters. In the industrialized countries, traditional type agro-activities have been partially replaced with modern industrial facilities exploiting new and, in some cases, advanced wastewater treatment processes. This manuscript describes several economically viable and environmentally acceptable technologies able to treat different waste/wastewater such as cattle manure, pig manure, table olive wastewater, and slaughterhouse waste. While the concept of the bio-refinery is not new, a bio-refinery based on a digestion system can become a successful operation through the

G. Pilidis (&) Department of Biological Applications and Technology, University of Ioannina, 45110 Ioannina, Greece e-mail: [email protected] I. Zarkadas Polyeco SA, Waste Management and Valorization Industry, 19300 Aspropyrgos Attica, Greece

recovery of biogas that can be used for electricity generation as well as for hot water to heat up the digester. While the digestate can be used as an organic fertilizer before or after composting, with the application of advanced technologies high value products can be generated improving the overall cash flow of the system. Additionally, in this work it presented a cash flow comparison between a digestion system operating with cattle manure as the only substrate and one with a mixture of different wastes. When considering the economics of the digestion of cattle manure as single substrate, the high transportation cost of waste with low bio-methane potential is rendering the system unsustainable. But, if the co-digestion opportunity is considered, the system has the potential to become a successful operation. Here, the capital and operational expenditures of a phenol oxidation facility are also presented for the treatment of the table olive wastewater. This system can reduce the phenol content by 85.8% with a total operational cost of 2.1 Euro/m3 rendering the wastewater suitable for further treatment and valorization. Keywords










Agro-industry Co-digestion Olives Pigs Wastewaters

© The Author(s), under exclusive license to Springer Nature Switzerland AG 2023 E. Debik et al. (eds.), Wastewater Management and Technologies, Water and Wastewater Management, https://doi.org/10.1007/978-3-031-36298-9_4




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G. Pilidis and I. Zarkadas

Acronyms

ACD AD ANPHOS BOD CAP CNG COD CSTR EU FOG HRT IPPC LNG MENA MEs OMW SEs SME SOPs TOC TOPW UASB VS VAT TS WAS

1

Anaerobic co-digestion Anaerobic digester Phosphorous removal in anaerobic effluents Biological oxygen demand Common agricultural policy Compressed natural gas Chemical oxygen demand Continuous stirred tank reactor European Union Fats, oil, greases Hydraulic retention time International Plant Protection Convention Liquefied natural gas Middle East North African Micro-enterprise Olive oil mill wastewaters Small enterprise Small medium enterprise Standard operation procedures Total organic carbon Table olive processing wastewaters Up-flow anaerobic sludge bed reactor Volatile solids Valued-added tax Total solids Waste activated sludge

Introduction

Agriculture is the main contributor to the resource security of most countries and one of the economic sectors with the highest waste biomass production. Driven by the increasing population, demand for food and food products has been steadily increasing and resulting in enormous amounts of wastewater generated by the food industries. In the industrialized

countries, traditional types of agro-activities have been partially replaced by modern and intense processes that are exploiting new and, in some cases, advanced wastewater treatment technologies. Nevertheless, frequently the environmental problems are not solved, since most of the farmers are, at least in the less developed countries, very small industrial units based on seasonal operation that are not able to apply sustainable and economical viable treatment methods. The current 15 million SMEs including MEs and SEs engaged in agro-industrial sectors throughout Europe and MENA countries notice that the EUs CAP has a number of limitations, such as budgetary constraints and the need to respond to the economic crisis and climate change. In the recent years, the CAP has become more market orientated, and interventions in agricultural markets throughout Europe have become less significant in budgetary terms than before. An indication of this is the lack of market commodity forecasting that leaves the farmers economically vulnerable to competition [1]. It is well documented that in the agroindustrial sector the scientific community was focused on the integrated waste management of bigger industries, which were able to co-finance investments and to cover operational costs. For this reason, the bigger industries in the USA and the EU countries, under the pressure of different ISO systems and certification regimes and regulations, adapted available treatment methods for the management of their wastes and wastewaters. Unfortunately, in less developed countries, most of the agro-industrial activities are, according to EU definition, micro-enterprises, facilities with less than ten employers, without scientific support and/or specific turnover. These companies are mostly spread over areas close to a city or village. Usually, they are operating and producing agro-industrial products on the traditional way with limited access to scientific information and technical progress. Conventional management methods for organic waste treatment such as incineration, composting, or landfilling require significant amounts of energy and have a significant environmental footprint. On the contrary, the anaerobic digestion, in which

Agro-Industrial Wastewater Management—Case Studies …

organic wastes are converted into biofuels, bioenergy, and bio-fertilizers, has become very important over the past two decades, since the digestion process exhibits many benefits and advantages compared to the usually applied treatment methods. For those micro-sized companies, the waste management requires cooperation to split the initial costs between different users as well as to achieve the required waste volume to operate the system. This can be achieved in two ways: • Relocation to an industrial park and • Establishment of a region servicing centralized bio-refinery. The most critical point for any technology that will be used on a sustainable way to protect the environment (soil, water, and air) is its business plan, and how over the years amortization and operational costs will be covered by sustaining a balance between expenses and income. For the industrialized countries, but also for less developed countries, many programmes exist, where financial instruments can cover part of the construction. However, the problems occur after the initial operation and when the interest wears out. This period coincides also with the highest financial contribution and the need for scientific support. The image of abandoned facilities is not uncommon in many countries resulting in loss of money and continued environmental degradation.

2

Industrial Parks

Industrial parks are land zones allocated closed to transport facilities and outside the main residential area of a city, clustering together different heavy industrial activities of a region. It is somehow an “intense” version of a business park, operating under strong regulations that should be accepted and followed by all company members of the park. By concentrating dedicated infrastructures in a limited area, the per-business cost of that infrastructure (transport facilities, high-power electric supplies, communication

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cables, large-volume water supplies, gas lines, waste/wastewater treatment facilities, etc.) is reduced. Among other regulations, the management of the waste/wastewaters resulting from each individual activity is the most important one. In some industrial parks, the effluents from each activity are accepted from the central management system after a separate pre-treatment step, which is applied by the wastewater producer. The central wastewater treatment plant accepts effluents with specific characteristics. For Greece, the following maximum concentrations for common environmental parameters are required for their acceptance (in ppm): BOD5 500; COD 1200; phenols 5; detergents 50; heavy metals 10; suspended solids 500. After treatment in the central plant, the effluents are discharged to surface water bodies (river, lake, coastline) if they are compatible with the EU regulations. Otherwise, they end up in the municipal wastewater treatment plant of the nearby city.

3

Anaerobic Co-Digestion

The anaerobic co-digestion (ACD) concept is not new, but it is a well-proven waste/wastewater treatment technology usually applied in Northern Europe. The anaerobic digestion is considered as an economically viable and environmentally compatible method for the utilization/ valorization of agriculture derived waste biomass. This waste biomass can be converted into bio-energy and bio-chemicals and helps the countries and communities to minimize their environmental footprint [2]. The process employs a consortium of different micro-organisms to convert the organic matter present in wastewater into biogas, which contains 55–70% methane [3]. Biogas can be used as a fuel source usually in internal combustion engines or in micro-turbines coupled to electric generators. The produced energy could be then directly injected into the grid. As a result, the produced electricity is CO2 neutral, while the operators can benefit from the financial

38

incentives offered, through grants, tax reduction, and low interest loans for energy production from renewable sources. Despite recent advancement in ACD processes, the effects of feedstock compositions and operating conditions on the bio-methane production processes are still in progress. The anaerobic digestion has been steadily developing during the last years with more than 3000 systems operating across Europe treating different waste/wastewaters. The cost of the generated energy depends on the size and type of the system employed and can be competitive to the conventional energy generation with additional advantages due to CO2 neutral emissions, protection of natural resources, and overall protection of the environment. Some of the advantages are • Developing a direct reduction of methane emissions from stored waste/wastewater, • Reduction the emissions of odorous compounds, • Enrichment of soil with nitrogen and phosphorus when the digestate is land applied as organic fertilizer, • Cost reduction or financial profit, since more than 80% of the calorific value of biogas can be converted into electricity and heat in a combined heat and power unit. The electricity can be directly injected into the grid and substitutes energy generation from conventional sources, and • Minimal energy requirements especially when compared to aerobic treatment systems. While the anaerobic digestion has many advantages, it has also several disadvantages, which must be addressed in the design of any treatment system. • The organic load and the nutrients available in the wastes/wastewaters are only partially removed, and thus, a land application is still necessary [4], • Storage space (lagoon) is required for the treated wastewater until final land application, which is a problem especially during plant

G. Pilidis and I. Zarkadas

growth and before harvest (more than 5 months/year), and • The physicochemical characteristics of the influent substrate should be monitored and adjusted when necessary. Other process limitations are • Inhibitory phenomena due to high ammonia levels already present and further produced during the hydrolysis stage of the high nitrogen content of wastes, including the pig/chicken manures and slaughterhouse wastes [5], • Inhibition due to high lipids content especially for waste/wastewaters of animal origin [6, 7], • Problems due to the alkaline pH of raw wastes [8], • Inhibition due to phenols present in olive mill wastewaters [9, 10], • Low availability of waste streams, which challenges the economic viability of the system, • Current treatment techniques focus on the recovery of biogas from wastewaters through anaerobic digestion, while little or no attention is paid to the final disposal of the digestate due to low or no economic value of the product, and • During the periods, where land disposal is not feasible, holding tanks are employed as storage units, increasing the operation costs.

4

The Concept of a Central Bio-Refinery

Bio-refining [1] has become a relevant technology over the past decades, used to produce a wide range of products, including biofuels, energy, and chemical. The objective of a biorefinery is to optimize the use of resources and to minimize waste production, thereby maximizing benefits and profitability. Several bio-refining methods are currently applied, including fermentation, trans-esterification, fast pyrolysis, gasification, hydrogenation, and anaerobic digestion.

Agro-Industrial Wastewater Management—Case Studies …

Production in a modern bio-refinery usually takes place in a single chain and not in a multistage–multi-product framework, resulting in a limited number of products, a higher production of wastes, and increased environmental footprint. The goal for the micro-sized enterprises is a full utilization of biomass from different activities with targeting production of value-added products and low environmental footprint. Bio-refining for the recovery of added value products and raw materials has been developed mainly with substrates rich in cellulose, semicellulose, and lignin aiming at the recovery of a limited number of products, with the most important being ethanol that can be readily upgraded to biofuel [11–13]. Current bio-refining systems present several limitations, which must be addressed in order to provide future systems with further possibilities in attaining their aims. As presented in [14], the limitations can be divided into technical and nontechnical ones. Among the technical barriers are harvesting, transportation, application of the best available technology, reliability of the system, and dependence upon oil prices in case of biofuels. The non-technical barriers include ecological and environmental deterioration, competition with available food sources, and lack of public awareness and skilled workforce. The lack of multiple products derived from today’s bio-refineries combined with the dependence upon a single raw material, and the possibility of waste generation makes bio-refineries too expensive to be utilized in substrates, where their characteristics are not stable throughout the processing cycle, or the required mass is not secured. Agricultural micro-enterprises lack the scale needed to apply bio-refining in a cost-effective manner on their own waste streams, although they produce biomass, which can be used as feedstock in bio-refining. Economy of scale can be achieved, when biomass is harvested from agro-enterprises in a radius of 20 km from a centralized bio-refinery [15]. ACD for different type of agricultural wastewaters at the heart of a bio-refinery could

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be an excellent option to solve the environmental problems of agricultural micro-enterprise activities in less developed and/or remote areas of countries worldwide. Through this option for the benefits are (a) A bio-refinery works throughout the year; problems related to seasonal nature of some agro-industrial wastes are solved, (b) Critical mass of waste, required for sustainability, can be guaranteed, (c) Unbearable costs for the micro-enterprises can be drastically reduced, (d) Compliance with international and national regulations also for MEs is ensured, (e) Protection of the environment (soil, water, air) is guaranteed, (f) New job opportunities will be created, and (g) Contributions to the local economy and creation of new knowledge are done. ACD processes applied on agricultural and food wastes together with activated sludge from wastewater treatment plants as substrates are an effective approach to create bio-energy. Some important parameters related to the composition of the waste/wastewater types, the pre-treatment of some polymeric materials, and operating conditions such as the C/N ratio of feed substrates at the condition of the maximum methane production should be studied in advance. The average methane production from the ACD of food wastes with other organics, such as agricultural wastes and WAS, was found to be 421 ± 45 m3CH4/ton VSadded [2] significantly higher compared to manures. Priorities for future research directions should be • Establishment of a regional circular centre utilizing agricultural wastes, • Development of optimization strategies for an anaerobic co-digestion system, including substrate pre-treatment, and system configuration and control, and • Maximization of economic benefits of digestate utilization. Studies on anaerobic co-digestion of for the region of Epirus (Northwestern Greece) typical

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agricultural waste/wastewaters have been performed in the Laboratory of Environmental Technology at the University of Ioannina with the following results: • By using table olive processing wastewaters with cattle and pig manure, the mesophilic codigestion resulted in a 50% increase of the methane production based on the volatile solids added, accompanied by an almost 30% phenol reduction and 80% removal of total organic carbon, while under thermophilic conditions a 7% increase of the TOC removal efficiency was achieved [8]. • Thermophilic co-digestion of pasteurized food waste in mixtures with cattle manure as feedstock for anaerobic digesters resulted in methane yields of 281–385 m3CH4/ton VSadded for the different mixtures at organic loading rates up to 6.85 kgVS/m3d with the TS content of the influent reaching up to 15.7% [16]. • The co-digestion of different types of fur farming and feed wastes offered specific methane productions ranging between 368 and 591 m3CH4/ton VSadded, corresponding to 67.4 and 91.1% of their theoretical methane potential [17]. • The thermophilic co-digestion of three pickling and canning semi-solid wastes as substrates gave a 32% improvement compared to the methane yield achieved under monodigestion conditions [18]. The construction and operation of a centralized bio-refinery, operated by an independent company or cooperative, will provide an integrated solution to the waste management of micro-sized agro-industries, which are widespread in a remote area. For the prefecture of Ioannina in Northwestern Greece, the following calculations have been made based on experiments in pilot scale digesters for two types of agro-industrial wastes for a bio-refinery of the same size: • One mixture of (a) 60% cow manure, (b) 20% pig manure, (c) 12.5% table olive processing

G. Pilidis and I. Zarkadas

wastewater, and (d) 7.5% slaughterhouse wastes (Table 1) and • One pure waste type (cow manure). The profit of 205,040 € refers to the net profit during the loan repayment period. This will be increased to 427,810 € after payment of the loan. On the contrary, if only one type of agrowastes will be used for the anaerobic digester (cow manure), the financial balance of the biorefinery will have the following picture: Income

439,400

Expenses

322,500

Transport costs

210,000

Operational costs

87,500

Maintenance

25,000

Lending Profit

139,230 − 22,330

The mono-digestion of manure cannot provide a profit until the loan is paid. Obviously, the low bio-methane production of the waste together with the high transportation costs for the waste collection minimizes the viability of the unit.

5

Anaerobic Co-Digestion and Added Value Products

Anaerobic digestion (AD) processes produce energy in the form of biogas, which contains up to 70% methane that can be used to power electrical generators reducing by this way the treatment plant’s energy bills. To increase the methane production, anaerobic digesters are heated to produce an environment suitable for the micro-organisms that digest the substrate, which can include manure, food scraps, fats, oils, greases (FOG), industrial organic waste, and sewage sludge. The biogas, a by-product of digestion, rises to the top of the digester and is then collected. It can fuel engines to produce mechanical power, heat, or electricity that can fuel furnaces, boilers, and digesters, as

Agro-Industrial Wastewater Management—Case Studies …

41

Table 1 Financial balance for a bio-refinery based on anaerobic co-digestion for a mixture of four agricultural waste types Parameter

Centralized treatment unit

Mono-digestion of manure

Substrate quantities per day

100 t (four loads per day  25 t)

100 t (four loads per day  25 t)

Digester type

CSTR

CSTR

Wastewater treatment per year

36,500 t

36,500 t

Hydraulic retention time

21 days

21 days

Digester operation–thermophilic zone

55 oC

55 oC

Methane, calorific value

9.8 kWh/m3 (efficiency 100%)

9.8 kWh/m3 (efficiency 100%)

Electricity sale price

200–220 €/MWh (https://www. dapeep.gr/eidikiti-timi-agoras-eta-022022-diorthotiki/)

200–220 €/MWh (https://www. dapeep.gr/eidikiti-timi-agoras-eta02-2022-diorthotiki/)

Sale price of treated waste as geoimprover

Zero € but with the obligation of farmers to collect and dispose it of

Zero € but with the obligation of farmers to collect and dispose it of

Cost for substrates

Zero (0)

Zero (0)

Gate fee for slaughterhouse wastes (today price payable for incineration 5–10 € cents/kg)

40 €/t

No slaughterhouse waste will be added

Gate fee for table olive wastewaters (today payable price 2.3 €/m3)

5 €/m3

No table olive wastewater will be added

Organic load rate (OLR)

3.9–4 kg VS/m3d 3

3.9–4 kg VS/m3d

Digester volume Vinx HRT

2100 m volume needed  1.2 = 2500 m3 (400 m3 for future activities or for longer HRT)

2100 m3 volume needed  1.2 = 2500 m3 (400 m3 for future activities or for longer HRT)

Mixture of agricultural wastewaters

60% cow manure, 20% pig manure, 12.5% table olive wastewater, 7.5% slaughterhouse wastes

100% cow manure

Energy consumption for heating the digester

Zero (0): use of warm water from power generator

Zero (0): use of warm water from power generator

Energy consumption for pasteurization

Zero (0): use of warm water from power generator

Zero (0): use of warm water from power generator

Selling price of thermal energy

Zero (0): heating the digester and the pasteurization system

Zero (0): heating the digester and the pasteurization system

Distance of waste generator to the treatment unit

20 km

20 km

Construction costs

5000 €/kWh installed. 400  5000 = 2,000,000 €

5000 €/kWh installed. 250  5000 = 1,250,000 €

Income from the energy sale per year

2350 m3/d = 383 kWh = 738,000 €

1400 m3/d = 228 kWh = 439,400 €

Income from gate fee per year

21,900 € for table olive wastewaters and 109,460 € for slaughterhouse wastes

No other incomes

Total income per year

869,360 €

439,400 €

Income

(continued)

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G. Pilidis and I. Zarkadas

Table 1 (continued) Parameter

Centralized treatment unit

Mono-digestion of manure

Installation costs 5000 € / kWh installed

2,000,000 €

1,250,000 €

Transport costs 100 € per hour

1 ½ hours per load 25 m3 (150 €) and 4 h per load (400 €)

1 ½ hours per load 25 m3 (150 €)

Cow manure 21,900 m3/25

876  150 = 131,400 €

1400  150 = 210,000

Slaughterhouse wastes 2,736 t/25

109  150 = 16,350 €

Table olive wastewaters 4,380 m3/ 25

175  400 = 70,000 €

Expenses

Pig manures 7,300 m3/25

292  75 = 43,800 €

Operational costs per year

140,000 € per year (7% of the installation costs)

87,500 € per year (7% of the installation costs)

Maintenance costs

40,000 € (2% of the installation costs)

25,000 € (2% of the installation costs)

Total operational costs per year

441,550 €

322,500 €

Lending sum

2,000,000 €

1.250.000 €

Interest rate

7%

7%

Years of repayment

15

15

Annual payment to cover loan

222,770 €

139,230 €

205,040 €

− 22,330 €

Loan

Financial results Profit/damages per year until payback of the loan

Financial balance of the bio-refinery (in €) Income

869,360

Expenses

441,550

Transport costs

261,550

Operational costs

140,000

Maintenance

40,000

Lending

222,770

Profit

205,040

well as power vehicles, and supply homes and business through the natural gas pipeline. To increase its value, biogas can be stripped from the carbon dioxide, water, and other contaminants. Lower-grade biogas is generally used in internal combustion engines, while higher grades are reserved for more efficient and sensi-

tive engines. More processing can upgrade biogas to compressed natural gas (CNG) or liquefied natural gas (LNG) to fuel vehicles. Digestate, the organic bio-solid left over from ACD sewage treatment, is valuable in agriculture, where it is used to amend soil. Nitrogen can be extracted from it to make more concentrated

Agro-Industrial Wastewater Management—Case Studies …

fertilizer. It can also be treated, dewatered, and used as bedding for livestock or even upcycled into biodegradable flowerpots. Co-digestion of food waste and sewage sludge provides benefits in terms of anaerobic process stability enhancing the buffer capacity of ammonia (for example) and biogas formation, which can be increased up to 80% when compared with mono-digestion [19]. Current situation: Duque-Acevedo [20] analysed the literature published from 1931 to 2018 and found that countries with high income and big agricultural industries have made remarkable advances in the implementation of bio-economic policies. This means that the deployment of ACD processes exhibits indispensable importance and significance. For MEs in remote areas of less developed countries, a bio-refinery can financially survive, when the variety of by-products normally generated in the downstream processing of biomass in a bio-refinery could be increased. A major route to do this is through high rate anaerobic digestion. However, a common problem is that highly loaded digesters become unbalanced by overloading or by the input of feedstocks rich in ammonia, salts, proteins, and fats. This causes the digestion process to slow down in its conversion rate and efficiency. When this occurs, the operator has only a few options to remediate the digestion system, which often has a volume of several thousand m3. Some actions that can be undertaken against those problems are • Reducing the organic loading rate until the digester accelerates again, • Replacing a substantial volume of the mixed liquor (often more than 50%) with mixed liquor from a well-functioning digester operating under similar conditions, or • Adding some additives such as iron sulphate or calcium hydroxide for controlling the inhibiting factors. These approaches are time consuming and costly and can lead to the overall downstream process component becoming disconnected from

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the other production processes. At present, there is no effective procedure or add-on technology available to remediate full-scale systems, which have become unbalanced, and which allows the overall bio-refinery to continue processing at the normal rate. Beyond the state of the art: Bio-refining based on a combination of physicochemical, enzymatic and microbiological treatments such as liquefaction and ACD, permits the production of high value products and energy. A plant processing 100,000 tons of bio-waste annually could produce biogas with a value of approximately 5.5 M €. The holistic concept should be that all solid, liquid, and gaseous materials at a bio-refinery site are potentially useful resources, even those traditionally considered as waste. To clarify it further, the possibility to modify agricultural waste materials and co-products (e.g. straw, peat residues, manure, slurry, sawdust, bark, and other litter materials) to biopolymers and bio-char is the most promising challenge. Using these novel products for wastewater treatment and further in biogas production enables an efficient nutrient recycling. The modification can include purification of materials, cross-linking, increasing porosity and surface area, making them more capable in binding the pollutants. Manure and slurry are wastes from agricultural animal husbandry (pigs, cattle, and poultry), which contain significant amounts of nitrogen and phosphorus that could not only be recycled as direct application in the field, but also in a stabilized form to minimize nutrient losses to the atmosphere (as ammonia), to groundwater (as nitrate) or in run-off (as phosphates). For this purpose, the residues should be separated into a wet solid fraction, a permeate fraction, and the residual water. Depending on the separation technology, the solid fraction must be additionally dried to facilitate the subsequent carbonization step. After modification, biopolymers are further treated in a wastewater treatment plant. The sludge generated in the plant can be transported to the anaerobic digester (AD) plant. The digestate is processed and stabilized to be used as fertilizer, i.e. substrate, and, after that,

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transported to the field. All phases (processing, transportation) are influencing the CO2 footprint. The pre-dried solid fraction can be the feedstock for a pyrolysis process. This carbonization process produces a manure-based bio-char, a carbon-rich material with higher plant nutrient contents than pure vegetation-based bio-char. Many studies have evaluated the use of bio-char as adsorbents for the removal of, e.g. heavy metals in wastewater. For example, anthropogenic emissions of ammonia need to be mitigated due to negative environmental impacts and economic losses. Ammonia and ammonium ions are the most encountered nitrogenous compounds in wastewaters. In addition to eutrophication, they also cause undesired odours and several diseases. Based on literature, bio-char’s capacity to take up ammonia is studied; however, research on nutrient removal, particularly ammonium and phosphate from water by biochar, is limited. The philosophy behind this is to stimulate the bio-refinery industry to deliver new products to end user industries. As example, the following processes could lead to a series of new products: • Production of value fatty acids from the hydrolysis process [21], which can be used as a building block for bio-polyesters, • Production of bio-sulphuric acid from the H2S from the hydrolysis process, • Production of CO2 from the hydrolysis process, which will be used in the ammonium carbonate production and/or in the bioreactor, • Production of high value succinic acid as a building block in biopolymers, • Capture of NH4+ with CO2 from the biorefinery to produce ammonium carbonate, which is a foaming agent for biopolymers, • Purification and upgrading of CH4 as biogas in the succinic acid reactor, • Production of fibres from the digestate, • Production of struvite from the liquid digestate, and • Use of the bio-refinery effluent as a forward enhancement to revitalize the anaerobic digester and to prevent clogging.

G. Pilidis and I. Zarkadas

6

The Case of Table Olive Processing Wastewater

For several decades, the scientific community focused on the management of olive oil mill wastewaters. Table olive processing wastewaters (TOPW) from the treatment either of black or of green olives remained in the background. In the beginning of the 1990s, the first serious papers on the management of table olive processing wastewaters were published [22–25]. In relation to the olive oil mill wastewaters (OMW), the effluents from TOPW have much lower BOD5 and COD concentrations [10]. Since table olive production is highly important for the economies of the Mediterranean countries (Italy, Spain, Greece, Turkey, Tunisia), constituting one of the major agro-industrial sectors, it was necessary to work on a financially manageable wastewater system. The technologies applied for the treatment of TOPW are physicochemical, chemical, and biological as well as combinations of them. Membrane filtration (ultra-filtration, microfiltration, nano-filtration, reverse osmosis) offers simple solutions and, therefore, is widely used in industrial wastewater treatment. But not always are membrane techniques the most suitable ones for any kind of wastes; limitations and some disadvantages connected with the use of filters are • Turbid water cannot be used in membrane filtration, • Glass filters can break, • Filters have high repair and installation costs, • Filters require high differential pressure, and • Clogging can easily occur. Advanced oxidation processes are treatments, where strong oxidation agents are applied (e.g. H2O2 or O3 or hydroxyl radicals), which are formed in situ by UV radiation. Ozone and hydroxyl radicals can decompose the phenolic compounds and to reduce the alkalinity of TOPW making them more acceptable to biological treatment [26]. By the Fenton reaction, Fe

Agro-Industrial Wastewater Management—Case Studies …

(II) breaks hydrogen peroxide into hydroxyl radicals that decompose a number of organics to smaller molecules. The reaction mechanism is not yet known, since due to the very short life of the intermediate products, it was not possible to document the reactions experimentally, and as such it is not known if the aqueous or the organic iron species are the real oxidation reagents. The first manuscripts on the application of Fenton’s reagent presented not very promising results [27]. The most spectacular results, however, (COD and phenol reduction of 65 and 92%, respectively) were achieved with the application of a combination of O3/UV/H2O2 [28]. One of the first attempts in the management of TOPW was the application of activated sludge from municipal wastewater treatment plants under aerobic conditions [28]. Even if a BOD5 reduction of up to 85% took place, no reduction of the phenolic compounds was achieved. Some years later, the application of a specialized strain of Aspergillus niger, isolated from older TOPW, was very successful in the degradation of phenolic compounds [25]. Due the relatively low COD and BOD5 concentrations, anaerobic biological treatment is not the best available technology for TOPW treatment [24]. To minimize inhibitory factors (high pH, high concentration of sodium chloride, and phenols) and to improve biogas production up to 350 m3CH4/ton VSadded, TOPW have been used as a co-substrate to pig and dairy manure [8]. Biological methods for wastewater treatment offer cost-effective and efficient processes; however, these technologies have the disadvantage of slow response and inability to decompose macromolecules and toxic organic compounds. TOPW consists of effluents from three washing stages with decreasing organic load. They belong to the “light” industrial wastewaters since the average COD concentrations from the three consecutive washing procedures are up to three times higher than the ones of municipal wastewaters. This leads to the idea that the problem of this wastewater can be solved in biological treatment plants, like the activated sludge systems.

45

In an aerobic biological treatment (12 L, upflow air lift fermenter; HRT 48 h) using a strain of Aspergillus niger isolated from undiluted TOPW, COD, total phenolic compounds, and some simple phenols were reduced by 85.8%, 58%, and 85%, respectively, while residual COD and phenolic compounds can be eliminated using a second advanced oxidation process with Fenton’s reagent (H2O2 and FeSO4  7 H2O) [9]. These laboratory and pilot scale results had to be confirmed in a large-scale treatment unit. As such, a table olive processing facility in central Greece was selected for the construction and operation of a 100 m3/day TOPW (Fig. 1). This was a small medium enterprise (SME) producing 5000 tons/year of green olives. During the operation over a period of three months (August to October), the viability of the plant was calculated taking into consideration operational, construction, and amortization costs (Table 2).

7

The Case of Anaerobic CoDigestion Processes in the Livestock Sector

Economic development in lower income countries is closely connected to the livestock sector and specifically to pig, cattle, and chicken breeding. Annually, these operations produce huge amounts of waste/wastewaters resulting in extensive pollution of soils, groundwater, and surface water bodies, because of high COD, nitrate, phosphate, and phenol as well as pathogen releases. Pig manure-based anaerobic digesters are usually presenting relatively low conversion efficiencies and biogas production, usually due to the influent streams, failing to meet the minimum required characteristics. While this has limited consequences at small digesters, in centralized systems, where post-treatment is applied, it could have catastrophic effects and result in complete failure of the system. Recently, a methodology was developed for choosing proper technological solutions for pig farms with due account for the whole variety of combinations of production and natural and

46

G. Pilidis and I. Zarkadas

Fig. 1 Treatment of 100 m3/day TOPW: from top left in circular direction— homogenization tank, acidification in three stages, biological aeration, sedimentation, Fenton oxidation tank, sedimentation, and pH correction

Table 2 Viability of a treatment plant for a table olive processing SME taking into account amortization, construction, and operational costs (all financial figures in €) Annual production of table olives

5000 tons/year

1. Construction cost

300,600

2. Amortization costs/year

33,400

On a 10-year basis

3. Inoculums cost

1600

Aspergillus niger strain A

4. Chemicals

10,020

Fe(II)sulphate; sulfuric acid; flocculation chemicals; CaO

5. Chemical and microbiological analyses

2000

6. Energy costs

5010

Variable

7. Operational costs

5010

200 man months technician

8. Maintenance costs

3340

9. Sludge management

2800

Total costs (2–9)

63,250

Wastewater volume/year in m3

22,500

Wastewater treatment cost per m3

2.10

climatic conditions. It includes an accurate accounting for nitrogen and phosphorus flows in the manure processing and utilization chain and considers the limiting factors of the farm and the end-product customer requirements [29]. It is important that anaerobic digestion as a biological treatment method can only be applied to wastes/wastewaters meeting several criteria. More importantly, the pH of the incoming stream must be between 6 and 8, the carbon to nitrogen ratio (C/N ratio) is in the range of 20–30: 1, the

Remarks

Transport to landfill (28€/ton) Circulating the effluent from the third to the first washing stage

ammonia levels must not exceed 1 g/L, and the total phenols levels must be less than 2 g/L [30]. The PIGMAN project [31] was the first initiative that provided a complete, integrated, technological solution to the treatment and recovery of nutrients from pig manure. The project that was centred to pig manure developed an anaerobic digestion technology (UASB) for over 50% COD reduction, followed by the ANPHOS process for the removal of phosphate and ammonia as struvite. Artificial wetland

Agro-Industrial Wastewater Management—Case Studies …

technology was a cost-effective final polishing stage for wastewater treatment after the ANPHOS process. The recovered solids from the anaerobic digestion processes as well from the ANPHOS treatment (struvite) were stabilized aerobically to produce compost, which can be used as fertilizer for soils poor in nitrogen and phosphorous. Anaerobic digestion produced biogas (73 m3 of biogas per pig, per year) which can be used on the farms and the balance traded. Additionally, a struvite fertilizer and compost mixture (about 75 kg dry matter with 2.5 kg P and 2.7 kg N per pig, per year) is generated, which has a better fertilizing value than untreated pig manure. Unfortunately, in low-income countries, there is the lack of critical mass, which could convert a treatment system into an economically viable operation. A solution for overcoming the problem of critical mass is the combined treatment of livestock with other agricultural wastewaters in a centralized system. This method not only provides the feed stock, but it will also provide the necessary balancing of the physicochemical characteristics of the influent streams resulting in lower operational cost and increased biogas yields for the same volume system. Obviously, combining waste is an economically viable solution that helps farmers and polluting operations to meet the IPPC criteria [32]. Therefore, in a modification of the PIGMAN project, co-digestion experiments in a CST reactor with four livestock waste types (slaughterhouse, pig manure, dairy cattle manure, poultry manure) were performed. The substrates were selected after preliminary analyses of possible feed stocks. The selection was based mainly on available quantities, seasonality, and physicochemical characteristics. Addition of the latter to the proposed system is not only desirable due to its abundance during production season, but it also provides pH correction and increases the C/N ratio of the other co-substrates. The first stage of the process comprises a mixing tank and a CST reactor. The incoming stream from the livestock operations, the co-

47

substrates, and the OMW for pH and C/N correction after mixing are introduced at the side of the tank, where they encounter the microorganisms thriving within the reactor. At this stage, approximately 50% of the organic load of the incoming waste mixture is converted into biogas. HRT as well as organic loading rate are two important operational parameters. Overloading the system due to short HRT or to higher organic loading rates should be avoided. Otherwise, it is possible that sub-optional growth of the micro-organisms as well as failure of the system due to accumulation of intermediate process products including volatile fatty acids is possible. The effluent of the CST reactor is transferred to a liquid–solid separator, from where the solids are transferred to the composting unit. The liquid will be forwarded to the ANPHOS process [33], where phosphates and ammonium precipitate as struvite (a magnesium ammonium phosphate) fertilizer [34]. The reaction is taking the following form: Mg2 þ þ NH4 þ þ PO3 4 þ 6H2 O ! Mg(NH4 )PO4  6H2 O: The effluent of the CST reactor and the ANPHOS processes are introduced to artificial wetland units for further nitrogen removal [35– 37]. The generated biomass at the artificial wetland harvested once a year (e.g. in December) can be used as a co-substrate for the anaerobic digester (Fig. 2). As the final stage of the process, all solids collected during the liquid–solid separation together with the precipitated solids derived from the ANPHOS process will be forwarded to the composting unit. The addition of the high phosphorus and ammonia solids from the ANPHOS process to the composting unit might not be necessary as the produced struvite is a wellknown fertilizer with similar fertilization value to the chemical fertilizers and has a commercial value. The production of high-quality irrigation water in the artificial wetlands provides some financial income to the user.

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G. Pilidis and I. Zarkadas

Fig. 2 Flow sheet of an anaerobic co-digestion process

8

Conclusion

Many MEs and SEs mainly in Southern Europe, the MENA countries and in poor, less developed regions without scientific experienced personnel and/or specific turnover have difficulties to comply with the standard operation procedures that have been set by the EU, USA, and other industrialized countries. These companies are mostly spread over areas close to a city or village and are producing agro-industrial products on traditional ways with limited access to scientific information and progress. Anaerobic co-digestion of different agricultural wastes was intensively studied over the past years. Combination of different wastewaters

addresses the issue of required critical mass of feedstock for an economically viable treatment system. Furthermore, co-digestion and yearround operation of the system due to diverse substrates are more efficient in biogas production tilting the balance between profit against construction and operation costs. Now, these small companies have several options to follow between doing nothing and continuing to degrade the environment, apply a simple method for the management of their wastes, relocate in an industrial park where the operation must follow the instructions of the operator of the park, or become part of a centralized management system (bio-refinery). In this review, two case studies were presented: one of an already applied relatively cheap

Agro-Industrial Wastewater Management—Case Studies …

solution for the table olive processing wastewaters in a medium-sized enterprise and a study based on laboratory experiments for some mixtures of agricultural and livestock wastes that can be the basis for the development of a central biorefinery in a remote area. The research field is open for further investigations, especially in the compatibility of different wastes/wastewaters from the agricultural and livestock sector towards higher production of biogas, extraction of more added value products, and treatment of more than one type of wastes.

References 1. European Commission, 2015. EIP-AGRI Workshop “Opportunities for Agriculture and Forestry in the Circular Economy”. Workshop Report 28–29 October, Brussels, Belgium. https://ec.europa.eu/eip/ agriculture/sites/agri-eip/files/eip-agri_ws_circular_ economy_final_report_2015_en.pdf 2. Pan, S. Y., Tsai, C. Y., Liu, C. W., Wang, S. W., Kim, H., & Fan, C. (2021). Anaerobic co-digestion of agricultural wastes toward circular bioeconomy. IScience, 24(7), 1. 3. Rasi, S., Veijanen, A., & Rintala, J. (2007). Trace compounds of biogas from different biogas production plants. Energy, 32, 1375. 4. Walker, L., Charles, W., & Cord-Ruwisch, R. (2009). Comparison of static, in-vessel composting of MSW with thermophilic anaerobic digestion and combinations of the two processes. Bioresource Technology, 100, 3799. 5. Sung, S., & Liu, T. (2003). Ammonia inhibition on thermophilic anaerobic digestion. Chemosphere, 53, 43. 6. Palatsi, J., Viñas, M., Guivernau, M., Fernandez, B., & Flotats, X. (2011). Anaerobic digestion of slaughterhouse waste: Main process limitations and microbial community interactions. Bioresource Technology, 102, 2219. 7. Hejnfelt, A., & Angelidaki, I. (2009). Anaerobic digestion of slaughterhouse by-products. Biomass and Bioenergy, 33, 1046–1054. 8. Zarkadas, I., & Pilidis, G. (2011). Anaerobic codigestion of table olive debittering and washing effluent, cattle manure and pig manure in batch and high volume laboratory anaerobic digesters: Effect of temperature. Bioresource Technology, 102, 4995. 9. Kyriacou, A., Lasaridi, K., Kotsou, M., Balis, C., & Pilidis, G. (2005). Combined bioremediation and advanced oxidation of green table olive processing wastewater. Process Biochemistry, 40, 1401.

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10. Parinos, C., Stalikas, C., & Pilidis, G. (2007). Chemical and physico-chemical profile of wastewaters produced from the different stages of Spanishstyle green olives processing. Journal of Hazardous Materials, 145, 339. 11. Moshkelani, M., Marinova, M., Perrier, M., & Paris, J. (2012). The forest biorefinery and its implementation in the pulp and paper industry. Applied Thermal Engineering, 50, 1427. 12. Huang, H., Ramaswany, S., Tschirner, U., & Ramarao, B. (2008). A review of separation technologies in current and future biorefineries. Separation and Purification Technology, 62, 1. 13. Fitzpatrick, M., Champagne, P., Cunningham, M., & Whitney, R. (2010). A biorefinery processing perspective: Treatment of lignocellulosic materials for the production of value-added products. Bioresource Technology, 101(8915), 2010. 14. Demirbas, M. F. (2009). Biorefineries for biofuel upgrading A critical review. Applied Energy, 86, 151. 15. Zarkadas I., 2012. Anaerobic digestion of agroindustrial wastes in the region of Epirus. PhD Thesis, University of Ioannina, GR, 179. 16. Zarkadas, I., Sofikiti, A., Voudrias, E., & Pilidis, G. (2015). Thermophilic anaerobic digestion of pasteurized food wastes and dairy cattle manure in batch and large volume laboratory digesters: Focusing on mixing ratios. Renewable Energy, 80, 432. 17. Zarkadas, I., Dontis, G., Pilidis, G., & Sarigiannis, D. (2016). Exploring the potential of fur farming wastes and byproducts as substrates to anaerobic digestion process. Renewable Energy, 96, 1063. 18. Zarkadas, I., Georgopoulos, N., Kaldis, F., Pilidis, G., & Sarigiannis, D. (2017). Assessing the Biomethane potential of three pickling and canning semi-solid wastes under thermophilic conditions. Fresenius Environmental Bulletin, 26, 39. 19. Morales-Polo, C., Cledera-Castro, M. D. M., & Moratilla Soria, B. Y. (2018). Reviewing the anaerobic digestion of food waste: From waste generation and anaerobic process to its perspectives. Applied Sciences, 8(10), 1804. 20. Duque-Acevedo, M., Belmonte-Urena, L. J., CortesGarcía, F. J., & Camacho-Ferres, F. (2020). Agricultural waste: Review of the evolution, approaches and perspectives on alternative use. Global Ecology and Conservation, 22, 1. 21. Patel, A., Mahboubi, A., SárváriHorváth, I., Taherzadeh, M. J., Rova, U., Christakopoulos, P., & Mastakas, L. (2021). Volatile Fatty Acids (VFAs) generated by anaerobic digestion serve as feedstock for freshwater and marine oleaginous microorganisms to produce biodiesel and added-value compounds. Frontiers Microbiology, 12, 135. 22. Beltran-Heredia, J., Torregrosa, J., Dominguez, J. R., & Garcia, J. (2000). Aerobic biological treatment of black olive washing wastewaters: Effect of an ozonation stage. Process Biochemistry, 35, 1183.

50 23. Rivas, F. J., Beltrán, F. J., Alvarez, P., Frades, J., & Gimeno, O. (2000). Joint aerobic biodegradation of wastewater from table olive manufacturing industries and urban wastewater. Bioprocess Engineering, 23, 283. 24. Aggelis, G., Ehaliotis, C., Nerud, F., Stoychev, I., Lyberatos, G., & Zervakis, G. I. (2000). Evaluation of white-rot fungi for detoxification and decolorization of effluents from the green olive debittering process. Applied Microbiology Biotechnology, 59, 353. 25. Kotsou, M., Kyriacou, A., Lasaridi, K., & Pilidis, G. (2004). Integrated aerobic biological treatment and chemical oxidation with Fenton’s reagent for the processing of green table olive wastewater. Process Biochemistry, 39, 1653. 26. Ayed, L., Assrs, N., Chammem, N., Othman, N. B., & Hamdi, M. (2017). Advanced oxidation process and biological treatments for table olive processing wastewaters: Constrains and a novel approach to integrated recycling process: A review. Biodegradation, 28, 125. 27. Rivas, F. J., Beltrán, F. J., Acedo, B., Gimeno, O., & Alavarez, P. (2003). Optimization of Fenton’s reagent usage as pretreatment for fermentation brines. Journal of Hazardous Materials, 96, 277. 28. Benitez, F. J., Acero, J. L., & Leal, A. I. (2003). Purification of storage brines from the preservation of table olives. Journal of Hazardous Materials, 96, 153. 29. Izmaylov, A., Briukhanov, A., Shalavina, E., & Vasilev, E. (2022). Pig manure management: A methodology for environmentally friendly decisionmaking. Animals, 12(6), 747.

G. Pilidis and I. Zarkadas 30. Chen, Y., Cheng, J., & Creamer, K. (2008). Inhibition of anaerobic digestion process: A review. Bioresource Technology, 99, 4044. 31. CRAFT Program, EU contract COOP-CT-2005017641 (PIGMAN) (2008). A sustainable solution for pig manure treatment. 32. Integrated Pollution Prevention and Control directive 96/61/EU 33. ANPHOS-Phosphorous removal in anaerobic effluents; EU Life Public Database; https://webgate.ec. europa.eu/life/publicWebsite/project/details/2115 34. Andrade, A., & Schuiling, R. D. (2001). The chemistry of struvite crystallization. Mineralogical Journal, 23, 37. 35. Akratos, C. S., & Tsihrintzis, V. A. (2007). Effect of temperature, HRT, vegetation and porous media on removal efficiency of pilot-scale horizontal subsurface flow constructed wetlands. Ecological Engineering, 29, 173. 36. Stefanakis, A. I., & Tsihrintzis, V. A. (2009). Performance of pilot-scale vertical flow constructed wetlands treating simulated municipal wastewater: Effect of various design parameters. Desalination, 248, 753. 37. Kotti, I. P., Gikas, G. D., & Tsihrintzis, V. A. (2010). Effect of operational and design parameters on removal efficiency of pilot-scale FWS constructed wetlands and comparison with HSF systems. Ecological Engineering, 36, 862.

Advanced and Novel Treatment Technologies

Electrode Materials and Their Effects on Electricity Generation and Wastewater Treatment in a Microbial Fuel Cell Andika Wahyu Afrianto and Sandhya Babel

electricity generated in this system is very small. The amount of waste degradation and electrical energy produced is influenced using electrodes. Modifying the electrode structure with polymers such as polyaniline can boost electrical energy generation by up to 63.6% (Yin 2019).

Abstract

Wastewater is an unwelcome by-product of industrial processes that has no economic value. Biological Oxygen Demand (BOD) and Chemical Oxygen Demand (COD) are two parameters affecting the water quality. Microbial fuel cell (MFC) is bio-electrochemical systems that enable direct energy harvesting from the wastewater via microbial activity while also oxidizing organic matter in the wastewater. MFC-based wastewater treatment can reduce environmental pollution parameters such as BOD and COD. MFC uses microorganisms’ catalytic activity to convert chemical energy to electricity. Bacteria will degrade the organic matter in the waste, and their catalytic activity will be able to reduce the contaminants from wastewater. This technology is very efficient at lowering BOD and COD levels. With the power generated about 4465 mW/m2, the COD and BOD levels of tempeh waste lowered by up to 88.9% and 34.0%, respectively (Sejati 2020). As an environmental technology that can reduce COD and BOD levels in wastewater and generate electrical energy, the amount of

A. W. Afrianto
S. Babel (&) School of Bio-Chemical Engineering and Technology, Sirindhorn International Institute of Technology, Khlong Nueng, Thailand e-mail: [email protected]

Keywords







Organic loading Electrode Microbial fuel cell Wastewater Power generation




1




Introduction

Water is a natural resource whose availability and abundance are crucial for the survival of all living things. The industrial revolution has led to severe water pollution problems as it produced many pollutants that were discarded and polluted the environment. This issue has caused an alarming danger to humans and all living organisms on earth. Water pollutants discharged into natural aquatic environments mainly come from industrial waste, agricultural runoff, and domestic waste [1]. Wastewater is an unwanted by-product of industrial processes that have no economic value. Wastewater needs to be treated through an adequate system so that it can be reused for some activities. Domestic wastewater management is vital in protecting the health of

© The Author(s), under exclusive license to Springer Nature Switzerland AG 2023 E. Debik et al. (eds.), Wastewater Management and Technologies, Water and Wastewater Management, https://doi.org/10.1007/978-3-031-36298-9_5

53

54

people. Most of this wastewater is directly discharged through open channels, polluting the environment [2, 3]. Energy security has become one of the central issues on the agenda of many countries in the world, along with problems related to wastewater that is getting worse. Fossil fuels are not a solution to meet the increasing energy demand because they have negative effect on the environment. Therefore, the development of renewable energy sources (RES) is a solution to overcome this problem. This diversification of energy sources is closely related to the inclusion of RES in the power generation sector [4]. Some options for renewable energy sources are batteries, supercapacitors, and fuel cells, which function as energy storage devices. A microbial fuel cell (MFC) is a bioreactor system based on electrochemical reactions that allow the production of electrical energy from wastewater through the microbial activity that degrades organic matter in the wastewater [5]. MFC-based wastewater treatment can reduce environmental pollution parameters such as Chemical Oxygen Demand (COD) and Biological Oxygen Demand (BOD). COD is defined as a unit of the amount of oxygen needed to oxidize organic substances in wastewater, and BOD is the amount of oxygen required by aerobic microorganisms to break down organic matter in wastewater at a certain temperature for a certain duration of time [6–9]. Bacteria will degrade the waste’s organic matter and help treating the wastewater. The MFC system has been employed for many wastewaters as a substrate, such as chocolate factory waste, household liquid waste, beer factory waste, tempeh industrial waste, dairy industry waste, synthetic wastewater, and more [10–12]. Since MFC can treat wastewater and produce electricity simultaneously by oxidation, they provide a low-cost and dependable option for tackling environmental pollution and energy shortages at the same time [13]. The MFC developed has low-power output as a good energy and environmental technology. There are many ways to boost MFC power output by isolating certain microbial species, selecting mediator-producing organisms, or

A. W. Afrianto and S. Babel

optimizing the electrode surface electrochemically [14]. One of the alternative ways to overcome this issue is by electrode modification. Modifying anodes can enhance the MFC performance to a large extent. One of the most critical elements affecting bacterial adherence and electron transmission between bacteria and the electrode is the surface characteristic of anode materials. The anode surface can be modified with various functional groups that may have varied selectivity for different bacteria and thus enhance electron transport from bacteria to electrodes [5, 15]. Therefore, this chapter focuses on modifying the electrode to improve electricity generation while treating wastewater.

2

Wastewater

2.1 Wastewater Pollution Water is essential for sustaining ecosystems and biodiversity, human health and well-being, energy generation, and food security. It also plays a crucial role in maintaining human growth. About 97% of the Earth’s surface is covered with water, but since most of it is in the seas with high salinity and, in most cases, large quantities of contaminants, it cannot be consumed directly. Only 3% of the world’s water is usable and present in lakes, rivers, water vapor, and soil moisture. The world’s population, which is presently over 7 billion, uses around 28.8 million km3 of fresh water annually [6, 16]. Water contamination has grown to be a significant concern in the current situation. Due to global industrialization’s acceleration and diversity, chemical contamination is one of the primary causes of water pollution. Drugs, phenolic compounds, microplastics, dyes, heavy metals, and pesticides are just a few of the tenacious substances that have been discharged into the environment [1]. Clean water is necessary to meet the population’s growing needs due to socio-economic progress. Due to economic activity, the deterioration in the quality of these resources becomes more evident (which require roads, dams, electric

Electrode Materials and Their Effects on Electricity Generation …

55

Fig. 1 Classification of wastewater pollutants

power, etc.). Without adequate treatment, wastewater is frequently dumped into neighboring rivers and streams, from where it travels far upstream to vast pools of freshwater and even impacts the groundwater. In most emerging nations, resource allocation for enhancing sanitary conditions and water treatment infrastructure has not kept up with economic expansion [6]. Figure 1 describes the types of possible pollutants present in wastewater. The source of the wastewater effluents determines the chemical composition and contaminants found in contaminated water bodies. Industrial effluents need to be treated before being released into water streams owing to the complex nature of their contaminants, which contain both organic and inorganic materials, including metals, dyes, and pharmaceutical residues. Aside from the industrial sector, other factors contributing to water pollution include population growth, economic development, agricultural intensification, and the incompatibility of wastewater treatment with wastewater generation. There is also an increase in the discharge of untreated wastewater. The quantity of liquid waste and environmental pollution generated by the household sector likewise rises with rising population density. Pathogens, suspended

particles, nutrients (nitrogen and phosphorus), and organic contaminants can be found in high concentrations in domestic wastewater. Largescale farming produces organic pollutants transported to the surface and groundwater, including pesticides and other agro-chemicals. Reducing these pollutants to acceptable levels can minimize environmental and human health risks. Since wastewater is reused in various ways, it is crucial to remove organic impurities and pathogens from it [3, 6, 16].

2.2 Wastewater Treatment Wastewater that is not appropriately handled and directly discharged into the environment, such as rivers and soil, will cause many adverse effects on the environment. Good wastewater management will maintain the sustainability of life in the ecosystem. Wastewater treatment has been done using various remediation techniques, including traditional treatment methods. These methods include biological treatments, coagulation/ filtration, ion exchange, sedimentation, activated carbon filtration, and adsorption [1]. Other wastewater treatment strategies were evaluated by various research teams and focused

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A. W. Afrianto and S. Babel

on recovering hazardous compounds from wastewater. Among these techniques are membranes-based treatment, adsorption with activated carbon, and electrodialysis. Two phenyl urea pesticides (diuron and isoproturon) were treated using a reverse osmosis membrane to remove them from farmland water. The results obtained from this method showed that the isoproturon elimination was satisfactory and reached 95% [1, 17]. Electrochemical wastewater treatment methods are also quite popular. Several researchers have developed and conducted research related to electrochemical wastewater treatment. One of the methods is to make a microbial-based bioreactor that utilizes microbial activity to reduce wastewater pollution parameters. Electrochemical bioreactors are widely used in wastewater treatment because they can produce electrical energy and reduce pollution levels.

3

Microbial Fuel Cell System

3.1 Background of Microbial Fuel Cell Microbial fuel cell (MFC) is one of the bioelectrochemical systems (BESs) that use microorganisms to convert biomass into electricity. MFC is a bio-electrochemical system that converts chemical energy from microbial activity in wastewater into electrical energy. The donation and acceptance of electron pairs in redox processes, such as the oxidation of organic matter at the anode and the reduction of substances of higher electrochemical potential at the cathode, are the basis for the operation of MFCs. Two electrodes are connected by an electric cable to collect the movement of electrons. Electrical energy is generated from the biodegradation process of organic matter. MFC is a possible long-term solution to satisfy rising energy demands, mainly when wastewaters are used as substrates [5, 14, 18]. MFCs have some distinct characteristics that make them superior to other technologies. (i) MFCs convert chemical energy to electric current more efficiently than other forms of batteries. (ii) Unlike other bioenergy technologies,

MFCs may provide good outcomes over a wide range of temperatures (20–40 °C). (iii) Because the cathode may aerate passively, an external power for aeration to provide oxygen (as an electron acceptor) is not required in the MFC system during the operation. MFCs are available in various materials and styles. These systems are frequently run under ideal circumstances to generate more energy, but they may also function under changed settings [19]. MFC technology provides an exciting solution for managing wastewater sustainability. Organic materials quickly oxidize as a fuel in the anode compartment, allowing MFCs to use for wastewater treatment. MFC-based systems have also been recently employed in various novel applications, including hydrogen generation, seawater desalination, biosensors, and microbial electrosynthesis. As an environmental technology, MFC can reduce ecological pollution parameters such as levels of COD and BOD [20, 21]. MFC system is a bioreactor system; hence, its operation will be strongly linked to biological reactions involving bacteria, such as glycolysis and electron transfer processes. Bacterial microorganisms in the MFC circuit will function as a catalyst, oxidizing organic and inorganic molecules to generate electricity. Microorganisms can be used as a biological catalyst, oxidizing organic substances in the electrolyte solution, and producing electrons flow from the anode to the cathode. In contrast, ions diffuse across the membrane in the opposite direction to maintain a neutral charge. This catalytic activity of microorganisms will reduce waste pollution parameters and generate the system’s electricity [22, 23]. Figure 2 illustrates the MFC principle. Elakkiya and Matheswaran [24] built a bioelectrochemical reactor to treat liquid waste from the dairy industry. An ion exchange membrane is used in this reactor. The results showed a maximum power density of 192 mW/m2 and a COD removal of around 91%. Faria et al. [12] used the same reactor system to treat dairy industry wastewater and produced a power density of 92.2 mW/m2 with a maximum COD removal of up to 63 ± 5%. Tamakloe et al. [25] studied a bioelectrochemical reactor system using an

Electrode Materials and Their Effects on Electricity Generation …

57

Fig. 2 General principle of MFC system

earthenware-type ceramic membrane as an ion exchange membrane. The reactor is arranged in a one-chamber system, where the pottery acts as a membrane and the anode chamber. The electrodes used are graphite/graphite and zinc/ copper. A power density of 118 mW/m2 for a capacity of 1.7 L with zinc/copper electrodes and 78 mW/m2 for a reactor with a capacity of 1 L with graphite/graphite electrodes was observed. The COD removal for this system is 87.0% for 1.7 L capacity and 88.1% for 1 L capacity. The reactor with a larger volume capacity produces higher power density, and the COD content strongly influences the power generation. In addition, the electrode also plays a role in the production of power density. The zinc/copper electrode used in the reactor system with a capacity of 1.7 L is a metal electrode having a better conductivity property compared to carbon materials such as graphite. In Indonesia, Sejati and Sudarlin [26] adopted this earthenware membrane in their reactor system to treat homebased tempeh industrial waste. This research uses carbon graphite as an electrode, tempeh waste as

an analyte, and KMnO4 as catholyte to obtain electricity. The maximum power density obtained in this study was 1448 mW/m2, with the removal of COD and BOD of 88.9% and 34.0%, respectively. A membrane often separates the two chambers in MFCs. Microorganisms in the anode chamber will oxidize organic compounds in the waste anaerobically. The electrons will reach the anode after being released from cellular respiration. The electrons pass through the external circuit across the resistor to the cathode. Protons and electron acceptors will combine with it. Oxygen is an example of an electron acceptor. This movement of electrons generates electricity. Electrode characteristics affect MFC efficiency [27, 28]. The redox reactions can link the MFC to generate electrical energy. In theory, any biodegradable organic material may be employed in an anode oxidation process. The compounds used in MFC research are sugars and organic acids such as glucose, acetate, and complex polymers (starch and cellulose) [14]. The reaction at the anode is shown in Eq. 1.

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C6 H12 O6 þ 6H2 O ! 6CO2 þ 24H þ þ 24e 0 Eo ¼ 0:43V versus SHE at pH 7, ð1Þ where E0' is the standard redox potential used to compare with the standard hydrogen electrode (SHE). Due to its favorable redox potential and quick availability in air, oxygen reduction is the most prominent reaction at the cathode for MFCs. The oxygen reduction reaction (ORR) is constrained by the low solubility of oxygen in the electrolyte. Hexacyanoferrate and ferric/ferrocyanide are other redox functional pairings for reactions at the cathode. The most helpful cathode reaction for power production in MFCs is characterized by a rising positive redox potential [29, 30]. Equation 2 depicts ORR, O:2 þ 4H þ þ 4e ! 2H2 O 0 Eo ¼ 0:82V versus SHE at pH 7:

ð2Þ

The Nernst equation governs an electrode’s equilibrium potential (Ee) in an open circuit, which is influenced by several reaction factors such as species concentrations, pH, and temperature. Ee ¼ Eo þ 2:303

RT Co log ; nF CR

ð3Þ

where the standard potential is expressed in Eo, and the R-value is 8.314 J/mol/K for the molar gas constant. The temperature (T) is in Kelvin, while the number of electrons transferred is represented by n. The Faraday constant, 96,485 C/mol, is expressed in F, and CO and CR are the

concentrations of the oxidized and reduced products. The calculation of the MFC cell potential is based on the difference between the cathode and anode equilibrium potentials. Electricity is created when Ecathode > Eanode, and the system is known as MFC [31]. Table 1 describes the oxygen reduction reactions in acidic and alkaline solutions.

3.2 Design of MFC Several types of MFC were developed, ranging from single-chamber MFC, dual-chamber, and stack MFC. A single-chamber MFC is a reactor system with only one room, in which the substrate and electrolyte solution are mixed. The anode chamber of a dual-chamber MFC contains the liquid waste substrate. In contrast, the cathode chamber includes an electrolyte solution, separated by a proton exchange membrane (PEM) or a salt bridge. A stack MFC is a collection of single- and dual-chamber MFC units organized in series, parallel, or series–parallel systems [32].

3.3 Single-Chamber MFC Figure 3 illustrates the single-chamber MFC system. The single-chamber MFC design with a cathode in direct contact with the air is attractive because it eliminates relatively expensive PEM use. This type of single-chamber MFC system is commonly used in wastewater treatment. Based on the high energy output, simple structure, and relatively low price compared to other kinds of

Table 1 Electrochemical O2 reduction Electrolyte

ORR reaction

Electrode potential at standard conditions, V vs. SHE

Acidic aqueous solution

O2 + 4H+ + 4e− ! H2O O2 + 2H+ + 2e− ! H2O2 H2O2 + 2H+ + 2e− ! 2H2O

1.229 (at pH 0) 0.70 1.76

Alkaline aqueous solution

O2 + H2O + 4e− ! 4OH− O2 + H2O + 2e− ! HO2− + OH− HO2− + H2O + 2e− ! 3OH−

0.401 (at pH 14) − 0.065 0.867

Electrode Materials and Their Effects on Electricity Generation …

59

Fig. 3 Schematic diagram of single-chamber MFC system

MFC, the MFC with a cathode in direct contact with the air is most likely to be scaled up in wastewater treatment. Removing the membrane can reduce reactor costs but has a disadvantage: the diffusion of oxygen to the anode [33].

3.4 Dual-Chamber MFC The anode chamber and the cathode chamber are the two vessel chambers that make up the MFC dual-chamber type. The rectangular dualchamber MFC system is depicted in Fig. 4. Protons generated in the anode chamber can travel to the cathode chamber through PEM used in the dual-chamber system, which acts as a separator for the two rooms. The anode and cathode are connected via an electrical circuit, which may be made of copper or titanium wire. The organic substrate of the waste is oxidized by microorganisms in the anode chamber, producing electrons, protons, and carbon dioxide.

Electrons produced by microbial metabolic activities are transported to the anode surface by redox-active proteins or cytochromes, where they are then transferred to the cathode through an electrical circuit. There will be a decrease in electrons in the cathode space. The electron acceptor in the cathode chamber is usually oxygen or ferric chloride, and the electrons combine with protons and oxygen at the cathode. This process can also be facilitated by catalysts such as platinum [19]. The dual-chamber MFC reactor with a rectangular shape is today’s most widely used reactor type. However, many other two-chamber reactor designs, including “H” shape and sedimentary reactors, may be utilized for MFC. As seen in Fig. 5, the reactor has an “H” shape that is relatively common for usage on a laboratory scale. The anode and cathode chambers are often connected via a tube to create this MFC. The connecting pipe is where the ion exchange membrane is put. The most common membrane

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A. W. Afrianto and S. Babel

Fig. 4 Schematic diagram of a rectangular dual-chamber MFC system

Fig. 5 Schematic diagram of dual-chamber “H” shape MFC system

in this system is nutrient agar. Because of its simplicity in building and operating, this MFC form is frequently utilized in laboratory investigations. The sediment MFC reactor (SMFC), also known as the marine MFC, differs from

traditional two-chamber MFCs because it uses a naturally existing oxygen gradient to divide the anode and cathode chambers. Figure 6 shows that the cathode is positioned above oxic water, and the anode is placed on organically rich marine sediments. Reactors of this kind are

Electrode Materials and Their Effects on Electricity Generation …

61

Fig. 6 Schematic diagram of sediment MFC system

appropriate for powering sensors, communication systems, and naval equipment.

3.5 Stacked MFC A stacked MFC system may be created by connecting multiple MFC reactors in series or parallel to increase the voltage output [34]. This technology works better in real-world settings to process large amounts of wastewater while generating electricity that can power other system elements like air or water pumps. Six reactors of stacked MFC with a total volume of 115 L were used to treat swine manure at 50 L per day for more than six months. The removal rate of organic matter and power densities were at 1.9 ± 0.3 kg COD/m3.day and 2–4 W/m3, respectively [35].

3.6 MFC Performance in Wastewater Treatment Currently, many studies related to MFC-based wastewater treatment have been carried out. So,

through the development of research carried out, the use of the MFC reactor is expected to be an alternative in the development of renewable energy sources that are not only able to generate electricity and to solve the energy crisis, but also become an alternative for wastewater treatment that is very safe and efficient for the environment. Table 2 shows a comparison of the performance of various MFC reactors in managing various types of wastewaters. Table 2 also shows that the electrical output generated from the MFC system is very small, but it is effective in wastewater treatment and can reduce COD and BOD levels from various types of wastewaters. In most studies, a COD reduction of about 80% was reported with different types of wastewaters. In the operation of the MFC using tempeh industrial wastewater, the MFC reactor uses a pot-shaped membrane made of clay. This pot is used as an anode chamber and a membrane for ion exchange. The COD level of tempeh liquid waste can decrease up to 88.9% from the original 119,750 mg/L to 13,320 mg/L and the BOD level decreased from 46,200 mg/L to 30,500 mg/L (33.9% removal).

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A. W. Afrianto and S. Babel

Table 2 Comparison of MFC performances Type of MFC

Type of wastewater

COD/BOD removal

Power density (mW/m2)

Reference

Single-chamber

Urine

COD removal: 46 ± 4%, No BOD analysis

5.4

[36]

Urine

COD removal: 50 ± 3%, No BOD analysis

6.2

[36]

Brewery industry

COD removal: 80%, No BOD analysis

483

[10]

Brewery industry

COD removal: 87.0% and 88.1% No BOD analysis

55 and 69

[25]

Dairy wastewater

COD removal: 63 ± 5%, No BOD analysis

92.2

[12]

Tempeh waste

COD removal: 88.9% BOD removal: 34.0%

4465

[26]

Dual-chamber

4.2 Electrolyte in Cathode Chamber

4

Electrodes in MFC System

4.1 Cathodes in MFC—Cathodic System

An electrolyte solution is used in a cathode chamber due to ion transfer mechanisms. Two electrolyte solutions that are commonly employed in MFC systems are potassium permanganate (KMnO4) and potassium hexacyanoferrate(III) (K3[Fe(CN)6]). Because it has a high standard reduction potential and belongs to the peroxide group, which can release oxygen during oxidation, KMnO4 can utilize as MFC electrolyte cathodes [11, 12, 24]. Logan [30] reviewed the use of K3[Fe(CN)6] as an electrolyte cathode because it acts as an electron acceptor in the MFC system. K3[Fe(CN)6] is a highly electroactive species capable of grabbing electrons with a standard reduction potential of +0.36. Electron acceptors receive the electrons and contribute significantly to the MFC’s performance. Electron acceptors have the potential to boost power output and lower operating costs while also broadening the application spectrum of MFCs. Some refractory substances can now act as electron acceptors in the cathode. These findings show that MFCs might be utilized to manage pollution [21]. The following are some electrolytes used as electron acceptors in MFC.

The carbon material is the most common type of cathode used in MFC systems because of its chemical stability, biocompatibility, low cost, and high conductivity. Carbon materials used as MFC cathodes include carbon felt, graphite plates, graphite fiber brushes, granular graphite, and stainless steel mesh. A hydrophobic diffusion coating layer is required on the aerated side of an air cathode to allow a high oxygen intake to the cell while simultaneously limiting water loss. The cathode material influences the resulting potential. The cathode material can also be used as an anode. Since the reaction at the cathode will be constant (abiotic) regardless of changes in microbial metabolism at the anode, the MFC’s efficiency depends on anodic oxidation rather than cathodic reduction. However, the more efficient the cathode, the greater the influence of the MFC in electricity generation. For efficient MFC design, a suitable cathode is now required. MFC technology’s most challenging aspects are the cathode material and architecture. Carbonbased materials as cathodes have desirable (a) Oxygen properties, such as good chemical stability and biocompatibility, a relatively lower cost than The electron acceptor commonly used at the other materials, and high conductivity [31]. cathode is oxygen. It is because oxygen has a

Electrode Materials and Their Effects on Electricity Generation …

high oxidation potential and produces a clean product in the form of water [37]. An external circuit in MFC transports electrons to the cathode compartment. The reaction between protons and oxygen will produce water in the procedure described in Eq. 4. O2 þ 4H þ þ 4e ! 2H2 O

E0 ¼ 1:23 V: ð4Þ

According to Eq. (4), the uses of oxygen continue to keep electrical generation potential alive. By using the air cathode, it can supply oxygen to the cathode. However, the limited interaction of oxygen with the electrodes and the slow rate of oxygen reduction at standard carbon electrodes are drawbacks that limit the use of oxygen in MFCs [21]. (b) Permanganate Permanganate is converted to manganese dioxide by receiving three electrons in acidic and alkaline environments. Permanganate is a possible electron acceptor because of its characteristics. Permanganate is predicted to produce more power under acidic settings, since its oxidation potential is more significant than in alkaline ones. As a result, permanganate research was conducted at various pH levels in MFCs [38]. þ  MnO 4 þ 4H þ 3e ! MnO2 þ 2H2 O

ð5Þ

  MnO ð6Þ 4 þ 2H2 O þ 3e ! MnO2 þ 4OH

The reduction of MnO 4 will be accelerated if the cathode contains large number of protons. As a result, according to the Nernst equation (Eq. 7), the concentration of MnO 4 at the cathode decreases and the potential at the cathode increases [11]. E ¼ E0 

RT ½MnO2  ln : nF ½MnO4 

ð7Þ

Previously, in You’s [38] study, permanganate was used as an electron acceptor in the MFC system. The use of permanganate as an electron acceptor showed promising results. The resulting

63

power density of this system is 115.6 mW/m2 and has a current density of 0.017 mA/cm2. The use of permanganate showed yields 4.5 and 11.3 times greater than those produced by using hexacyanoferrate(III) (25.6 mW/m2) and oxygen (10.2 mW/m2), respectively. Furthermore, in the same investigation, MFC with permanganate as an electron acceptor obtained a maximum power density of 3986 mW/m2 at 0.6 mA/cm2. (c) Hexacyanoferrate(III) Hexacyanoferrate(III) has a concentration that is not limited by solubility like oxygen, so it is widely used as another electron acceptor in MFC research. Equation 8 shows that the standard redox potential of hexacyanoferrate(III) is not as high as that of oxygen. It has a much lower overpotential and results in a much faster reaction rate and a greater power output. Due to the increased mass transfer efficiency and more significant cathode potential, hexacyanoferrate(III) with carbon electrodes produces 50–80% more power than oxygen with a Pt-carbon cathode [21]. 4  FeðCNÞ3 6 þ e ! FeðCNÞ6 :

ð8Þ

Although hexacyanoferrate(III) is an efficient electron acceptor for power production, it is well known that potassium hexacyanoferrate(III) is not a realistic long-term solution. It is poisonous, and chemical regeneration and recycling are complex. As a result, the use of hexacyanoferrate (III) is only done in basic laboratory experiments. Due to its stability and good system performance, hexacyanoferrate(III) is still a significant cathodic electron acceptor for proving certain fundamental principles in the laboratory [30]. Elakkiya and Manickam [24] used hexacyanoferrate(III) as an electrolyte in an MFC reactor to treat liquid waste from the dairy industry. The ion exchange membrane used is the Nafion 117 type which was assembled in a dualchamber reactor system. The maximum power density obtained was 192 mW/m2. Another study conducted by You [38] using permanganate as an electron acceptor generated a power density of 116 mW/m2 at a current density of 0.017 mA/cm2.

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A. W. Afrianto and S. Babel

4.3 Anodes in MFC—Anodic System Several factors can give impact the power density of MFC, including electron movement from microorganisms, ohmic loss, over-potentials, microbial inoculum, and indirect effects such as electrode cost and cell design. The composition, shape, and surface features of anode materials will influence the bacterium adhesion, electron transport, and substrate oxidation in MFCs. The optimum anode materials for MFC applications must have excellent conductivity, increased biocompatibility, high stability of electrodes in chemical and physical properties, and cheap cost. Carbon-based materials are the most frequent and appropriate anode materials in MFC systems. It is due to the carbon material’s high conductivity, stability, strong structure, sufficient surface area, and optimal surface characteristics for biofilm formation. Carbon can be found as an anode material, including bars, foams, slabs, granular, and activated carbon [39].

4.4 Anode Materials (a) Carbon Materials Research on MFCs using carbon materials focuses on optimizing power density based on membrane volume or area. Table 3 shows the

types of anodes based on carbon materials and their characteristics and performances in the MFC system. Other studies have focused on the power density on the surface area of the electrodes making quantitative comparisons across carbon materials challenging. According to published data, the carbon brush topology provides better MFC power density than planar configurations such as paper, mesh, felt, and sheet [39]. One of the significant drawbacks of using carbon material as an MFC anode is the loss of electrocatalytic activity for microbial reactions because the biofilm will clog pores and reduce efficiency. The low specific area, lack of durability, brittleness, and high cost of carbon materials prevent these electrodes from being used in large-scale MFCs [39]. (b) Metal Materials Table 4 shows the types of metal materials in the MFC system. Metal materials are utilized as MFC anodes because they have better conductivity than carbon materials. The non-corrosive nature of MFC anodes limited the metal materials employed in MFC applications. Several studies have investigated stainless steel and titanium as suitable anode materials [39]. Compared to carbon electrodes, stainless steel has extraordinary mechanical qualities for long-term operation and

Table 3 Types of carbon anodes Anode materials

Properties

MFC performance 2

Reference

Carbon paper

Very thin, low specific area, easy to connect the wire, high cost, lack of durability, and slightly brittle

600 mW/m (bottle-MFC)

[40]

Carbon cloth

Thin, expensive, high flexibility, and more porous compared to carbon paper

1040 mW/m2 (cube-shaped MFC)

Carbon brush

High surface area, high porosities, and efficient current collection

2400 mW/m2 (cube-shaped MFC)

Graphite plate

Relatively smooth surface, low specific area, and high cost

3290 mW/m2

[41]

Graphite felt

Thick, large porosity, large resistance, and good for bacteria growth

386 W/m3

[42]

Graphite granular

High specific area and low porosities after long-term running

175 W/m3

Electrode Materials and Their Effects on Electricity Generation …

65

Table 4 Metal materials applied as anode in MFC Anode materials

MFC system

MFC performance

Reference

Stainless steel

Marine MFC with stainless steel plates as electrode immersed in the surface of seawater. Zinc wire employed as a pseudo-reference electrode for the system

23 mW/m2

[43]

Cu

Copper wires as anode and gold-covered copper wires as cathode. Chambers separated by salt bridge membrane and system used for bakery wastewater

40 W/m2

[44]

Zn

Zinc used as electrodes in the dual-chamber MFC system with earthenware pot membrane made by clay. Brewery wastewater sludge used in this system

69 mW/m2

[25]

Au

Gold used only in the anode chamber with carbon as a cathode. Nafion 117 was used as a membrane in this micro-MFC with bacterial culture from Shewanella oneidensis

95 lW/cm2

[45]

scale-up. MFC may use stainless steel as both an anode and a cathode. Stainless steel has the potential to be a competitive material for creating bio-anodes with natural microbial consortia when used in conjunction with proper inoculation protocols and potential control.

5

Modification of Electrode Materials

5.1 Cathode Catalyst Platinum has been widely used as a cathode catalyst, although it is expensive and limits its use. It is due to platinum having a low potential for reducing oxygen. Several studies have been carried out to investigate the catalytic potential of precious metal-free cathodes in MFCs to lower the total cathode charge while improving the ORR kinetics. Base metals and low-cost metal oxides have shown performance on par with Pt cathodes, such as cobalt, manganese dioxide, cobalt-tetramethylphenylporphyrin, and pyrolysis of Fe(II)phthalocyanine, lead dioxide, metal porphyrins, and phthalocyanines [31]. However, metal discharge from the cathode is a potential stumbling block that might limit the use of metal oxide catalysts in MFC cathodes.

Another flaw in metal-based catalysts’ performance is their susceptibility to increased pH values in the cathode chamber caused by cations crossing the membrane. Rhoads et al. [46] conducted one of the biocatalyst investigations, in which manganese-oxidizing bacteria were added to the cathode to compare the generated current to an untreated graphite electrode. However, significant levels of activation over-potentials in the cathode restrict biocatalyst production. With a maximum power density of 126 mW/m2, the current density provided by manganese oxide as the cathodic reactant was nearly twice that of oxygen. Furthermore, increasing the number of metabolites and ions passing across the membrane might stifle bacterial development by donating electrons to bacteria, counterbalance the biocatalyst function, and lower overall efficiency. Alternatively, treating pollutants (such as nitrates) in the cathode chamber might benefit biocatalysts [31].

5.2 Cathodic Surface Several beneficial surface treatments were proven effective, such as treating carbon fabric electrodes with 5% ammonia gas or oxidizing carbonic substrates on the surface. In alkaline and

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A. W. Afrianto and S. Babel

acidic environments, nitrogen surface integration on carbon has boosted oxygen reduction selectivity and catalytic activity via four-electron pathways. In addition, it is possible to immerse the cathode material in acid to increase the power gain. Based on research, soaking carbon fiber brushes using sulfuric acid can increase power gain by up to 8%. In addition, combining the acid with heating can also increase the power by 25% compared to the untreated material. The ratio of protonated nitrogen will increase to total nitrogen. It will potentially impact the biocatalyst, which can increase power production by acid and heat treatment. It will also increase the area and accumulation of positive charge, specifically on the electrode surface [31].

5.3 Composite Anode Materials Creating novel anodes based on composite materials will necessitate more research to assess prices, long-term stability, and the mechanism of bacteria–electrode interaction. Understanding the bio-electrocatalytic mechanism of the two oxidation processes and their kinetic characteristics can substantially aid MFC anode optimization [39]. Mn2+-graphite anode and neutral red (NR) covalently connected woven graphite anode were created by Park and Zeikus [47] to improve electrical energy generation compared to standard graphite electrodes. Mn2+-graphite anodes outperform NR-woven graphite and outperform standard graphite electrodes. The maximum power density obtained was 787.5 mW/m2 with Mn4+-graphite anode, 5.32 mW/m2 with NR-covalently linked graphite anode, and 0.65

mW/m2 with a woven graphite anode. Other types of anodes based composite materials are shown in Table 5. Table 5 describes the performance of composite materials applied as anodes in MFC. Two graphite composites were applied in the marine sediment MFC system, while the aluminum composites in the MFC system were with Escherichia coli. In this system, the graphite pastes with an incorporated Sb(V) complex produces the best results in terms of energy production, with a maximum power density of up to 115 mW/m2 when compared to graphite-ceramic composites containing Mn2+ and Ni2+, which only produce a maximum power density of 105 mW/m2. The production of power density can be affected by differences in the compounds or ions contained or coated in the composite of graphite electrodes.

5.4 Anode Based Carbon Nanomaterials Due to their remarkable physicochemical features, such as high surface area, conductivity, and mechanical strength, carbon nanotubes (CNTs) and graphene (G) are prominent carbon nanostructures. In micro- and macro-sized MFCs, CNTs have demonstrated excellent performance as anode-modifying materials. Surface modification of CNT-based anodes is required to reduce activation losses and cellular toxicity. CNTs’ surface activity, processability, and biocompatibility have increased due to advances in chemical modification and functionalization processes, paving the door for a wide range of bioapplications [39].

Table 5 Composite materials applied as anode in MFC Anode materials

MFC performance

References

Graphite-ceramic composite anodes containing Mn2+ and Ni2+

*105 mW/m of composite anode compared with graphite anode *20 mW/m2

[39]

Graphite paste with an incorporated Sb(V) complex

*115 mW/m2 of composite anode compared with graphite anode *25 mW/m2

Aluminum-alloy mesh composite carbon cloth electrode

2966 mW/m2 of aluminum-alloy mesh composite carbon cloth compared with aluminum-alloy mesh 22.4 mW/m2

2

Electrode Materials and Their Effects on Electricity Generation …

CNTs have gotten much interest because of their unusual properties. CNTs are long (up to millimeters), constricted (1100 nm) carbon atom cylinder structures with a cap at either end. CNTs refined to a high degree of purity are widely employed as building blocks in sophisticated materials with exceptional features. CNT-based structures are used in many applications, including microelectronics, tissue engineering, biosensors, and energy storage materials, due to CNT's magnetic properties, high surface area to volume (SAV), adsorption properties, and biocompatibility [48]. The examples of modified anode materials with carbon nanostructures and their performances in generating electricity in MFC systems are shown in Table 6. Various studies using carbon nanotubes (CNTs) as electrode materials show that CNTs provide advantages for MFC and nanomaterials’ researchers. Carbon nanotubes effectively improve the performance of MFCs, but CNTs need to develop for more outstanding biocompatibility in bacterial growth and adherence. Modifying the nature of the functional groups on the CNT surface can change the chemical properties of the surface and make it more suitable for practical applications [5]. The critical factor influencing the improvement of MFC performance is the abundance of C–OH and C=O groups. It can be obtained by modifying graphite using polyhydroquinone (PHQ) [5]. The increase of the adhesion strength on the CNT

67

surface with bacteria was influenced by the presence of the –OH as functional group on the CNT surface. This is known based on an investigation by Thepsuparungsikul et al. [53] related to the role of -COOH or –OH groups on the CNT surface. Acid treatment is one of the most common CNT modifications, which increases cell adhesion. The oxidation process in the acid treatment will create the presence of carboxylic or hydroxyl groups on the walls and ends of the CNT. Most of the research relates the conductive layer on the CNT to increase the surface area and conductivity of the anode, which will improve the performance of the MFC. These findings indicate that the addition of CNTs will increase the capacity of microorganisms to deliver electrons to the anode surface while lowering the internal resistance [39].

5.5 Anode Based Polymers Coating Anode modification is an effective method to overcome the low electricity production of the MFC system. The effectiveness of electron transfer from microorganisms can overcome problems related to MFC performance. Modifying the anode material can provide high electrical conductivity and increase the anode material's surface area, porosity, and biocompatibility. Furthermore, anode materials’ availability is critical for MFC commercialization [39].

Table 6 Modification of several anode materials with carbon nanostructures Carbon nanostructures

Anode

MFC performance

Reference

CNT

Carbon paper

*260 mW/m of maximum power density production

[49]

CNT

Glassy carbon electrode

82 times greater compared with non-modified anode with maximum current is 9.7 mA/cm2

[50]

Graphene

Stainless steel mesh

2668 mW/m2 power density production (18 times improvement compared to unmodified stainless steel mesh anode that only produced 142 mW/m2

[51]

Graphene oxide

Carbon paper

34.2 mW/m2 of power density and 30 A/m2 of current

[52]

Reduced graphene oxide

Carbon cloth

2.7 W/m3 compared with unmodified 1.7 W/m3

[53]

2

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A. W. Afrianto and S. Babel

Bacterial adhesion and electron transfer are affected by the nature of the anode surface used. Modification of the anode surface can be done by adding a functional group that may have selectivity against bacteria so that it will be able to facilitate the process of electron transfer from bacteria to the anode [5, 15]. Table 7 shows several anodes’ modification with conducting polymers. Anodes coated with conducting polymer may be used to improve MFC performance. Polyaniline (PANI) and polypyrrole (PPY) are two conductive polymers that have gotten a lot of interest due to their high conductivity, outstanding redox characteristics, good stability for the environment, and relatively easy synthesis of these two polymers [39]. PANI is the most common conducting polymer used to modify MFC anodes because of its environmental stability, controlled conductivity, superior biocompatibility, and ease of processability. Furthermore, its nanocomposites are an ideal coating material for MFC electrodes, resulting in significantly improved power generation [15]. PANI has been widely studied in electrochemical applications. Its unique chemical and physical properties depend on oxidation and

protonation. PANI is a polymer that is easy to synthesize even in its nanostructure. There are several anodes modified by PANI in the MFCs, such as carbon and platinized carbon cloth coated with PANI, glass carbon modified with PANI nanostructured, electrochemically deposited PANI on graphite fiber, carbon cloth coated with H2SO4-doped PANI, and the modification of indium tin oxide conductive glass by using polyaniline nanowire network (PANI-NN) [39]. Electrode coating can be conducted by using PANI as the base nanomaterial. The conductivity properties of PANI will increase the surface area of the electrode and reduce the resistance of extracellular microbial electronic transmission to the electrode surface, increasing the production of electricity in an MFC [15]. Yin et al. [58] investigated the performance of MFCs by comparing the PANI nanocomposite TiO2-modified carbon paper (TiO2-20PANI/CP) and the TiO2 nanosheets modified carbon paper anode (TiO2NS/CP). The research shows that 813 mW/m2 is the maximum power density for MFC with TiO220PANI/CP. The results showed an increase of 63.6% of MFC with TiO2-NS/CP anode (without adding PANI). It is because the charge transfer resistance at the anode interface is significantly

Table 7 Types of anode modification with conducting polymers Anode materials

Polymers

Electricity generated

Resources

Carbon felt

PANI

Maximum power density production of carbon feltmodified PANI is 27.4 mW/m2 compared with 20.2 mW/m2 of unmodified carbon felt

[54]

Carbon felt

Poly(aniline-co–oaminophenol) (PAOA)

Maximum power density production of carbon feltPAOA is 23.8 mW/m2 compared with 20.2 mW/m2 of unmodified carbon felt

Carbon cloth

PANI

Maximum power density production of carbon clothPANI is 5.16 W/m3 compared with 1.94 W/m3 of unmodified carbon cloth

Carbon fiber

Fibrillar and granular PPY

3.4 mW/m2 and 3.1 mW/m2 are the power density of modified-carbon fiber versus unmodified one, respectively

[55]

Graphite felt

Electrochemically deposited PANI

Maximum power density production of graphite feltPANI is 4 W/m3 compared with 1.7 W/m3 of unmodified graphite felt

[56]

Reticulated verified carbon (RVC)

PPY

Maximum power density production of RVC-PPY modified is 1.4 mW/cm3 compared with 0.42 mW/cm3 of unmodified RVC

[57]

Electrode Materials and Their Effects on Electricity Generation …

reduced due to the synergistic effect of vertically oriented TiO2-NS and PANI. It demonstrates that anode modification with polymer compounds can provide bacteria with good environmental stability.

6

Conclusion

The MFC has several components: an ion exchange membrane, electrodes, and a wastewater substrate. The main problem of the MFC system is the minimal energy output. The electrode material is a determining factor for the overall performance of MFCs in addition to other considerations since its composition, shape, and surface qualities directly influence microbial adhesion, electron transfer, and substrate oxidation. Electrode modification is an alternative method to increase the electricity production of the MFC system. The electrode modification leads to effective transfer of electrons from microorganisms to the anode due to high electrical conductivity of the anode material. In addition, increasing the surface area, porosity, and biocompatibility of the anode can reduce the problems associated with MFC performance. There are many ways to modify the electrodes in an MFC system. Modification with conductive polymers is one option. Conductive polymers offer outstanding electrical and optical characteristics and high conductivity/weight ratios. They can be made biocompatible, biodegradable, and porous. Incorporating antibodies, enzymes, and other biological components into conductive polymers can be carried out according to the demands of the application to improve chemical, electrical, and physical characteristics. Based on the review, it can be concluded that the carbon materials are the best electrode material that can be applied to the cathode and anode chambers. Carbon material has high conductivity and stability, a strong structure, and is suitable for the microbial catalytic environment. Modifying carbon electrodes with polymer coating can improve environmental stability, which is suitable for microbes. Polyaniline (PANI) and polypyrrole (PPY) are two

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conductive polymers that have attracted much attention due to their high electrical conductivity, excellent redox properties, environmental stability, and ease of manufacture. Composite materials such as carbon composite materials with conductive polymers are alternative electrode materials that can be used and provide excellent effects in electricity production and wastewater management. In addition, the system’s operation with a combination of two types of electrode material can also be carried out on the MFC system. Combining metal materials with carbon materials or composite materials in different anode and cathode chambers can effectively treat wastewater and generate electricity in the system.

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Microbial Electrolysis Cells for Biohydrogen Generation and Wastewater Treatment—A Short Review and Current Trends Victor Alcaraz-Gonzalez, René Alejandro Flores-Estrella, Marcelo Nolasco, Vitor Cano, and Victor González-Alvarez

this wastewater treatment process focuses mainly on the removal of organic matter and nitrogen and neglects the valorization of by-products as an energy resource. In this context, Microbial Electrolysis Cells (MEC) have recently gained significant interest due to their potential for energy recovery in the form of hydrogen from renewable soluble organic matter. Hydrogen promises to be a viable alternative as a clean energy soon. In this way, MEC provide a new alternative to simultaneously treat wastewater and to produce bioenergy. In this chapter, MEC systems are reviewed. Contribution on modeling, experimental lab scale, achievements at pilot scale, barriers to industrial-scale adoption, latest contributions, and the next stage of research lines are considered.

Abstract

The world growing population and industrial activities are demanding energy at increasing rates in the last decades. It is well known that the demand for hydrocarbons, which are one of the most energy sources used today, exceeds the reserves forecast for the next decades. Thus, the whole world is searching for alternative, environmentally friendly energy sources able to satisfy the current and future energy demands. Nowadays, high-rate conventional wastewater treatment technologies rely on biological and physicochemical processes to meet the stringent environmental policies on disposal. Aerobic activated sludge stands out for its high treatment efficiency, but

V. Alcaraz-Gonzalez (&)
V. González-Alvarez Departamento de Ingeniería Química, Universidad de Guadalajara, Guadalajara, México e-mail: [email protected] R. A. Flores-Estrella Departamento Ciencias de la Sustentabilidad, El Colegio de la Frontera Sur, Tapachula, Chiapas, Mexico M. Nolasco Sustainability Graduate Program, School of Arts, Sciences and Humanities of University of Sao Paulo, Sao Paulo, Brazil V. Cano Department of Hydraulic and Environmental Engineering, Polytechnic School of University of Sao Paulo, Sao Paulo, Brazil

Keywords







Biohydrogen Industrial application Microbial electrolysis cell Modeling Lab-scale application

1







Introduction

Nowadays, despite global efforts to move toward renewable energy sources and the commitment of the industrialized countries to alleviate climate change, fossil fuels still account for approximately 80% of current global energy demand. On

© The Author(s), under exclusive license to Springer Nature Switzerland AG 2023 E. Debik et al. (eds.), Wastewater Management and Technologies, Water and Wastewater Management, https://doi.org/10.1007/978-3-031-36298-9_6

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the one hand, it is well known that reserves of these fuels are limited and expected to last soon. If extraction rates remain at their current levels, it is estimated that proven coal and oil reserves could last at most 110 and 50 years, respectively [1, 2]. On the other hand, it is also well known that the use of fossil fuels represents the main source of CO2 emissions, which contribute mostly to global warming. It is for these reasons that the entire world scientific community is in the process of implementing and adopting global policies to reduce energy demand, to reduce dependence on fossil fuels, and to avoid carbon dioxide emissions. The World Energy Outlook Report 2021 [3] highlights four key measures that can help close the gap between today’s pledges for the next ten years and support further emission reductions after 2030: (a) a massive additional push for clean electrification, (b) a relentless focus on energy efficiency, (c) a broad drive to cut methane emissions from fossil fuel operations, and (d) a big boost to clean energy innovations. Regarding the latter ones, currently there exist several alternatives of renewable and clean energy sources, e.g., nuclear, hydropower, geothermal, ocean, solar, wind, and biomass, which can be stored in the form of electricity or as chemical energy in the form of hydrogen. Certainly, hydrogen is one of the greatest promises as a renewable fuel and energy source in the medium and long terms. Hydrogen can be obtained from other sources, but biomass has the highest energy and exergy efficiencies (60.5% and 52.5%, respectively), even above nuclear energy [4]. In the universe, the most abundant element is hydrogen, and it is well known that this element has the highest calorific value of all fuels currently used (120 MJ/kg) that almost triples other fuels, such as gasoline, which has an energy content of 44 MJ/kg. Hence, the interest in this element as a source of clean and renewable energy increases. Other advantages are that their production, storage, transport, and end-use do not harm the environment. Furthermore, considering that one of the largest sources of biomass is organic waste and wastewater, obtaining energy from these raw materials could

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allow to mitigate environmental deterioration and to contribute as an alternative solution to the global energy problem [5, 6]. However, the energy available in biomass cannot be recovered by traditional chemical processes because the organic matter in the waste has a complex composition, and the use of bioreactors is necessary [7–9]. Some of the methods that use anaerobic microorganisms have been well known for several decades, for example, anaerobic digestion (AD) to produce CH4 and dark fermentation (DF) to produce biohydrogen [9]. Nonetheless, a novel and potentially more efficient technology to produce biohydrogen, discovered just two decades ago, is the use of Microbial Electrolysis Cells (MEC). In fact, before the discovery of Bioelectrochemical System (BES) as potential producers of biohydrogen, the best-known method was dark fermentation of waste carbon-rich organic effluents, which is relatively easy to implement. However, DF has relatively low efficiency due to simple thermodynamic issues, in which much of the recoverable theoretical hydrogen is still trapped in volatile fatty acids (VFAs). Unlike the widespread belief of the scientific–technological–academic community that BES is relatively recent, the truth is that they were already known since the beginning of the last century, although it was certainly not until the end of the century that their potential as an alternative source of energy as well as other uses were recognized. In [10], a magnificent historical recapitulation is made, in which three key moments are highlighted: (1) the discovery of the electrical activity of certain microorganisms by Potter in 1910 [11], (2) the first Microbial Fuel Cells (MFC) System for producing electrical energy in by Cohen in 1931 [12], and (3) the discovery of the capability of this system to produce hydrogen in 2005 (i.e., MEC) [13], and since then, several examples of other uses such as selective production of methane or hydrogen, nutrient recovery, desalination among others. Particularly with respect to MEC, Dange et al. [14] also present a timeline showing the chronological development of MEC technology

Microbial Electrolysis Cells for Biohydrogen Generation …

from 2005 to 2021. Indeed, MEC turns out to be a highly promising alternative for producing hydrogen from wastewater and has encouraged vigorous research over the past two decades to develop the technology to bring this process to an industrial scale. According to Zhao et al. [15], the number of publications on MEC almost quadrupled in the last decade only (2009 to 2019) and suggests that this trend in MEC has also contributed significantly to the large increase in publications in the top ten related research areas, particularly in Energy Fuels. Thus, although some fundamentals are common among different BES, this chapter focuses mainly on their application in MECs.

2

Bio-electrochemical Systems— Fundamentals

Main exoelectrogenic (or electroactive) bacteria involved in BES are Geobacter strains (e.g., G. sulfurreducens, G. metallireducens), but other species like Roseivivax sp., Chlorobium, Desulfuromonas, Pseudomonas, Enterobacter cloacae, and Rhodopseudomonas can be also present depending on the type of organic substrate [16, 17]. Biochemical reactions are catalyzed by these exoelectrogenic bacteria on the surface of an electrode, called anode, producing protons and electrons from the oxidation of organic or inorganic substrates. The electrons are transferred from the bacteria to the anode in a process named extracellular electron transfer, while the protons migrate from the anode to the cathode through a selective ion system (e.g., proton exchange membrane) [18, 19]. Then, in MEC, a small potential applied externally in the circuit allows the production of H2 in the cathode under anaerobic conditions by reducing the protons produced in the anode [20, 21] (see Fig. 1). Potential of electrodes can be calculated by using the Nernst equation:   RT Products E¼E  ln nF Reactants 0

with

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E (V) is the cell potential, E0 (V) is the standard potential, R is the ideal gas constant, T is the temperature in K, n is the number of electrons transferred in the reaction, and F is the Faraday Constant. Considering acetic acid as a model organic molecule, the following reactions and their theoretical half-cell potentials (calculated with the Nernst equation at pH = 7 and standard conditions: pressure; P = 1 atm and temperature T = 278 K) take place in the anode and cathode, respectively: Anode: CH3 COO þ 4H2 O þ  ! 2HCO 3 þ 9H þ 8e Ea ¼ 0:3 V Cathode: 2H þ þ 2e ! H2ðgÞ Ec ¼ 0:414 V The potential difference for these two reactions can be easily calculated as: E ¼ 0:3  0:414 V ¼ 0:114 V. The negative sign means that the overall reaction is not spontaneous and, therefore, it is necessary to apply 0.114 V for hydrogen production to take place at the cathode. Then, external energy applied in MEC is lower (by a factor of 10.7) in comparison to the hydrolysis of water (1.220 V), the industrial method currently used to obtain hydrogen. Of course, these data are theoretical, and in practice, different carbon sources other than acetate may be present, especially if the feedstock for these systems comes from agroindustrial wastewaters still rich in organic matter. In such a case, it would be necessary to perform the above calculations for each organic molecule and to consider the concentrations for each type of feedstock. Different internal resistances can also be present, and there are other factors to consider such as the type of bacterial culture, the type of electrodes, the system configuration and architecture. In [22–26], detailed lists of all factors that affect the efficiency of MEC can be found, while Kadier et al. [27] present a detailed list of various substrates used in different reactor configurations and their respective half-cell potentials. According to this last work, 0.5–

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Fig. 1 Microbial electrolysis cells’ system

1.06 V may be needed in practice to activate the reaction even using acetate as substrate. However, several of the factors mentioned also affect the process of water hydrolysis, and in practice, it is always necessary to apply an overpotential. For example, still using supercapacitors, Yu et al. [28] have reported a range of 1.4–2.0 V for hydrolysis of water to take place. Therefore, if a comparison is made again between these two processes, MEC still require much less energy than the hydrolysis of water.

3

Modeling Microbial Electrolysis Cells

From a phenomenological approach, modeling microbial electrolysis cell systems involves different aspects, i.e., physical, chemical, biological, and electrochemical (Fig. 1). The basic principles of some of them have already been outlined in the preceding section. Thus, once the basic principles of BES have been established, in the following paragraphs, a short review of how these aspects is considered in mechanistic models is further made. For other kind of models, e.g., based on neural networks or fuzzy logic, the reader is invited to consult [29] and the references herein.

3.1 Electron Transfer Mechanisms and Kinetic Modeling The first models to explain and to interpret the kinetic phenomena involved in the anode were proposed in the first decade of this century. For instance, in Marcus et al. [30], a Nernst–Monod type equation was proposed to express the current density in the anodic biofilm, while Torres et al. [31] reviewed the three main mechanisms of electron transfer in the anode, namely (a) direct contact, (b) oxidative and reductive mediators, and (c) nanowires. Then, precisely using the concept of Nernst–Monod kinetics, the presence of nanowires has been established as the most plausible mechanism [19].

3.2 Electrochemical Phenomena The classical electrochemistry principles have been successfully applied to MEC with virtually no variation, except perhaps on the voltage losses by activation, Eact , that appear in the potential balance. Normally, these losses are represented and described by the Butler–Volmer equation, but in some MEC models, they were oversimplified in an artificial way by representing them with the sinh1 ðcurrentÞ approach, proposed by

Microbial Electrolysis Cells for Biohydrogen Generation …

Noren and Hoffman [32]. Some models that have used this approach are depicted in [33–35]. This approach may work well for a limited set of operational conditions [32]. Nevertheless, for models with fewer restrictions, it may lead to mathematical problems (e.g., Differential Algebraic Equations’ (DAEs) models, imaginary solutions, or sudden discontinuities), which have no physical meaning. In more recent works, the Butler–Volmer equation has been used without any artificial simplification [36, 37], which does not only avoid the mathematical difficulties mentioned above but reduces the number of parameters involved as well [37].

3.3 Microbial Populations The first and simplest dynamic models (without spatial considerations) emerged at the end of the last century. In addition to the balances for reagents and products, they evolved in complexity in relation to microbial populations. In 1995, Zhang et al. [38] proposed a simple model only for a monobiomass–monosubstrate system. However, complex organic compounds may not be directly utilized by exoelectrogens, requiring the development of a microbial consortium to degrade such compounds into acetate and other molecules utilized by electroactive bacteria. This, in the following decade, two microbial populations present in the anode (i.e., anodophylic and methanogenic) were proposed [30, 39, 40], and subsequently, more microbial groups were included, i.e., fermentative, electricigenic, hydrogenotrophic [34, 41, 42] as well as different specialized microbial species for each type of metabolite [43]. These last two approaches, i.e., two-populations, multi-populations, continue to be applied for adequately representing the microbial population dynamics in the anode (see for instance [40, 44–47], respectively).

3.4 Membrane Phenomena The need to explain not only phenomena in the bulk phase but in the membrane itself has led

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researchers to propose 1D, 2D, and 3D models using different approaches, including simple multilayer models [34], and models that formally incorporate diffusive phenomena [39, 48, 49]. More recently, the use of effectiveness factors to capture all these diffusive phenomena was also proposed [50]. Comprehensive and extensive reviews of all these models can be found in [37, 51, 52]. However, the relative complexity of all these models depends on the use or the purpose, for which they will serve. In any case, the use of one or the other approach depends on the degree of complexity, with which a particular BES is wanted to model. For example, a detailed and indepth understanding of each one of the phenomena involved will require multi-population and multi-dimensional models. These can be useful for scaling-up or for monitoring purposes, but difficult to handle for control and optimization purposes, where time-only dependent dynamic models are preferred.

4

Experimental Lab-Scale Microbial Electrolysis Cells

Research on lab scale is one of the principal contributions today in MEC systems. Indeed, several reviews have been reported in the last decade concerning advances in lab-scale systems [53–56]. The contributions in this direction are mainly focused on studying the effects of reactor configuration, operating conditions, substrate, and electrode materials on hydrogen production and coulombic efficiencies [54–56]. Concerning electrode materials, one of the most common anode electrodes used is graphite and carbon due to low overpotential, low cost, easy availability, and chemical stability. Graphite (felt and brushes) and carbon (felt and cloth) are commonly used on lab-scale MEC systems [53]. For other materials, Cardeña et al. and Park et al. [17, 57] make an extensive and complete review of different materials used as both anode and cathode including the performance of each. Nonetheless, as mentioned before, the reaction producing hydrogen occurs on the cathode, and

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then, the cathode is an essential part of the MEC system. Indeed, the cathode material and the catalyst applied significantly affect the performance of the MECs [53, 58]. Platinum is commonly used as a catalyst in MEC systems, mainly due to low overpotential. However, there are two main disadvantages: the high cost and the negative environmental impacts associated with mining and extraction processes [53]. As an alternative, several other materials have been used. One of these materials is stainless steel due to its low cost, low overpotentials, and chemical stability. In addition to material, the size of the electrodes has been explored to study the influence of several architectures on individual electrode resistances and limiting currents by varying the applied potentials [59]. The application of nanotechnology has been a promising research area. Nanoparticles can improve both electrodes and the membrane performance as well as economic viability [60]. In general, the standard components of nanoparticles are carbon, metals, noble metals, polymer nanofibers, and some doping elements [60, 61]. Among all alternatives, graphene-based nanomaterials are considered as a promising approach. Graphene has exciting properties such as high electrical and thermal conductivities, superior specific area, high electromobility, and mechanical strength [62]. Unfortunately, nanoparticle components generally have expensive processing, purification, and raw resource costs [61, 62]. Thus, research groups have focused their efforts on developing different types of efficient and low-cost MEC cathodes or cathode catalysts for hydrogen generation [58]. These efforts include the use of natural or recycled raw materials and some nanoparticle alternatives already mentioned [61]. The catalysts commonly applied to the cathode are metals, materials (e.g., Tungsten carbide powder, nickel oxides (NiOx), electrodeposited Ni-based salts, among other), and nanomaterials [63]. However, it is considered that the microorganisms in the cathode will also act as a catalyst. This is commonly known as the concept of biocathode [53]. In recent years, several research groups have studied the application of

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cathodes catalyzed by microorganisms [63, 64]. While in MFCs biocathodes are generally applied to catalyze the oxygen reduction reaction or other spontaneous reactions under anoxic conditions (e.g., denitrification), in MECs, the biocathodes are applied to improve the efficiency of the hydrogen evolution reaction. Both—MEC and MFC are BES. Nonetheless, MFC uses enzymes and microorganisms that act as biocatalysts transferring electrons generated in their respiration to an external solid electron acceptor, to convert the chemical energy of a substrate directly into electrical energy instead MEC [65]. The start-up of MEC systems includes biofilm formation on the anode. Studying the combined effects of the used parameters at this stage is crucial to reducing the time of the biofilm growth. In this sense, initial substrate concentration and potential application are common key factors reported [66]. Once the biofilm is fully developed, several operating conditions like inlet substrate, feeding regimes, and concentration levels could be studied. Indeed, hydrogen production is strongly related to process operating conditions [23], and they should be studied to optimize the process and to enhance the energy efficiency [27, 67]. Concerning MEC microbial population, undesirable microorganisms should be avoided in some specific applications. For example, if the primary goal is only hydrogen production, the methanogenic population could cause a problem. Various techniques for eliminating and reducing methanogens activities in MECs have been discussed for waste remediation and simultaneous hydrogen generation [26, 68, 69].

5

Recent Achievements at Pilot Scale and Barriers to IndustrialScale Adoption

Nowadays, it is considered that MEC is a technology at readiness level five, which means that it is ready for industrial development [14]. In this sense, modeling studies, as they have been reviewed in Sect. 3, are a reliable alternative to ensure the implementation at pilot and industrial

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scale [52]. However, the scaling-up from laboratory to pilot scale and from there to industrial scale is not a simple task, and indeed, MEC is still mainly limited to lab-scale applications [70]. Furthermore, before even dimensioning, it is necessary to find the appropriate parameters to be considered as a point of comparison between different scales. The basic principles of scalingup in other biological reactors indicate that certain relationships remaining constant between them must hold. For example, in activated sludge wastewater treatment systems, it is typical to find scaling criteria such as Power per Volume unit, P/V, constant, volumetric mass transfer coefficient, kLa, constant, among others. However, in BES, to the best knowledge of the authors, similar parameters have not been established to homogenize scaling-up criteria. In this sense, Park et al. [57] point out that one of the great challenges is that there is not enough modular experience in enlarged MEC. However, in the absence of such parameters and criteria, different studies and works on the state-of-the-art have been able to identify some variables and certain aspects to be considered. For example, the size of the reactor has been pointed out as one of the main challenges when scaling MEC, since increasing the volume causes resistances and losses that seem to be inevitable. Park et al. [57] show that maximum hydrogen production rates of almost 10 L/(L.d) are relatively easy to achieve in small reactors of less than 100 mL, but when the size is increased, 1– 1000 L, reactors do not exceed a rate of 1 L/(L. d). In this sense, Leicester et al. [70] identified one primary challenge to face, i.e., the depth of the reactor, the solution of which would come from the field of engineering. Another aspect that has slowed down the comparative study of MEC at all scales is that hydrogen production yields are reported in different units that are not comparable to each other. Moreover, Rousseau et al. [71] point out that the volumetric production rate that is often used for assessing performance is not a good indicator if the area, and especially, the current density is not considered. Thus, current density is pointed out as the most important

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parameter to be considered to characterize and to compare MECs. Furthermore, they also point at the Faradaic performance (Ucath), which expresses the ratio of the number of electrons provided to the electrical circuit to the total number of electrons that could be extracted from substrate oxidation, as a useful tool for identifying operational problems in MEC and improving coulombic efficiency. Other important aspects that are usually highlighted are the configuration, e.g., “one chamber” versus “two chambers separated by a membrane”, as well as the architecture [54]. Regarding the configuration, most authors agree that if it is desired to produce only hydrogen, it is better to use the configuration of two separate chambers to prevent protons be consumed by hydrogenotrophic microorganisms eventually present in the bulk phase. However, the use of membranes can cause pH gradients and potential losses. For this reason, some authors highlight a single-chamber configuration as more effective [58]. In terms of architecture, most of the pilot MEC plants consist of anode–cathode planar panels arranged in series or in parallel. However, Rousseau et al. also point out cylindrical architecture [72, 73], where the cathode is in the center surrounded by the electrode on the cylinder wall, as more suitable than planar for reducing resistances. In terms of other parameters, and under the paradigm that an MEC plant must serve both to treat waste effluents and to produce energy, Leicester et al. [70] make a comparative study of 12 pilot plants, in which their characteristics compete with an activated sludge (AS) plant. In this study, eight parameters are identified as the most important to be considered for an effective pilot scaling. While these parameters are not declared in the form of constant scaling criteria, as P/V or kLa constants, maximum or minimum values are established instead. These values are chosen based on other studies (see references herein), and although a precise criterion for their choice is not established, at least it is the first attempt to meet these criteria in a single study and to give them a quantitative assessment. Such

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parameters are: (1) Feedstock Complexity, (2) Reactor Depth, (3) Organic Loading Rate (OLR), (4) Volume Treatment Rate (VTR), (5) Chemical Oxygen Demand (COD) Removal Efficiency, (6) Energy Balance, (7) Conductivity, and (8) Temperature. In their conclusions, they also point out that nowadays there is no individual MEC plant that has successfully performed at the required level against all parameters. Nevertheless, an extremely interesting point is that in the 12 plants studied, the energy balance turned out to be always positive or at least equal to that required in an AS plant. Authors note that an important economic aspect to consider in larger plants is the cost of electrode materials. Indeed, one of the main concerns, regardless of the practical implication of MEC systems, is to identify a highly efficient and low-cost cathode [53]. The cost of these materials can be too expensive or even prohibitive for scaling-up. In fact, in [57], it is noted that continuous evaluation, improvement and optimization of electrode materials in large-scale MEC play a key role in their commercialization. The good news is that electrode costs have gradually decreased. For example, Pant et al. [74] conducted a cost estimation study of BES plants and found a reduction of €4000/m3 of electrode compartment to just €1137/m3 between 2006 and 2010. The estimated costs certainly included those of the electrodes. Although today they are still expensive, it is expected that as technology advances, these will become more and more affordable. Finally, it should be noted that although the industrial scale of the BES in general is still far from being a reality [57], the production of renewable energy from wastewater is possible, if research in the future is properly directed [70]. For example, in the future, a better understanding of the complex microbial metabolic pathways will be necessary, whose solution would come from different areas such as microbiology, chemistry, and even molecular biology and metabolic engineering [75], always aided by electrochemistry [70].

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6

Future Directions on Microbial Electrolysis Cell Systems Development and Research

The first major challenge that this technology faces in the immediate future is to move from the lab and pilot scale to the industrial scale. Although most authors agree that the technology is not yet fully mature, many others also accept that it is just a matter of an extra effort in research to be able to make that step [10, 70]. In the previous section, several of the problems and drawbacks in this regard were discussed. Ohmic and potential losses must be reduced to obtain acceptable coulombic efficiencies on the industrial scale. The losses are related to the use of membranes, the conductivity of the medium, the materials used in the electrodes and the architecture itself. Materials science and nanotechnology can provide a solution in terms of new materials in membranes and electrodes, while the configuration and architecture of reactors must be studied and improved from a more systematic scaling approach than is currently developed, in which constant criteria are established to be maintained between the different scales. Moreover, the mechanisms of electron transfer between microorganisms and the anode are not yet entirely clear. It is necessary to do more research in this regard not only to clarify these mechanisms, but to find ways to favor the one or those that are more efficient. This could be a function of the materials used in electrodes or even the type of feedstock used [76]. The latter could involve different metabolic pathways and, therefore, influence the electron transfer mechanism. In addition to this, it is also necessary to advance faster and more efficient starting protocols. Both microbiology and biochemistry still have a very important role to play in these areas. In terms of dynamical modeling, although most authors agree on modeling microbial growth kinetics in the anode with Nernst–Monod type expressions, it is also possible that the kinetic constants present in such expressions are different, when bacterial populations are subjected to

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electric fields than when they are not. Even the kinetic expressions themselves could be different. Bioengineering is the science that can answer this question. As progress is made in all these aspects, there will be better and more useful models for each specific application. This can also lead to the implementation of different approaches in control and optimization that have already given excellent results in other bioreactors such as anaerobic digestion, dark fermentation, activated sludge, among others, which in turn will further encourage researchers and society, in general, to seek faster progress in the development of this novel and promising technology. By using biocathodes, it is also possible to sequestrate the CO2 produced in any other biological process to produce value-added chemicals, e.g., methane, acetate, formate, ethanol [6, 77, 78], as well as to recover nutrients and metals [79, 80]. Also, by implementing integrated systems that combine AD, DF, MFC, and MEC [6, 79] and under a biorefinery concept [76], these systems have already been proposed for obtaining a maximum of bioenergy available in organic waste and wastewater in the form of electricity, hydrogen, and methane [81, 82]. For instance, recent studies showed that MEC-AD in a single bioreactor could accelerate the degradation of the substrate [83–86]. The main feature is the alteration of the AD microbial population by simultaneously enriching the exoelectrogens and methanogens [86]. This dual application has recently gained attention due to the energy efficiency and economic benefit [87]. Finally, it is important to remark that BES in general is already being used in applications other than the production of only one by-product (like hydrogen) from waste and wastewater. However, the study of the integration of MEC with other bioprocesses, particularly with other BES systems for value-added applications, is an emerging area of research [54]. For example, two or more BES processes can provide an alternative for cost reduction (i.e., MFC-MEC coupled systems) [26]. For specific chemical removal applications (i.e., nitrogen, metal, and salinity

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removal), combined microbial desalination cell with MEC system could be a feasible alternative [88]. It is important to remark that coupling MEC systems are not limited to another BES process. As an example, simultaneous dark fermentation and MEC system in a single bioreactor for hydrogen production from microalgae have been reported [89]. Moreover, the coupling idea is not limited to the simultaneous interaction at microbial level. From a self-sustainable point of view, it is possible combining conventional MEC system with advanced carbon–neutral technologies to provide external power source (i.e., solar, microbial, osmotic, or thermoelectric power) [90]. Without doubt, the MEC system coupled with an additional bioprocess is a promising research area. In this direction, innovative principles applied to bioprocess and equipment design could be considered. As an alternative, the process intensification, that has been successfully applied to chemical and petrochemical research areas, could be a powerful design tool in future MEC coupled systems.

7

Conclusions

Microbial Electrolysis Cell Systems for hydrogen production are a promising research area on biofuels, while being a sustainable alternative to mitigate the so-called climate change, since being based on waste and wastewater treatment as source of feedstock. Integrated with other biochemical and biotechnological systems, it can be also an important approach for nutrient recovery, metals removal, bioremediation, and added value metabolites. Scientific fundamentals upon this approach are advanced enough today, but it is still necessary a more intensive technological contribution on material sciences, bioengineering, and modeling for being a real industrial reality. The achievements and advances obtained so far leave no doubt that this technology can become in the very near future a real alternative to contribute to the growing energy demand and to move toward a more sustainable development.

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Novel Membrane Technologies in the Treatment and Recovery of Wastewaters Mehmet Emin Pasaoglu, Recep Kaya, and Ismail Koyuncu

from the pressure prerequisites, are the pore sizes of the membranes. Considering the membrane studies in the literature, such pressure-driven membrane technologies have long been implied with a variety of scenarios for wastewater recovery. Besides the pressure-driven membranes, innovative hybrid water recovery solutions rely on concentration, electrical potential, thermal difference, and vacuum-driven membrane processes. The advanced membrane processes that can be explored are pervaporation (PV), forward osmosis (FO), membrane distillation (MD), electrodialysis (EDI), membrane bioreactors (MBR), and a combination of these technologies to be used in zero-liquid discharge systems for wastewater recovery and reuse.

Abstract

In residential areas with limited water supplies or unstable water sources, water reuse was first suggested more than 20 years ago. Pollution, increasing global urban population, and climate change have all had an impact on sustainable water supplies, increasing the demand for efficient wastewater reuse and recovery technology. Wastewater reuse and recovery can be applied with different membrane technologies. The most extensively used membrane applications in the treatment of wastewater and recovery from pretreatment to post-treatment stages are pressure-driven membrane processes. These approaches rely on hydraulic pressure to create separation. These procedures are divided into four categories. The four techniques are microfiltration (MF), ultrafiltration (UF), nanofiltration (NF), and reverse osmosis (RO). The fundamental differences between all these techniques, apart

Keywords




1 M. E. Pasaoglu
R. Kaya
I. Koyuncu (&) Civil Engineering Faculty, Environmental Engineering Department, Istanbul Technical University, Sariyer/Istanbul, Turkey e-mail: [email protected] National Research Center on Membrane Technologies (MEM-TEK), Istanbul Technical University Ayazaga Campus, Sariyer/Istanbul, Turkey




Electrodialysis Forward osmosis Hybrid membrane processes Membrane distillation Wastewater treatment and recovery







Introduction

Water is essential to all human activity. Millions of m3 of wastewater are created daily as the human population grows in agriculture, households, and industrial sectors. However, freshwater resources cannot be renewed to meet the requirements of this continuously increasing

© The Author(s), under exclusive license to Springer Nature Switzerland AG 2023 E. Debik et al. (eds.), Wastewater Management and Technologies, Water and Wastewater Management, https://doi.org/10.1007/978-3-031-36298-9_7

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population. Moreover, there has been a fierce rivalry and an unequal allocation of scarce freshwater resources among the various industries. As a result, many people worldwide, particularly in developing countries, have lack of access to safe drinking water. Evidence of these scenarios may be seen throughout the world, particularly in the Middle East, Africa, Asia, and Latin America [1]. Over the years, many efforts have been made to implement different wastewater treatment technologies, such as conventional filtration, coagulation-flocculation, and biological treatment systems. Existing technologies are also being improved to meet current discharge or reuse regulations. Membrane technology is one of the wastewater treatment technologies that has witnessed significant growth during this time. Membrane technology has significantly expanded in the last several decades due to its advantages in water and wastewater treatment. Membrane technology offers several opportunities in wastewater treatment due to considerable reductions in equipment size, energy requirements, and affordable capital costs [1, 2]. Figure 1 shows a scheme of different membrane-based processes. There are different types of membrane processes in the literature, but in this chapter, the authors will detail the most novel technologies in wastewater treatment and reuse applications.

Fig. 1 Schematic representation of membrane processes [3]. MF: microfiltration; UF: ultrafiltration; NF: nanofiltration; RO: reverse osmosis; PV: pervaporation; GP: gas permeable; ED: electrodialysis

M. E. Pasaoglu et al.

These novel processes can be specified starting with the forward osmosis (FO) process. FO is based on the conventional osmosis system, in which H2O molecules are transferred from one solution to another one over a semipermeable membrane. FO has been used to treat and concentrate a variety of wastewater streams. ED and reversal electrodialysis (EDR) are techniques that extract dissolved ions from water via the combination of electricity with ionexchange membrane technology. An electric potential is used in these processes to transport ions from a diluted solution to a concentrated one across an ion-exchange membrane. PV separation is a membrane combined with evaporation primarily used to separate liquid mixtures. This technique is generally used for the separation of ethanol and water. PV has been investigated in a range of sectors for wastewater treatment. There are also hybrid membrane processes combining different technologies to make more benefit from the water recovery. One example of those hybrid systems is the combination of forward osmosis with reverse osmosis (FO/RO), introduced for wastewater treatment and desalination of seawater simultaneously. Another method is the membrane bioreactor for industrial and domestic wastewater treatment and reuse. To treat wastewater or to recover resources from

Novel Membrane Technologies in the Treatment and Recovery of Wastewaters

wastewater, a membrane bioreactor can be described as a combination of a biological process, such as activated sludge, with membrane technologies with MF, UF, NF, and other types. Concentrated streams are difficult to recover for wastewater recovery systems, especially in industrial applications. Membrane distillation (MD) exhibits growth in membrane technology that treats concentrated streams as part of zeroliquid discharge systems. This hybrid membrane process has been used for over 50 years with limited progress toward large-scale or commercial applications [4]. MD is the process of using heat to separate compounds depending on the compound’s volatility. MD is a process in which water vapor is delivered onto a microporous hydrophobic membrane by utilizing the vapor pressure difference across the membrane. Membrane distillation consists of four types that include vacuum membrane distillation (VMD), direct contact membrane distillation (DCMD), sweeping gas membrane distillation (SGMD), and air gap membrane distillation (AGMD). Recently, primarily hybrid combinations have been focused such as liquid gap membrane distillation (LGMD) and thermostatic sweeping gas membrane distillation (TSGMD). However, recent MD studies have concentrated on hybrid combinations such as LGMD and TSGMD [1].

2

Forward Osmosis (FO)

In the coming decades, water scarcity will substantially impact civilization, ecological systems, food security, and environmental sustainability. It might also represent a severe danger to economic growth [5]. If existing trends of water consumption continue, most of the world’s population, approximately two-thirds, are estimated to suffer from water scarcity by 2025 [6]. Therefore, it is pretty concerning when water is used for issues other than sustaining life, such as industrial production activities [7]. Desalination, waterless technologies in industrial operations, reservoir water storage, the protection of wetlands, and other approaches are among those used to overcome water shortage.

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Reusing wastewater using membrane-based or pressure-driven filtering processes is one viable solution to worldwide water scarcity [8]. RO technology is considered a highly efficient and broadly applied membrane technology for treating wastewater nowadays [9]. Despite its popularity, RO brings several drawbacks, such as CO2 emissions, irreversible membrane fouling, brine management, high cost, and the necessity of considerable pretreatment [9, 10]. Furthermore, because significant hydraulic pressures are used, RO cannot directly treat extremely salty streams and is energy intensive [11]. FO, a novel osmotic-driven membrane process, has recently piqued the interest of academics and scientists worldwide as an emerging membrane process and a surrogate for the RO process. Because FO occurs naturally and does not require hydraulic pressures, it offers a significant advantage over other pressure-driven membrane processes [12]. Some disadvantages of draw solutions include lack of highly selective membranes, reverse solute diffusion, determination of a proper draw solution, concentration polarization (CP), membrane fouling, and particularly, the regeneration of the draw solution. FO process flux prediction was designed according to the models [13]. These models changed in time to incorporate new physical variables showing the growing insight into the osmosis flow phenomena in FO membrane processes. A model which is considered the first flux model for pressure-retarded osmosis (PRO) coexisting with CP was published by Lee et al., in which FO and RO values were used to forecast PRO performance (1981) [14]. Later, McCutcheon and Elimelech (2006) developed this model for the FO process [15], and they made additional modifications to this model to include the impact of CP (Fig. 2) on flux behavior. When the draw solution is directed toward the active layer while the feed solution is directed toward the support layer (called PRO mode), the effects of dilutive external concentration polarization (DECP) and concentrative internal concentration polarization (CICP) show a combined effect. In this case, flux is predicted as given by Eq. (1)

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Fig. 2 Solute concentration profiles in a forward osmosis mode and b pressure-retarded osmosis mode through a TFC membrane at steady state. Reprinted with allowance of Elsevier from Bui et al. [16]

 
 Jw JwPRO ¼ A pD;b exp   pF;b expðJw K Þ : k ð1Þ Water flux calculation is performed by using the FO membrane equation, with the support layer facing the draw solution and active layer facing (AL-FS) the feed solution (FO mode) (2). 
 Jw JwFO ¼ A pD;b expðJw K Þ  pF;b exp  : k ð2Þ Here, k represents the mass transfer coefficient— the level of internal concentration polarization (ICP)—while K indicates the rate at which a solute may diffuse through the support layer [15]. While considering ICP and external concentration polarization (ECP), the calculations mentioned above neglect the effects of reverse salt diffusion (RSD) and are only accurate if the FO membrane is rejecting solutes.

The impacts of RSD on the FO membrane are not considered in the McCutcheon and Elimelech [15] model. Tan and Ng [17] discovered that the model defined by McCutcheon and Elimelech [15] overpredicts water flow for concentrations of draw solution higher than 1.0 M sodium chloride (NaCl). Yip et al. [18] created a model for estimating water flux in the FO process considering the RSD (draw-to-feed-solution) as well as internal and external concentration polarizations: (

FO JW

  ) pD;b expðJW K Þ  pF;b exp JkW    ¼A 1 þ B=JW exp JkW  expðJW K Þ (

PRO JW

ð3Þ

)   pD;b exp  JkW  pF;b expðJW K Þ    : ¼A 1 þ B=JW expðJW K Þ  exp  JkW

ð4Þ As is apparent from the denominator, Yip et al. [18] assessed the effects of both internal and

Novel Membrane Technologies in the Treatment and Recovery of Wastewaters

external concentration polarization; however, the explanation of mass transfer resistance at the porous support layer was missed in the formula. The impacts of ECP on the porous support side of the membrane have gained importance and cannot be neglected when the membrane is operated at higher water fluxes or slow water flow velocities [19, 20]. FO JW

h

i

 9 8 < pD;b exp JW k1 þ S=DD  pF;b exp JkW = D F n

 h

io ¼A :1 þ B=J exp  JW  exp J ; 1 S W W KD þ DD kF

ð5Þ PRO JW

 h

i 9 8 < pD;b exp  JkW  pF;b exp JW k1 þ DS = D F F n h



 io : ¼A JW :1 þ B=J exp J ; 1 S W W KF þ DF  exp  kD

ð6Þ DD is the solution’s draw diffusion coefficient, and S is the parameter concerning the membrane structure. Technically, less than a 10% drop in water flow throughout the FO process may be related to the influence of ECP at the porous support layer [16, 21]. The FO process also incorporates salt flux, which passes backward from the draw solution to the feed and the water flux. All osmotically driven membrane processes have reverse salt flow, which has a negative impact on the membrane performance in the FO process [22]. The following equation can be utilized to determine the reverse salt flux in the FO process [23]. "

   # CDb exp  JkW  CFb exp JWDS     : ð7Þ Js ¼ B 1 þ JBW exp JDw S  exp  JkW CDb and CFB indicate the draw and feed solution’s bulk concentrations, respectively. When specific draw solutions are utilized, RSD can contaminate the feed solution, reduce the net driving power across the membrane, increase draw solution loss, and promote membrane fouling. To decrease fouling, the FO membrane is essential in wastewater treatment. Figure 3 depicts the proportion of various types of membranes employed in wastewater treatment

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experiments. Authors’ research indicates that nearly half of the experimental work used cellulose triacetate (CTA) FO membranes to treat wastewater because of their high chlorine resistance, resilience to biological deterioration, and low fouling potential [24, 25]. CTA membranes have significant drawbacks despite their various advantages, including high salt chloride permeability, low pH tolerance, and low water permeability [13, 26, 27]. Thin-film composite (TFC) FO membranes are competitive due to their higher water permeability than CTA membranes [27, 28]. Several companies, such as HTI, Porifera, and Oasys Water Inc., supply TFC FO membranes with high rejection rates for nitrates, silica, and organic compounds [29]. TFC membranes, unfortunately, exhibit easy-to-foul properties in wastewater treatment. Additionally, since TFC membranes are open to chlorine attack, various studies with complicated wastewaters integrating CTA membrane, which can survive up to 1 ppm of chlorine residues, were conducted [30, 31]. Furthermore, due to the high-density properties of carboxylic acid functional groups on their surface, TFC membranes are prone to fouling. In contrast, CTA membranes are more resistant to gypsum and silica scaling than TFC membranes [32]. Fouling remains a significant operational issue in membrane processes, even though membrane technologies present more benefits than proven methods in wastewater treatment [33]. Membrane fouling significantly impacts water flux and may have long-term effects that affect the membrane. In the FO process of treating wastewater input solutions, colloidal membrane fouling and organic, inorganic, and biological fouling have been discovered [9, 34–38]. Researchers have created fouling mitigation strategies for the FO, including chemical treatment, modifying membrane orientation, osmotic backwash, air scouring, ultrasonic cleaning, and increased cross-flow velocity. The suggested approach for membrane fouling mitigation was not often applicable in recovering water flux at the FO process, particularly in real wastewater media.

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Fig. 3 Fraction of various FO membranes utilized in various FO wastewater applications

3

Pervaporation (PV)

This process separates liquid mixtures based on preference via evaporation and membrane permeability. On one side of the membrane, the liquid mixture is seen in Fig. 4 being supplied; on the other side, the permeate is evaporating. During this stage, the permeate is absorbed upward. This way, the more permeable liquor component is adsorbed on the membrane (molecularly porous inorganic or non-porous polymeric membrane). Then, under a diffusing species concentration gradient, these components diffuse across the membrane and evaporate at the membrane’s downstream phase. Eventually, the liquid is recovered following the condensation of the vapor obtained. Mass is known to be transferred across the membrane via the solutiondiffusion method [39]. The primary use of this technology is the separation of ethanol from water. However, several companies are looking into it for the wastewater treatment process. Pervaporation was employed for the micro-irrigation of plants using

wastewater by Quiones-Bolaos et al. [40]. In order to perform the experiment, certain spots in the soil were covered with a thick hydrophilic pervaporation barrier. Synthetic wastewater was cycled over the membranes in a feed tank, while the permeate flow and wastewater enrichment (contaminant rejection) were observed. Based on the results, this technology may be used to treat wastewater or groundwater, brackish, for microirrigation.

Fig. 4 Schematic diagram of pervaporation [1]

Novel Membrane Technologies in the Treatment and Recovery of Wastewaters

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Table 1 Pervaporation applications for the removal of specific contaminants Application

Results

Reference

Removal of toluene from aqueous solution

42% removal

[43]

1.0 mol of aqueous volatile organic compound (diethyl ether, acetonitrile, ethyl acetate)

90.35%* removal

[44]

Methyl tert-butyl ether removal from aqueous solution

95% removal

[45]

Removal of 0.5% by weight of pyridine from water

Effective removal stated

[46]

Removal of 0.39% by weight of Isopropyl acetate from aqueous solution

Effective removal stated

[47]

Removal of 0.1–0.4% by weight of chlorophenol and phenol from aqueous solution

Effective separation stated

[48]

*

wt% = Percentage by weight

Wijmans et al. [41] employed membranes, which are 100% organophilic, to extract and concentrate organic solvents (ethyl ether, toluene, benzene, butane, naphtha, etc.) from dilute aqueous streams in a small-size experiment. It was discovered that the organic solvents might be concentrated at least 50–100-fold, resulting in a cleaner flow of effluent to discharge or reuse. Similarly, Kondo and Sato [42] employed an aromatic hydrocarbon selective polyether block amide (PEBA) membrane to extract phenol from industrial effluent discharged from a process of phenolic resin production. Less than 10% phenol and other pollutants were found in the effluent. Following the studies, the phenol amounts observed were less than 300 mg/L. Pervaporation’s unique properties make it suitable for removing certain pollutants. Table 1 depicts some target uses of pervaporation in wastewater treatment. Owing to their specific function, pervaporation membranes are mainly developed to exhibit a better affinity to the component to be separated. This means the membrane structure and chemical composition are essential in obtaining the desired separation [49]. The pervaporation process is affected by the temperature, feed concentration, partial pressure, and feed stream velocity [50]. Pervaporation is an energy-saving and ecologically advantageous process in addition to its potential for separating fluid blends when

standard separation procedures are constrained. However, there are several drawbacks to this technology. Due to the very delicate working conditions, large industrial applications have yet to be realized. Again, a lack of suitable membranes and their high price restrict the application of pervaporation beyond dehydration [1, 51].

4

Electrodialysis (ED)

To extract salt from a feed stream into a concentrate stream, electrodialysis uses direct current (DC) power to transport ions across selective ionexchange membranes, leaving behind a highervalue product. This method is employed in various industries and applications, including demineralization of the dairy industry-whey demineralization and sugar, desalting of glycerin and amine, and deacidifying of juice. General benefits of the ED process are given below; • Increases the value of the product, produced in the manufacturing process, lowering the cost of processing, and getting value out of waste streams. • Merely takes away ionized species, leaving behind valuable components. • Compares favorably to ion exchange in salt effluent and chemical use.

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• Because operations may be adjusted to match any product set point, wasteful energy, time, and product are eliminated. • Uses a skid-mounted, mechanical design that works in batch or continuous mode to operate efficiently and dependably. Some applications of the ED can be listed below; • Dairy: whey and permeate demineralization, • Demineralization of dextrose and other cornderived sugars, • Wine stabilization, • Glycerine desalting, • Brine concentration and wastewater recovery, • Removal of heat stable amine salts, demineralization of glycol solutions. A simple scheme of the ED process is given in Fig. 5. Ion movement is used in electrodialysis, a sophisticated membrane technique, to desalinate water. The ions flow across membranes used in selective ion exchange that only permit the

Fig. 5 Simplified scheme of the ED system [52]

M. E. Pasaoglu et al.

passage of cations or anions. It produces two distinct streams: a desalinated stream (dilute) and a concentrated salt stream (brine) by switching these membranes with spacers in between [27, 28, 52, 53]. Salt ions are moved via ion-exchange membranes from one solution to another using ED, which is influenced by an applied electrical potential. This is carried out in an arrangement known as an electrodialysis cell. To create the feed (dilute) and concentrate (brine) compartments, two electrodes are placed between an anion-exchange membrane and a cationexchange membrane, respectively. With the alternate anion and cation-exchange membranes created, the numerous cells of electrodialysis are organized as a composition known as an electrodialysis stack in almost all practical electrodialysis operations. Other membrane-based processes such as RO and distillation differ from electrodialysis procedures in that dissolved ions are transported away from the feed stream rather than toward it [54, 55]. In an electrodialysis stack, the electrode (E) stream, the brine or concentrate (C) stream,

Novel Membrane Technologies in the Treatment and Recovery of Wastewaters

and the dilute (D) input stream are all permitted to flow through the proper cell compartments made by the ion-exchange membranes. The electrical potential difference causes the negatively charged ions (e.g., Cl−) in the diluted stream to flow in the direction of the positivecharged anode. These ions pass through the positively charged anion-exchange membrane, but they cannot proceed in the direction of the anode due to the negatively charged cationexchange membrane. As a result, they continue to circulate in the anion’s concentrated C stream. The charged, which is a negatively cationexchange membrane, allows the positively charged species (such as Na+) in the D stream to pass through as they move toward the cathode [56, 57]. The cations remain in the C stream because the positive charge, the anion-exchange membrane, prevents them from moving further toward the cathode [57]. The migration of cations and anions causes an electric current to flow between the anode and cathode. Only an equal number of cation and anion charge equivalents from the D flow are transferred into the flow of C to maintain the charge balance in each stream. The concentration of ions in the C flow rises overall due to electrodialysis, whereas ion concentration in the D solution input stream decreases [58]. The electrode stream or E stream traverses each electrode in the stack. This stream may be a different solution for different species, like Na2SO4, or it could have identical content as the influent, such as NaCl [52]. Depending on the stack configuration, anions and cations can be transported from the E stream to the carbon stream, or anions and cations from the D stream may be transferred into the E stream. In each instance, this transport is required to move current across the stack and keep stack solutions electrically neutral. Reactions occur at each electrode in electrodialysis [54]. At the cathode, 



2e þ 2H2 O ! H2 ðgÞ þ 2OH ; while at the anode,

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H2 O ! 2 H þ þ 1=2 O2 ðgÞ þ 2e or 2 Cl ! Cl2 ðgÞ þ 2e : A small amount of hydrogen gas is produced at the cathode. Depending on the E stream’s composition and the placement of the end ionexchange membranes, few amounts of oxygen or chlorine gas are produced at the anode. The E stream effluent from each electrode compartment is mixed to maintain a neutral pH and discharged or recirculated to a separate E tank. These gases are typically released later. However, others have suggested gathering gas of hydrogen to use it in the creation of energy, for instance. Efficiency in the electrodialysis process can be calculated by Spiegler and Laird [59]. The effectiveness of ions being transported across ion-exchange membranes for a specifically applied current is measured by the current efficiency. In commercial stacks, current efficiencies >80% are often preferred to save energy running expenses. Low current efficiencies suggest the possibility of back-diffusion of ions from the C to the D, shunt currents between the electrodes, or water splitting in the D or C streams. Current efficiency is calculated as:  d  d zFQf Cinlet  Coutlet e¼ NI

ð8Þ

with e z F Qf d Cinlet d Coutlet

N I

current utilization efficiency ion charge constant of Faraday, 96,485 A-s/mol dilute flow rate, L/s dilute electrodialysis cell influent concentration, mol/L dilute electrodialysis cell effluent concentration, mol/L cell pairs number current, A.

Current efficiency is generally a function of feed concentration. ED and RO are the most frequently used membrane technologies for desalination and water treatment (deionization). Figure 6 presents

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a

M. E. Pasaoglu et al.

a

Fig. 6 Overview of ED and RO. a In RO, pressure is used to push water through a membrane while most of the solutes are retained on the retentate side. b In the ED,

current flows through channels and membranes generated by electrodes placed at either end of a stack containing many cells [60]

a brief schematic of both techniques. Desalination and deionization remove ions by removing salts, whereas water treatment often refers to removing impurities other than salts, such as organic micropollutants (OMPs). The pressuredriven RO process keeps the bulk of the ions and other solutes on the retentate side of the membrane, while freshwater is created on the permeate side. A complementary method to RO called NF employs membranes with bigger pore sizes than those used in RO and lower pressures. The retention of monovalent ions in NF is considerably lower than that of divalent ions, allowing for the selective removal of divalent ions [60]. In ED, ion-exchange membranes (IEMs) are placed next to thin channels of water, and when a current is supplied, the ions are drawn through the IEMs and into the other channels (Fig. 7). The growing world population and the consequent increase in the need for non-traditional water sources (such as salt water, brackish water, produced water, and wastewater), both potable and non-potable, will need to be included as elements of our supply of future water portfolio [62]. Multistage treatment trains, which are required to treat these unusual and occasionally severely degraded waters, now frequently include membrane technologies. For instance, non-porous membranes, which are RO powered by pressure, have been widely utilized to treat salinity fluids that are high concentrations. They have the potential to reject up to 99.5% of the salt in synthetic seawater while using ten times less

energy than thermal systems. NF is another popular membrane separation technology method, with somewhat more porous membranes and fewer operating pressures than RO. NF has been operated in the treatment of drinking water for both water softening (the removal of Ca+2 and Mg+2) and the removal of natural organic matter (NOM) [62]. Another effective pretreatment method is ED, which uses IEMs to extract charged solutes from feed fluids and has been used to desalinate saline and brackish waters for over 50 years. Although electrodialysis has been proven to eliminate boron from brackish water, seawater, and wastewater at high pH levels, borate is less mobile than other anions that coexist, such as chloride and sulfate. ED at circumneutral pH can be utilized as a part of an NF/RO pretreatment strategy to remove anions that are charged and cations, excluding separating the neutral boric acid species. Because the pretreatment of ED reduces the ionic composition of the NF/RO water of feed, the pH may be increased, and borate can be denied in an NF/RO process with decreased hydraulic pressure needs and less possibility of inorganic scaling [63]. Landsman et al. integrated the ED process with the NF/RO process for the removal of boron in a more efficient way. Their example schematic can be found in Fig. 8 [62]. It is shown in their study that ED pretreatment promotes rejection of boron by NF/RO through improving membrane-solute interactions to illustrate repulsion of electrostatic and by enhancing the speciation of charge, which is

Novel Membrane Technologies in the Treatment and Recovery of Wastewaters

Fig. 7 ED stack in an exploded perspective with its primary components identified. (1) anode; (2) cathode; (3) steel frame; (4) plastic end plate; (5) inlet anode compartment; (6) anode chamber; (7) inlet cathode compartment;

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(8) cathode chamber; (9) inlet‐concentrating compartment; (10) inlet‐diluting compartment; (11) cation‐exchange membrane; (12) spacer‐sealing frame; (13) spacer net; (14) anion‐exchange membrane; (15) screws [61]

Fig. 8 Illustration of the planned ED-NF/RO system [62]

negatively boron because of a decline in complexes of salt-borate [62]. Additionally, ED pretreatment lowers the osmotic pressure of the NF/RO water of feed, improving the flow of the NF/RO permeate and/or enabling the use of EDNF for high-salinity waters that would commonly need RO. Different amounts of total carbonate, alginate, and calcium in the NF/RO water

of feed affected how fouling was mitigated, how the fouling layer was made up, and how well the system worked overall. Calcite and alginate are two (in) organic substances that interact in ways that affect NF/RO membrane fouling. In an integrated treatment train, improving both the ED and NF/RO processes can increase the system’s adaptability and resilience while

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desalinating non-traditional waters. The technological benefit of this hybrid system of EDNF/RO depends on several variables, including the feed water composition (e.g., pain of ion) and treatment objectives [62]. Wang et al. studied white carbon black wastewater treatment by ED concerning silicon sol transport, salt separation, and wastewater recycling. White carbon black (WCB) is employed extensively in various industrial processes as a significant silica product. Silica and sodium sulfate-containing effluent are created in huge quantities during the manufacturing of WCB using the precipitation process. It is crucial and essential to develop effective and affordable wastewater treatment technologies to extract sodium sulfate from wastewater as a value-added by-product and return this wastewater to the WCB process. This study treated WCB wastewater using the ED technique, a tried-and-true salt-separation technology. The presence of silica in the wastewater makes its ED treatment more challenging [63]. Hussain et al. investigated acid recovery from molybdenum metallurgical effluent with the usage of selective ED and NF. The molybdenum mining industry produces a massive quantity of acidic wastewater including, copper, molybdenum, and different non-precious metals. Metal ion separation and sulfuric acid recovery from molybdenum metallurgical wastewater were investigated in this study using selective nanofiltration and electrodialysis (SED). The comparison of NF and SED indicates that SED is more competitive than NF for this wastewater treatment because it consumes less energy, has a greater acid recovery rate, and has a higher purity for the recovered acid. Furthermore, combining NF and SED might achieve an acid recovery rate of more than 94.2% [64]. Lin et al. [65] investigated an integrated loose ED-NF technique to obtain sustainable resources from highly brackish textile effluent. An integrated loose ED-NF method was investigated in this work for the simultaneous recovery of coloring agents, sodium chloride, and clean water from high-saltiness textile effluent, thereby completing the material loop and decreasing waste emission. A loose membrane, NF (weight

M. E. Pasaoglu et al.

representing a molecular cutoff of 800 Daltons), has been proposed to fractionate dyes and sodium chloride in high-salinity fabric effluents. The coloring agent was recovered from highsalinity textile wastewater using a nanofiltrationdiafiltration (NF-DF) unit that included a preconcentration phase and a constant-volume diafiltration phase, and it was enriched by a factor of 9.0 from 2.01 to 17.9 g/L with 98.4% purity. The sodium chloride concentration and clean water were efficiently retrieved from the salt-containing permeate resulting from the loose NF-DF, thanks to the following application of ED. At the same time, the produced clean water was recycled to the NF-DF unit [65].

5

Membrane Distillation (MD)

A microporous hydrophobic membrane is used in the membrane distillation (MD) process to separate two aqueous solutions at various temperatures. A gas–liquid interface is generated because the membrane’s hydrophobicity limits the mass transfer of the fluid. The volatile ingredients in the supply blend evaporate through the pores (10 nm–1 µm) and are transported from the section with high pressure of vapor to the section with low pressure of vapor, where they are condensed in the cold fluid/phase of vapor via diffusion and/or convection in consequence of the pressure of vapor difference led by the temperature gradient on the membrane. Supply solutions that solely include non-volatile components, for example, salts, will carry water vapor through the membrane, resulting in demineralized water on the part of distillation and a more concentrated flow of salt on the part of the supply. The specific module layout controls how the membrane’s vapor pressure difference is produced. Direct contact membrane distillation (DCMD), the most common kind, uses a condensation liquid (typically clean water) directly contact via the membrane to create the permeate side. A cold, inert sweep gas can also release the evaporated solvent. An option is to collect the evaporated solvent on a surface that is separated from the membrane by an air gap (AGMD) or a

Novel Membrane Technologies in the Treatment and Recovery of Wastewaters

vacuum (VMD) and sweeping gas membrane distillation (SGMD), respectively. Vapor molecules condense away from the membrane module. The total preservation of non-volatile elements like ions, macromolecules, and colloidal particles is theoretically possible due to the kind of driving force and the membrane’s capacity to reject water. The potential benefits of MD over traditional separation methods are principally due to its lower operating temperature and pressure, which results in cheaper energy expenditures and fewer demanding mechanical requirements. Against distillation and RO, supply solutions can be separated at a temperature much under the point of boiling and at atmospheric pressure. Average supply temperatures between 30 and 60 °C enable the use of renewable energy sources like solar, wind, and geothermal energy, as well as the recycling of waste heat fluxes. Since a higher matter concentration is attained on the supply side, MD is also less impressionable to concentration polarization-related flux constraints than RO. In theory, MD permits full retention for nonvolatile substances which are dissolved and limitless supply concentrations.

Fig. 9 Direct contact membrane distillation (DCMD), air gap membrane distillation (AGMD), and the derived “feed gap” version feed gap membrane distillation

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MD provides the fundamental advantages of membrane separation over conventional distillation, including simple operations, easy scaling up, the use of high membrane surface-to-volume rates, and the capacity to process flows containing heat-sensitive components and/or a high concentration of suspended particles at atmospheric pressure and temperatures under the point of boiling of the supply. There are several distinct MD methods. The configuration of their distillate channel or how this channel is managed is the key difference between the fundamental four approaches. The most popular technology are as follows: • • • • •

Air gap membrane distillation (AGMD) Direct contact membrane distillation (DCMD) Permeate gap membrane distillation (PGMD) Sweeping gas membrane distillation (SGMD) Vacuum multi-effect membrane distillation (V-MEMD) • Vacuum membrane distillation (VMD). Simple illustration of MD processes is given in Fig. 9.

(FGMD) and feed gap air gap MD are the basic channel designs (FGAGMD). Adopted from [66]

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Table 2 Implementations of membrane distillation for treatment of wastewater and recovery MD configuration

Types of membrane

Feed Solution

Purpose

DCMD

PTFE

Simulated water

Removal of chromium

DCMD

TF200

Distilled water and humic acid

Solution of humic acid’s treatment

DCMD

Hydrophobic ally modified FS PS or PES

Wastewater of synthetic radioactive

Removal of radioactive elements

DCMD

PTFE

Persian Gulf’s seawater

Desalination

DCMD

PP

Zablocka hot salty water

Concentration of brine

DCMD

PVDF

Sea water

Boron removal

DCMD

PTFE

Wastewater from olive mill

Concentration of olive mill wastewater

DCMD

FS PP

Cooling tower blowdown water

Desalination

SGMD

FS PTFE

Wastewater of diluted glycerol

Glycerol's concentration

SGMD

PP

Distilled water, liquid form of NaCl

Desalination

DCMD

PTFE

Produced Water

Desalination

DCMS

FS PTFE

Wastewater of olive oil mill

Phenolic compounds’ concentration

AGMD

FS PTFE

Produced Water

Desalination

DCMD

PVDF

NF’s and RO’s retentate

Enhancement of RF of water and crystallization of salt

VMD

PP and FS PVDF

Wastewater consists of arsenic

Arsenic removal

VMD

PTFE

Human urine

Regeneration of water

Table illustrated from [67]

One of the main obstacles to the industrialization of MD for seawater desalination applications is the absence of an appropriate membrane that could display long-term agreeable salt rejection and flux. As a result, this section goes into great length into the requirements for a membrane’s use in MD. The first step is to go through the material selection, fabrication, and characterization methods. Appropriate values or scope to main influencing criteria are supplied in every category. The impact of every criterion on MD performance is explained. Afterward, methods for improving the currently accessible membrane to improve performance will be examined. The cost

of membrane distillation will be focused on next, and the session will finish with a view on MD’s future [67]. Some examples of the different applications for membrane distillation systems in wastewater treatment and recovery are numbered in Table 2.

6

Membrane Bioreactor (MBR)

Membrane bioreactor (MBR) research to advance wastewater treatment is a prominent topic of membrane technology research. The MBR is a hybrid membrane technique that combines membrane separation with aerobic or

Novel Membrane Technologies in the Treatment and Recovery of Wastewaters

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Fig. 10 Configuration of MBR systems: a submerged MBR, b side-stream MBR configuration [71]

anaerobic bioreactors. In either case, membrane filtration is employed to retain biomass completely within the reactor to remove or change pollutants in wastewater via biodegradation while providing high effluent (permeate) quality [68]. In general, retained contaminants, usually solids containing bacteria, are separated using low-pressure-driven membranes like MF or UF [69]. To reuse wastewater, other membrane types, such as NF or RO, are also used downstream of the MBR process [70]. External circulation (side-stream design) or submerged membranes are used in the two primary MBR configurations (Fig. 10). Numerous research has been conducted since the mid-1990s to broaden the use of technology of MBR by producing permeate, which has high quality on a broad scope during lowering capital expenditures (CapEx) with the introduction of immersed configuration. The form of the membrane module, the distribution of the membrane’s pore sizes, and operating conditions to reduce

membrane fouling phenomena while developing novel methods to clean fouled membranes may be summed up as the examined parameters. The global economic crisis after 2008 slowed the commercialization of MBR plants. However, MBR technology has a promising future since cleaner water environments are much needed [72]. The typical process for conventional activated sludge (CAS) consists of two phases. First, one uses a blower tank where active microorganisms are used to clean wastewater (i.e., activated sludge). The tank of sedimentation, also known as the secondary clarifier, separates the treated water and the active sludge [72]. The lighter fraction of activated sludge is often processed with the treated effluent since it cannot be entirely separated in the sedimentation tank. However, most of the activated sludge may be separated while utilizing MBR due to membrane existence of various pore sizes. Table 3 summarizes the pros and cons of MBR against CAS [72].

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Table 3 MBR’s pros and cons compared to CAS [73] Advantages

Disadvantages

• Lower footprint due to reduced bioreactor size and removal of sedimentation tank

• Disregarding the membrane’s operational complication and process installation, fouling phenomena is a typical issue with MBRs that call for various operational techniques to lessen the fouling propensity of the membrane

• There are no restrictions on the amount of mixed liquor suspended solid (MLSS) in the MBR, which lowers the production of waste activated sludge (WAS) • The highest volume of MLSS in CAS is approximately 5000 mg/L because of restrictions with clarifiers which is second. However, the ideal range of MLSS concentration for MBR is between 8000 and 12,000 mg/L

• Due to the price of the membrane and antifouling techniques, the capital expenditures (CapEx) and operating expenses (OpEx) increase the cost of MBR system

• The solid retention time (SRT) may be used to assess the quality of treated water and the bioreactor’s MLSS • Due to the removal of the secondary sedimentation tank, fine SRT control may be performed in the MBR

• The procedure complexity is mostly brought on by techniques for cleaning and maintaining the membrane

• Longer SRT often increases wastewater efficiency • Higher SRT application in MBR (longer duration than approximately 3 weeks) compared to CAS (typically 1 to 2 weeks) results better effluent character while treatment

• The easy tendency of foaminess is yet another issue partially brought on by MBR's higher aeration need

• As a result of the availability of a narrow gap membrane compared to suspended particles, highquality treated effluent is produced. However, the average SS concentration for effective secondary clarifiers is about 5 mg/L

• Higher power usage while in use

7

Technology Evaluation

As shown in most of the pressurized applications mentioned above, MF, UF, and NF are typically used as pretreatment stages before RO. This is done to decrease fouling of the RO membrane and to improve consistent flux maintenance. This also functions as a multi-barrier treatment for pollutant removal from wastewater [74, 75]. Water recovery from wastewater has clearly become a viable alternative thanks to pressuredriven membrane technologies. However, the energy needs owing to the pressure remain an issue. With all of its intriguing characteristics, FO does have certain shortcomings that must be addressed. Aside from specialized uses of draw solution, where the draw solute is included into

the end product, further separation is required to recover freshwater. Another disadvantage of FO is low permeate flow owing to concentration polarization (CP). This CP reduces permeate flow by influencing net osmotic pressure. Again, the energy needs for FO rise as molecule weight decreases (MWCO). This is since regeneration of draw solutes would necessitate membranes with smaller holes and higher pressure, such as RO. As a result, the overall energy requirements grow [1]. In wastewater treatment, ED is particularly helpful for removing total dissolved solids (TDS) and other ionized component particles. ED has extremely high-water recovery rates and needs very little preparation for feed water. There is also reduced membrane fouling because of the process reversal and the technology being paired with renewable energy supplies. However, ED is not ideal for high-salinity wastewater streams

Novel Membrane Technologies in the Treatment and Recovery of Wastewaters

since energy requirement in desalination is proportional with removed ions [1]. Pervaporation membranes are uniquely developed to have a stronger affinity for the component to be separated due to their specialized purpose. This means that the membrane's chemical composition and structure are important in obtaining the desired separation. Other parameters that influence pervaporation processes include feed concentration, partial pressure, temperature, and feed flow rate. Pervaporation is recognized to be an energysaving and environmentally beneficial technique, in addition to its capacity to separate liquid mixtures when traditional separation methods are restricted. However, there are several downsides to this technique. Due to the very delicate working conditions, large industrial applications have yet to be realized [1]. Submerged MBR technology quickly became popular owing to simplicity and minimal setup and gained popularity owing to its lower energy consumption as well as simplicity as compared to the side-stream MBR. Moreover, it has own method of cleaning the membrane units while they are immersed in the bioreactor [76, 77]. Also, MD has several disadvantages. For starters, a lack of consumption data leads to unpredictability in water production costs (WPC). Second, the lack of membranes particularly built for MD puts the process at danger when membranes intended for other processes are used. This can result in membrane wetting, which allows organic deposits to accumulate and necessitates rigorous pretreatment. This raises the cost of the process. Finally, because heat and mass transmission occur concurrently, a fluid boundary layer forms, resulting in temperature polarization (TP) [1].

8

Conclusion

Due to strict water quality laws and regulations, growing non-traditional pollutants, physical limitations, and the need for enhanced pathogen and contaminant removal, the effectiveness of conventional water treatment techniques has become

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constrained. Therefore, cutting-edge technology, such as membrane treatment technology combined with other hybrid technologies, will be needed. Since the 1990s, membrane technology has quickly advanced thanks to its operating simplicity, low chemical input, high energy efficiency, and space-saving capabilities. Due to the creation of high-quality membranes and declining prices, their usage in water treatment is expanding quickly. Nevertheless, issues including fouling, life expectancy, and selectivity-permeability trade-off are preventing their widespread adoption. Thus, further improving the performance of membrane technology can assist the globe in dealing with contemporary issues, for instance, drinkable water shortage, contamination of water, and water demand.

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Application of Supported Fenton and Fenton-Like Catalysts in the Degradation of Pharmaceuticals in Wastewater— A Review of New Technologies in the Last Decade Sanaa Rashid, Dominic Bale, and Katherine Huddersman to their improved efficiencies compared to traditional Fenton’s. This review will analyse some of the new technologies developed in the last decade, the problems they were designed to solve, their applicability to the treatment of pharmaceutical wastewater as well considering their effectiveness and relative cost. This review will also consider potential future applications as well as any outstanding issues in pharmaceutical wastewater treatment, to which more research needs to be conducted.

Abstract

The last decade has seen vast technological advances in the pharmaceutical industry with many new drugs being developed and brought to the market. Although this advancement within the pharmaceutical industry brings many positives, the most difficult issue facing pharmaceutical companies is the amount of harmful waste generated during their manufacture and subsequent accumulation of these pollutants in water systems. In many cases, the municipal wastewater authorities are unable to remove all pharmaceuticals which result in their discharge into the environment. Pharmaceutic companies that produce drugs and other products may try to deal with waste by incineration. However, in less developed countries, this waste stream may well be discharged to the environment. To combat this problem, significant research has been invested into the development of new technologies with the purpose of degrading these pollutants to less harmful levels. Advanced oxidation processes involving Fenton and Fenton-like catalysts supported on secondary matrices have been especially researched due

S. Rashid
D. Bale
K. Huddersman (&) Faculty of Health and Life Sciences, School of Pharmacy, De Montfort University, Leicester, UK e-mail: [email protected]

Keywords








Advanced oxidation Catalysis Fenton Fenton-like catalysts Pharmaceuticals

1




Introduction

Water is arguably the most highly valued and indispensable commodity on earth. It is crucial for the survival of all lives on earth and its use in industry during the development and manufacture of almost every material product. In developed countries, water is generally considered a necessity of daily life; however, this is not the case in developing countries which struggle to provide clean drinking water for their residents. Generally, water on earth is not in short supply

© The Author(s), under exclusive license to Springer Nature Switzerland AG 2023 E. Debik et al. (eds.), Wastewater Management and Technologies, Water and Wastewater Management, https://doi.org/10.1007/978-3-031-36298-9_8

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with approximately 75% of earth surface being covered in water; however, the issue is that almost all this supply is not suitable to produce drinking water. To some extent, access to clean drinking water for developed countries is also a problem, as a vast number of pharmaceutical compounds are released into the environment each day directly from wastewater treatment plants (WWTPs) due to the lack of adequate treatment technologies. One of the most predominant challenges today in treating wastewater is the degradation of persistent pharmaceutical compounds which resist destruction by conventional treatment methods, e.g. biological treatment [1]. Accumulation of pharmaceuticals in water systems and bodies (municipal water, surface water, groundwater, etc.) presents a very real threat to the environment and human health, and as such, the treatment of pharmaceutical wastewater is a topic of great interest, with research continually being carried out to find better solutions to current wastewater treatment methods. Due to the complex composition of most pharmaceutical compounds, conventional wastewater treatment plants cannot effectively degrade these pollutants [2], hence why pharmaceutical water pollution is so prevalent. Surface water and ground water contamination is largely due to the discharge of wastewater effluent without being sufficiently treated. The presence of pharmaceutical compounds in water systems is not only due to direct contamination from pharmaceutical wastewater effluent, but also due to the use of pharmaceuticals in livestock farming where contaminated wastewater filters down through the soil to reach groundwater sources or are washed directly into nearby surface waters by rain downfall. As each pharmaceutical manufacturing plant produces different types of pharmaceutical products: active pharmaceutical ingredients (APIs), complete formulated medicines, etc., many of which produce a combination of these products, it is difficult to identify pharmaceutical wastewater as a single class of pollutant, and so trying to develop a single treatment method to degrade all pharmaceutical pollutants is not advisable. This means that different treatment technologies will

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be needed depending on the type of compounds present in the wastewater. It is estimated that around 50% of the pharmaceutical wastewaters produced around the world are discharged to the environment prior to receiving sufficient treatment [3], and these untreated wastewaters typically contain a multitude of complex pharmaceutical compounds [1], each with the potential to be highly damaging to the environment.

2

Sources of Pharmaceutical Pollution in the Environment

Sources of pharmaceutical pollution can be organised into two major classes: (1) point source pollution; (2) non-point source pollution (diffuse pollution) [4]. Point source pollution can be tracked from a single identifiable source, for example, waste pipes feeding directly from a pharmaceutical manufacturing plant or WWTP into a nearby water source. This differs from nonpoint source pollution of which the source of environmental contamination is not easily identifiable. An example is the contamination of groundwater and surface waters from nearby agricultural land. Most farms make use of pharmaceuticals in the form of veterinary medicines, which are administered to animals and eventually end up in the soil after excretion. Herbicides can also be included here as some have shown to have pharmaceutical therapeutic effects [5]. Pharmaceutical manufacturing plants, whether they are involved in medicine dosage formulation or active pharmaceutical ingredient (API) synthesis, produce large amounts of wastewater daily. The composition of wastewater can vary significantly depending on the processes carried out, meaning that wastewater produced during a medicine formulation manufacturing process will differ in chemical composition and concentrations compared to wastewater produced in a plant’s research and development (R&D) department. For instance, wastewater from a manufacturing process will contain fewer pollutants, usually all of which are known, whereas wastewater form an R&D process is likely to contain more pollutants,

Application of Supported Fenton and Fenton-Like Catalysts …

some of which are not likely to be known [6]. This is important to note from an industrial point of view, as the wastewater produced from an API R&D department may likely require more treatment processes compared to a final dosage formulation manufacturing department.

3

Major Sources of Domestic Pollution

Healthcare institutions such as hospitals and senior residences have been identified as a major source of pharmaceutical contamination to the environment [7, 8] along with pharmaceutical manufacturing facilities. Wastewater effluent produced by healthcare institutions has been found to contain pharmaceuticals in concentrations greater than 0.01 µg/L, which is above the European Medicines Agency (EMA) threshold for environmental risk evaluation [7]. This is a cause for concern as no distinction is typically made between the wastewater treatment of health care effluents, specifically hospital effluents, and urban wastewater effluents [9]. Effluents from either source are sent for treatment at their nearest WWTP, which is not sufficient to remove the higher concentrations of complex pharmaceutical compounds generated by healthcare institutions, primarily hospitals [8].

are typically employed at WWTP’s are not powerful enough to degrade most antibiotic compounds, meaning that any effluent and/or sludge produced by WWTP’s will contain antibiotics to some extent [10]; therefore, more powerful processes are required to reduce the release of antibiotic compounds. Biological treatment processes in WWTPs have been reported to be almost negligible in the degradation of antibiotics, with some papers reporting a degradation rate of 0%. Titouhi and Belgaied [11] reported a heterogeneous Fenton process for the degradation of the antibiotic ofloxacin in which a degradation rate of 94% was achieved. This emphasises the power of a Fenton/Fenton-like Advanced Oxidation Processes (AOP) for the degradation of recalcitrant pollutants such as antibiotics. Action needs to be taken soon with regard to properly treating antibiotic-containing pharmaceutical wastewater because we are currently in a vicious cycle; higher concentration of antibiotic in the environment leads to increased bacterial resistance, which then requires a higher consumption of antibiotics to treat infections, and so forth. Without implementing powerful processes such as heterogeneous Fenton/Fenton-like AOPs into the treatment of pharmaceutical wastewater, this cycle is likely to progress and worsen.

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Fenton and Fenton-Like Oxidation

Antibiotic Resistance

One of the most prevalent problems with insufficiently treated pharmaceutical wastewater is the presence of undegraded antibiotic compounds in water bodies and municipal water systems. A global issue associated with antibiotic compounds present in the environment is the occurrence and progression of antibiotic resistance within bacteria. This has a direct impact on human (and animal) health with the decrease in efficacy of antibiotic medicine against infection, leading to a higher mortality rate [10]. This highlights a great need for proper treatment of pharmaceutical wastewater to reduce an emerging problem such as antibiotic resistance. Conventional processes that

Throughout this review, both Fenton and Fentonlike oxidation processes for the treatment of wastewater will be discussed. Fenton-like oxidation does not have an absolute definition and as such has been defined slightly differently in different papers. For the purposes of this review, Fenton oxidation is defined as hydrogen peroxide coupled with ferrous iron (Fe2+/H2O2) and Fenton-like oxidation as hydrogen peroxide coupled with ferric iron (Fe3+/H2O2). Fenton and Fenton-like oxidation processes promise to be extremely effective tools in the destruction and removal of pollutants found in wastewater, namely due to the production of powerful oxidising species, hydroxyl radicals (

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OH), as well as other reactive oxygen species (ROS). Hydroxyl radicals are powerful enough to, in the correct conditions, degrade organic pollutants to carbon dioxide and water (complete mineralization). Recently, more research on the incorporation of AOPs into wastewater treatment has been carried out due to the inadequacies of conventional biological treatments. Thus, as hydroxyl radicals are some of the most powerful oxidising species, Fenton and Fenton-like processes are some of the best AOPs available for degrading organic pollutants, particularly pharmaceutical compounds [12], hence the interest for the advancement of wastewater treatment research. Carbamazepine, an anticonvulsant, is known to be a particularly recalcitrant pharmaceutical with conventional biological wastewater treatment processes having little to no effect on the degradation of this compound. Ternes [13] reported data from a sewage treatment plant (STP) in Frankfurt, Germany, that removed the drug carbamazepine by just 7% using conventional wastewater treatment processes, including physical/mechanical methods such as sorption and flocculation, and activated sludge (biodegradation). This is an exceptionally low removal rate and demonstrates the need for more powerful techniques if we are to make a serious change to the treatment of recalcitrant pharmaceuticals in wastewater. The data reported by [13] correlate with similar data published by [14] wherein carbamazepine was removed by no more than 10% in a conventional WWTP. Keen et al. [15] also reported the resistance of carbamazepine to biodegradation methods, which suggests that any removal of the drug from wastewater is mainly, or solely, due to physical/mechanical methods such as flocculation. Sun et al. [16] investigated the use of a heterogeneous Fenton-like oxidation catalyst for the degradation of a mixture of carbamazepine and ibuprofen. The catalyst was comprised nanoparticles of nano-magnetite (Fe3O4). The magnetite nanoparticles were added to aqueous solutions of the sample pollutants, and the pH of the solutions was adjusted within the range of 5.32–8.68. The results from this experiment

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achieved degradation efficiencies of carbamazepine and ibuprofen of 90.0% and 80.1%, respectively. This demonstrates the benefit of Fenton-like oxidation in the degradation of certain recalcitrant pharmaceuticals’ pollutants when compared to conventional biological processes. Another study conducted by [17] evaluated the photocatalytic degradation of carbamazepine and ibuprofen utilising a MIL-53(Fe) metal–organic framework (MOF) as the supported Fenton catalyst. In this study, it was found that complete degradation of both carbamazepine and ibuprofen can be achieved, when the pH is maintained in the range of 6.0–9.0. When the pH was increased from 9.0 to 10.0, the degradation efficiency of the carbamazepine and ibuprofen decreased from 100% down to 40%. In [18] a combination of Fenton and Fenton-like reactions for the degradation of diclofenac, a non-steroidal anti-inflammatory drug (NSAID) was used. The group synthesised Fe3O4 nanoparticles onto an aluminium oxide (a-Al2O3) ceramic substrate to create a heterogeneous Fenton/Fenton-like catalyst in the form of a microfiltration membrane. Degradation reactions were carried out at two different pH values, 3.0 and 6.6. A degradation efficiency of diclofenac to be 69.0% at pH 3.0 after 24 h, compared to a removal efficiency of 30.0% at pH 6.6, was reported [18]. This can be attributed to the increased generation of hydroxyl radicals at more acidic pH, resulting in enhanced degradation of diclofenac using a Fenton/Fentonlike heterogeneous catalytic system. Further to this explanation, the increase in pH was reported to generate superoxide radicals (
O2−) which have a lower oxidation capacity compared to hydroxyl radicals. Removal rates of diclofenac from WWTPs using conventional biological treatments such as activated sludge and membrane bioreactors are noted to be no greater than 20% [19] and in some cases have been reported to be completely recalcitrant [20], making diclofenac an emerging environmental pollutant. In [21] the use of a silica-supported Fenton-like catalyst in the degradation of pharmaceuticals and drugs of abuse (DA) found in surface waters was

Application of Supported Fenton and Fenton-Like Catalysts …

investigated. Iron oxide, in the form of hematite (Fe2O3), was supported within the mesoporous structure of the silicate support, SBA-15, used in conjunction with hydrogen peroxide. The target drug compounds used in this study were reported as very recalcitrant against conventional biological activated sludge treatment. Liquid Chromatography—Mass Spectrometry (LC/MS) was used as the analytical technique to quantify the degradation results, and after treatment with Fenton-like oxidation, almost all the drug compounds were in concentrations below the limit of detection for the LC/MS analysis. This emphasises the value of advanced oxidation processes such as Fenton and Fenton-like oxidation in the degradation of pharmaceuticals.

6

Assisted Fenton/Fenton-Like Processes

The oxidation capacity of conventional Fenton/ Fenton-like processes can be enhanced by the inclusion of additional energy into the reaction, most typically in the form of UV light, solar light, or electricity, known as photo-Fenton, solar-Fenton, and electro-Fenton processes, respectively. Both photo-Fenton and solar-Fenton are photocatalytic processes which utilise light radiation to enhance the production of hydroxyl radicals, resulting in a more concentrated reaction medium. The major difference between these two processes is that photo-Fenton utilises lamps to generate artificial UV light, whereas solarFenton captures light radiation directly from the sun, inclusive of UV and visible light. There are few ways in which the addition of light radiation can enhance the Fenton process: (1) the electromagnetic energy of the light directly decomposes hydrogen peroxide into hydroxy anions and hydroxyl radicals; (2) the target substrate absorbs energy from the light, making the molecule more susceptible to the oxidation potential of the hydroxyl radicals. The use of a photo-Fenton system in the degradation of ofloxacin from wastewater was studied by Du et al. [22], with the catalyst being

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an iron–magnesium oxide (FeMnOx), and was found to be highly effective with removal rates over 98%. Compared to the standalone Fenton system used by Du et al. [22] (absence of UV light), which achieved only a 64% degradation rate of ofloxacin, the increased efficacy of the photo-Fenton system can likely be attributed to the inclusion of UV light. In [23] the research group also utilised a photo-Fenton system in the degradation of drug compounds. As expected, the degradation efficacy of the photo-Fenton system superseded that of a non-assisted Fenton system, illustrating that the inclusion of UV light enhances the oxidation potential of the Fenton reaction. The parameter which had the biggest effect on the degradation efficacy was the catalyst iron loading concentration, with the hydrogen peroxide concentration having seemingly no effect at concentrations above 25%. From an industrial perspective, this is an interesting result due to the cost reduction associated with using lower concentrations of hydrogen peroxide compared to a traditional Fenton system. This coincides with what [11] reported, stating that their Fenton reaction for the degradation of ofloxacin was highly influenced by the variation of hydrogen peroxide concentration. Electro-Fenton reaction is an electro-catalytic method that incorporates electrical current into the reaction process to enhance the production of hydroxyl radicals using two cathodes where one is iron doped. There are four typical variations in the way this system works, but the idea that ROS production is increased remains constant between these variants. The principal difference between electro-Fenton and other Fenton processes is the in-situ generation of hydrogen peroxide and subsequently ROS compared to more traditional Fenton processes, in which the reagents (hydrogen peroxide and Fe2+/Fe3+) are added externally to the reaction solution. In essence, compressed air is pumped into an electrochemical cell within the reaction vessel, resulting in the two-electron reduction of O2 molecules, which then bond with free H+ to form hydrogen peroxide. This is shown in Eqs. (1) and (2).

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O2 þ 2e ! 2O ;

ð1Þ

2O þ 2H þ ! H2 O2 :

ð2Þ

Fluidized and Fixed-Bed Reactors

Fluidized and fixed-bed reactors are a relatively new assisted Fenton/Fenton-like AOPs, in which catalyst particulates or inert solids (both referred to as carriers) are held inside columns within the reactor, providing either an activation site (catalyst material) for reaction catalysis or a site for crystallisation and/or precipitation (inert solids) of leached iron. Inert solids such as SiO2 can be incorporated into the reactor to precipitate out any dissolved iron in solution, either from a homogeneous Fenton setup or any iron that has leached into solution from a heterogeneous iron catalyst [24]. This has the potential to reduce the need for more complicated post-treatment purification stages with the aim to reduce any post-treatment secondary pollution. Moreover, iron catalyst solids can also be used providing an active site for chemical catalysis to take place. Both fluidized-bed and fixed-bed reactors perform the same job, and there are some significant differences which give each reactor type advantages depending on the intended job. Carriers used in fluidized-bed reactors, whether catalysts or inert solids, are not fixed but are suspended within the reactor by the upward flow of solution, whereas, in fixed-bed reactors, the carrier particulates are fixed in place, hence the name. The two main advantages to fluidized-bed reactors are the higher efficiency in heat transfer and the ability to perform inline catalyst exchange [25]. The inline catalyst exchange ability is of most interest when referring to the treatment of pharmaceutical wastewater on an industrial scale as this would slow down time needed to replace spent catalysts, subsequently reducing costs. In [26] the research group utilised both fluidized and fixed-bed reactors for the degradation of six recalcitrant pharmaceuticals using a diatomite supported heterogeneous iron catalyst. The results from this research show that on

average (across the six compounds treated) the fluidized-bed configuration was slightly more effective in terms of degradation compared to the fixed-bed reactor setup with an increase of 0.8%. However, deactivation of the catalyst after 10 h of operation was far greater in the fluidized-bed reactor at 52.1%, compared to 27.9% for the fixed-bed reactor. This indicates that the fixedbed reactor setup is more appropriate for commercial application, as it will require fewer catalyst changes, and this is also emphasised by Ulloa-Ovares et al. [26] reporting that this configuration was also more energy efficient. An explanation for the deactivation of the catalysts is iron leaching; however, this was assessed and the iron concentration in the reaction solution after 10 h was found to be no more than 8 mg/L for both reactor types. This is a relatively low concentration of iron with regard to leaching and is a factor to be considered when scaling up to industrial scale.

8

Effect of pH

It is traditionally documented that the Fenton oxidation process, both heterogeneous and homogeneous, has optimal generation of the ROS hydroxyl and hydroperoxyl radicals in the pH range of 3–5 [27]. Reaction kinetics are typically noted to decrease as the solution becomes less acidic. This is likely due to the more rapid decomposition of hydrogen peroxide into oxygen and hydrogen at higher pH, especially when the solution pH becomes alkaline (pH > 7). This reduces the concentration of hydrogen peroxide available for radical generation. At more alkaline pH, there is also the preference for superoxide radical (∙O2−) generation, which is a less powerful oxidation species compared to hydroxyl and hydroperoxyl radicals [28]. Furthermore, an alkaline solution will cause iron species to precipitate out of solution usually as some form of iron hydroxide, which reduces the capacity of the catalyst to function as required. This is more applicable to homogeneous Fenton reaction rather than a heterogeneous system.

Application of Supported Fenton and Fenton-Like Catalysts …

Wastewater effluent usually has a circumneutral pH to minimise any chemical changes and to protect the ecology of the receiving water bodies. This is an important factor to consider when developing and implementing Fenton/ Fenton-like processes within WWTPs; as mentioned previously, these generally work best at acidic pH. However, there has been recent research into the use of assisted heterogeneous Fenton processes in the removal of pharmaceutical compounds form wastewater at near neutral pH. This is an important development as it would mean that fewer (if any) pH adjustment stages would need to be implemented to reduce the environmental impact of the wastewater effluent.

9

Supported Fenton Catalysts

Supported Fenton catalysts have been explored in many studies over the last decade. Particularly from an industrial point of view, the use of a catalyst that can be easily removed and recycled is of interest compared to homogenous catalysts. Supported heterogenous catalysts also have the added advantage of having the potential to be used at higher pH values as well as reducing the amount of secondary pollution via iron leaching. Choosing the correct type of support is pivotal to optimising the degradation of pollutants in wastewater. This is due to the mechanism of Fenton catalysis, where the first step is the adsorption of pollutants onto the catalyst surface. A good support will facilitate the movement of the substrate to its surface, allowing the reaction to take place and ultimately allowing the substrate to be desorbed. Thus, this review will investigate in depth the supported Fenton catalysts utilised in the last decade to treat pharmaceutical wastewater. In particular, the applications must have potential for use in an industrial scenario, which must be demonstrated by either extensive lifetime studies, cost analysis, or real-world applications. In Table 1, the properties of supported catalysts are listed along with their lifetime and cost comparisons.

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Methods of Incorporating Fenton Catalysts onto Supports

There are several methods in literature used to incorporate Fenton catalysts onto supports. The first is the sol–gel method, which can form composites by (1) hydrolysis of the precursors in acidic or basic mediums and (2) polycondensation of the hydrolysed products to form the solid oxide composite. This method has allowed the production of iron and titania composites [29, 30], copper/iron composites [31–33] among others. Solvothermal methods are also well researched to produce Fenton catalysts such as magnetite/graphite composites [34] and iron biochar composites [35]. In particular, the use of the method to produce nanoparticles [36] is well documented. These nanoparticles can be easily modified by changing solvents, temperatures, reaction times, and chemicals used. Another similar method is hydrothermal synthesis, which differs only in its use of solvent that is usually water. However, in many studies, the terms solvothermal and hydrothermal are used interchangeably [37]. Dip coating of glass with goethite [38], carbon nanotubes with titania [39], and the deposition of manganese oxide nanoparticles on graphite are some of the many uses of this functionalization method. One of the advantages of dip coating a Fenton support using an iron oxide is the 100% purity of the deposited iron oxide, which cannot be achieved easily with other methods [38]. The method can also be modified depending on the requirements for thicker or thinner layers whilst maintaining reproducibility and uniformity [40]. Ion exchanging, chelating, or complexing metals onto a support are also an efficient method of catalyst deposition. This is usually dependent on the functional sites available on the sorbent as well as pH and concentration. Well-known ion exchangers are zeolites, which will be discussed further in this review, as well as modified polymers.

Acetaminophene Petrochemical wastewater

Ciprofloxacin

Photo-Fenton reactor

Ozonation in fixed bed reactor

Photo-Fenton reactor

Cellulose acetate/MOF films

Titania-supported sand and iron-rich clay

Ag–Fe–TiO2

Sulfamethoxazole

Stir tank reactor

Surface-functionalised PAN mesh

PANI-clay-Fe composite

Naproxen

Ceramic membrane reactor with electro-Fenton reactor

MIL-88(Fe)

Atrazine

Estrone B-estradiol a-ethinylestradiol

Ornidazole Ofloxacin

Al2O3-supported copper oxide

Hospital wastewater

Electro-Fenton

Photo-Fenton reactor

Mixed pharmaceuticals

Nitrogen-doped biochar

Fluidised bed reactor Fixed-bed reactor

Magnetite-doped diatomite

Target pollutants

Iron-alginate spheres

Reactor types and other assistance

Catalyst

Table 1 Catalysts with lifetime and cost comparisons

100% in 120 min

100% dropping to 85% after 6 cycles (720 min)

2 months Regenerated with 1 g/L H2O2

86% removal at 450 min

* 90% in 90 min > 80%

13% reduction after 4 cycles

12.8% reduction over 30 cycles

96% reduction to 85% after 5 cycles

6 cycles: 91.3%

Lifetime

95% 92%

81.6% in 60 min

60.7–26.0%

96% removal

95.7%

40.1% dropping to 19.2% over 10 h 35.8% remaining stable over 10 h

Efficacy % 3

N/A

N/A

N/A

N/A

$2688 per tonne for fresh catalyst $3162 per tonne to replace deactivated $474 per tonne regeneration

N/A

81.4 Wh/m 1.01 Wh/m3

Cost

(continued)

[105]

[99]

[69]

[110]

[81]

[107]

[77]

[50]

[43]

[26]

Ref

114 S. Rashid et al.

[111]

[101]

Ref

[90]

Application of Supported Fenton and Fenton-Like Catalysts …

11

115

Natural Catalyst Supports

Natural sorbents and their use as catalyst supports are a growing research area due to the increasing focus on sustainability.

N/A 98.2–91.3% after 8 cycles 98.2% removal in 120 min Electro-Fenton reactor Activated carbon cathodesupported magnetite NPs

Phenazopyridine

6 cycles 84% TOC removal after 6 cycles 100% removal of ofloxacin (1 h) 88% TOC removal after 8 h Electro-Fenton reactor Iron(II)(III)hydroxide-doped carbon felt cathodes

Ofloxacin

15% loss of activity after three cycles 98.3–90.2% in 5 h at pH neutral

Cost Lifetime Efficacy % Target pollutants Reactor types and other assistance

Baffled reactor Fe–Cu-doped carbon nanotubes

Acetaminophene

Biochar is an interesting carbon rich material, which is currently being explored for CO2 trapping in soils [41]. It is produced via pyrolysis of biomass, which can be used in soil remediation [42] and in water treatment due to its absorbent and porous nature. Due to its good conductivity, many researchers have attempted to use biochar as an electrode in electro-Fenton process [43].

11.2 Chitins

Catalyst

Table 1 (continued)

N/A

11.1 Biochar

Animal-based materials such as chitins are also interesting due to their easy availability and strength. As chitins are composed of acetylglucosamine units, and chitosan is formed of both dglucosamine and N-acetyl-d-glucosamine, this allows its functionalisation. Due to their strength, chemical stability, and antimicrobial activity, chitins are currently being explored as sustainable food packaging. The use of chitins, especially chitosan in pharmaceutical wastewater treatment, has been seen in many studies. Chitosan films, which are base hydrolysed and then impregnated with iron, have been used to remove triclosan from water [44]. The functionality of chitosan can also allow the material to be crosslinked before doping with iron [45]. The use of nitrogen doped chitosan electrodes in electro-Fenton degradation of acebutolol has also been studied due to its costeffectiveness compared to traditional sacrificial cast iron electro-Fenton cathodes [46] as well as its use as catalytic membranes to degrade pharmaceuticals in wastewater [47].

11.3 Alginate Alginate is a naturally occurring polymer that can be found in brown seaweeds. The two main

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copolymers that make up alginate and bestow both strength and plasticity to the polymer are guluronic acid and mannuronic acid [48]. As alginate is biodegradable, it is a promising support as a Fenton catalyst. In many studies, the iron is encapsulated within alginate beads via crosslinking in a calcium chloride solution [49, 50]. The ease of forming the alginate beads along with its relative availability means that it can be combined in many studies with other materials. In the study by Kong et al. [51], the antibiotic tetracycline was removed from water via a graphene oxide/alginate iron composite.

11.4 Diatomite Diatomite, which is the silica rich fossilised remains of diatoms, has traditionally been used as building materials [52]. Due to its low density, high availability, and porosity, using diatomitebased materials as sorbents in wastewater treatment has been explored. The photodegradation of ciprofloxacin using a LaFeO3-doped diatomite disc was achieved by [53], whilst a MnFe2O4 diatomite composite was used to degrade tetracycline in [54]. The porous nature of diatomite means that it can be easily pelletised and used in both fixed-bed and fluidised reactors [26].

11.5 Zeolites Zeolites are aluminosilicate materials that can be easily modified to act as metal ion exchangers [55]. The materials are also highly porous, which can aid the sorption of substrates to its surface to facilitate pollutant degradation. Once adsorbed on the zeolite surface, the zeolite structure holds the substrates via electrostatic attraction [56]. There are many types of zeolites available, and as they all contain varied SiO2/Al2O3 ratios and particle sizes, they can influence the catalytic efficiency of the held iron. Their ion exchanging abilities also mean that they can be easily

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modified to incorporate new active sites, such as the doping of iron [57] or iron oxide nanoparticles [58] for use in Fenton degradation. In the study by Anis and Haydar [59], an analcime zeolite (NaAlSi2O6
H2O) was utilised to degrade caffeine in water alongside an iron sulphate and hydrogen peroxide solution. The zeolite was able to remove 80% of the stimulant at a relatively high pH range of 5.8–7.0 [59]. The authors suggested the cost of treatment using the analcime zeolite would be $23.20 per m3 of waste compared to the cost of homogenous Fenton, which is $29.4 per m3 [59]. Fe-ZSM5 catalysts, whilst not occuring naturally, are commercially available and have been used in a number of studies to degrade pharmaceutics in wastewater [60, 61] as well as being incorporated into membranes [62]. The appeal of this zeolite is its commercial nature as well as its low leaching content of 0.2% over three hours of reaction [61]. Zeolite Y, occurs naturally as the mineral Faujasite, a hydrophobic zeolite with a high silica content, has been used in many studies to remove antibiotics from water. The study by Braschi et al. [63] used zeolite Y in the degradation of sulfa-antibiotics. An interesting study of real water treatment by Ayoub et al. [64] using an iron(III)-impregnated zeolite Y managed good removal of diclofenac, estrone, lidocaine, triclosan, ketoprofen, ibuprofen, etc., after 6 h treatment with UV assistance [64]. Unfortunately, this trend was not possible for sulfamethoxazole [64]. Recently, sulfamethoxazole was reportedly removed 100% at pH 7 using an Fe(II) NaY zeolite alongside peracetic acid [65]. Although zeolites are well researched in Fenton reaction, there are some disadvantages to their use in water remediation. For example, pore size is extremely important and potentially limiting, if an organic pollutant is larger than the pore size of the zeolite [66] and thus blocks its diffusion and consequent breakdown. The pore size may also help to speed up deactivation via site blockage or loss of active sites via leaching.

Application of Supported Fenton and Fenton-Like Catalysts …

12

Metal Organic Frameworks (MOFs)

Metal Organic Frameworks (MOFs) are structures of a crystalline nature that are made up of both metal ions and organic components. The porous nature of the MOFs as well as their ability to be modified in terms of size and functionality makes them a desirable Fenton support. Similarly to zeolites, the expansive surface area and porosity of MOFs can facilitate the sorption of pollutants and their consequent destruction. MIL-type (Materials Institute Lavoisier) MOFs are interesting due to their ability to react to their surrounding environments; hence, MIL types such as MIL-53(Fe) are slowly being recognised as interesting Fenton catalysts [67, 68]. MIL-type MOFs, such as MIL-88B(Fe), have also been supported on ceramic membranes such as in the study by Ye et al. [69], which degraded naproxen in wastewater with electroFenton assistance. The ability to tune MOFs by the addition of different metal ions can also be beneficial to the Fenton process. For example, MOFs that contain both iron and copper ions have been used to degrade sulfamethazine [70], whilst iron and nickel MOFs in photo-Fenton reaction have been used in tetracycline degradation [71]. Studies have also investigated the introduction of expensive palladium in MOFs [72, 73]. A disadvantage of MOFs is the difficulty in recyclability. To counter this, some researchers have further supported MOFs onto other materials such as glass. In the study by Samy et al. [74], MOF-808 carbon nanotube composites supported on glass plates were mounted into a photocatalytic reactor to degrade carbamazepine and diazinon. The glass-mounted MOF composites were able to be recycled 5 times before deactivating due to physical loss of active sites and the possible poisoning of the catalyst [74]. MOFs have also been supported on carbon nitride polymer, which adds stability as well as improves photocatalytic activity by reducing the recombination of electron holes [75]. Other supports for MOFs include graphite felt for the electro-Fenton degradation of ciprofloxacin [76]. An alternative to coating the MOFs on a support

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is embedding them within a material such as a cellulose film for the photodegradation of paracetamol [77] or mesoporous carbon for the Fenton degradation of amoxicillin [78].

13

Titania and Alumina

Titania photocatalysts have been utilised in many studies over the last decade due to their synergistic effects when coupled with iron oxides during Fenton oxidation. Although their use as a Fenton photocatalyst is well investigated, titania remains an expensive choice for a support. When in the presence of UV light, electron–hole pairs are formed on the titania surface, which allow redox reactions. Thus, some researchers attempt to offset this cost in other areas. For example, in the study by Talwar et al. [79], a titania support was doped with sand and iron-rich clay (Fuller’s Earth). The use of low-cost bentonite clay has also been explored to remove acetaminophen in both a Fe/titania composite [80], magnetite/ titania [38], and Ag–Fe composites [81]. Although the use of titania with UV assistance is well documented and known to be quite efficient, the use of synthetic UV light can be an added operating cost and is unsustainable. Thus, many researchers are utilising solar irradiation as an alternative energy source. The incorporation of titania and alumina in ceramic membranes has also been explored. Alumina as a support material has been used to remove diclofenac in a ceramic membrane reactor [18]. Although there are many studies, which utilise immobilised iron on alumina for Fenton degradation, these are mostly in the degradation of phenols [82, 83] and dyes [84, 85]. Thus, more research into the use of alumina/iron composites for pharmaceutical degradation should be explored.

14

Carbon-Based Supports

As shown in the study by Sopaj et al. [86], the type of carbon cathode material can have an impact on the degradation capabilities of the electro-Fenton process. In particular, the type of

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material used can have an impact on the amount of hydrogen peroxide produced in the system, which can ultimately contribute to the OH radical production [86, 87]. Carbon materials in other formats such as felts have also been used as porous cathodes for the electro-Fenton degradation of acetaminophen [88].

14.1 Carbon Nanofibers Carbon nanofibers are cylinders made of carbon and graphite layers. Their composition allow them to be strong and having good stability and electrical conductivity. Thus, they make promising materials for cathodes in electroFenton use. The carbon nanotubes can easily be modified to introduce metal active sites, which have included magnetite [89], iron and copper [90]. The most popular metal active sites are the use of nanoparticles, which can either be doped on the surface of the nanotubes or alternatively within the walls [34, 91]. As many of these nanoparticles have magnetic properties, they are easily removed from solution and recycled [91].

14.2 Graphite and Graphene Graphite and graphene supports have been used in aerospace, architecture, and energy materials due to their strength and thermal and electrical conductivity. Graphene-based cathodes have been used in a variety of studies for the removal of pharmaceuticals from water [92]. Graphene oxide, which was functionalised with ferrocene, was used as an electrode in the electro-Fenton degradation of both ciprofloxacin and carbamazepine [93]. As graphite cathodes can be very expensive to utilise and to replace, a sustainable source is of great interest. A study by Kadji et al. [94] reused the graphite rods from used batteries for electroFenton degradation of amoxicillin. However, the reuse of the battery rods did not significantly reduce the energy costs of the electro-Fenton process [94]. Graphite composites with magnetite, which are formed by solvothermal methods, have also

S. Rashid et al.

been explored for their use in degrading levofloxacin [95]. Graphite and magnetite composites have also been used as cathodes for use in electro-Fenton degradation of tetracycline [96].

14.3 Polymer-Based Supports Polymeric catalyst supports are increasingly popular due to their easily functionalisation and chemical stability. With many polymers available, supports must be functional for metal deposition, cost effective, and having a long lifetime. Polymeric membranes have been utilised to remove amoxicillin from wastewater using a goethitedoped polyacrylonitrile (PAN) membrane [97]. It is thought that the incorporation of Fenton catalysts helps to counter the effects of fouling. Functionalised PAN fibres have also been doped with iron to degrade pharmaceutical wastewater. The study by Hosny et al. [98] used a novel iron-complexed surface-functionalised PAN fibrous mesh to remove the alkaloid colchicine from wastewater, whilst [99] successfully degraded 90% of endocrine disrupting compounds in wastewater. Electro-Fenton process is an efficient method of supporting Fenton degradation. However, it must be noted that the high current utilised during electro-Fenton reaction can destroy the graphite anode, which is costly to constantly replace [56]. Thus, the utilisation of polymer electrodes [100] could possibly be an interesting substitution. The use of a magnetite-doped PAN nanofiber fabric cathode has been utilised to degrade carbamazepine in wastewater [89]. Other iron oxide-doped carbon fibre cathodes include Fe(II)/ (III)hydroxides to remove ofloxacin [101]. Another polymer, which has been studied frequently for dye removal but less so for pharmaceutical degradation, is polyaniline (PANI). As PANI is a highly conductive polymer due to its benzene rich backbone, its use as an electroFenton support would have been thought to be widespread. However, due to its poor solubility in solvents, its application without another material is limited. Thus, many studies incorporate PANI along with the materials such as

Application of Supported Fenton and Fenton-Like Catalysts …

zeolites [102, 103], carbon nanotubes [104], and clay [105] to name a few.

15

Deactivation and Regeneration of Fenton Catalysts

All catalysts have a limited lifetime, and as explained in the study by Argyle et al. [106], deactivation can be caused by degradation, fouling, poisoning, wear of the catalyst, and leaching. The deactivation and consequent regeneration of Fenton catalysts is a subject explored very little. However, this is key for scaling up the technology into industry. A recent study by Jin et al. [107] found that an industrially used Al2O3-supported copper oxide catalyst deactivated over the course of its fiveyear usage with a degradation reduction from 60.7 to 26.0%. The mechanism of deactivation was determined to be due to the blockage of catalytic active sites by the electrostatic adhesion of colloidal compounds [107]. Ultimately, the catalysts were regenerated by calcination at 500 ° C [107]. In many studies, the main mechanism of catalyst deactivation is loss of the iron active sites. The study by [108] utilised an iron-rich clay, Hangjin2#, which suffered from leaching instigated deactivation with an iron concentration drop from 12 mg/L on the first cycle to 7.13 mg/L on the third. The extensive study by Akinremi et al. [109] into both the deactivation of a functionalised PAN catalyst and its consequent regeneration found that iron leaching was the primary mechanism of deactivation.

16

Conclusions and Future Considerations

Over the last decade, many new technologies have been investigated for the removal of pharmaceuticals in wastewater. In many cases, existing materials have been repurposed for use as supports. A particular interest and aim are to ensure that the materials used as supports are both

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sustainable and have as little impact on the environment as possible. The next decade is crucial for turning the tide on the impact of water pollution. Thus, the authors suggest the following areas of focus. Firstly, the lifetime of supported catalysts should be increased further to reduce the effects of secondary pollution and to improve their ability to be scaled up in industry. Many current studies do not discuss lifetime or only provide limited batch work, which is not representative of real-life applications. Furthermore, efficient methods of regeneration should be explored to better exploit the emerging technology. Another desirable ability is for the supported catalyst to be used in a wider pH range. Many studies are already trying and succeeding; however, extensive lifetime studies need to be carried out. Another interesting area of research is the insitu production of hydrogen peroxide. Lastly, future supported Fenton technologies need to be green with a low carbon footprint and sustainable.

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Combined Ferrite Treatment of Multi-component Wastewaters Under the Elevated Temperature Gheorghe Duca, Victor Covaliov, Olga Covaliova, and Lidia Romanciuc

Abstract

Ferrite treatment of wastewater is a promising approach, since it allows to efficiently treat the polluted water containing at the same time dissolved heavy metal complexes within a broad pH range, and organic compounds. Application of ferrite treatment makes it possible to resolve a series of environmental and economic challenges due to the high degree of water purification and final formation of poorly soluble sediments with crystal oxide structure and ferromagnetic properties. This paper contains a review of ferritization treatment methods of multi-component wastewaters. It was shown that the main factor ensuring the formation of ferrites in the studied electrocoagulation treatment process is temperature, which needs to be not less than 70 °C and up to 85 °C. To carry out electrolysis with converting iron (III) hydroxide into Fe3O4, an optimal power consumption of 3–5 Coulombs per 1 mg of iron (III) contained in the sediment

is required. The optimal conditions for this process are: ia (current density)—up to 2.5–3.0 A/dm2, temperature 70–80 °C, and pH values 4.0–8.0. Application of high-frequency currents for studied processes was tested, and the ferrite sludge with electric susceptibility of 2000–2200 Oersted was obtained, with higher index of hydraulic fineness of sludge particles which reached 0.86 mm/s. The value of the specific magnetic susceptibility of sediment in the dry state is 2.0–3.0 cm3/g. At the same time, the removal of heavy metal ions within 0.5 h treatment time reached 98–100%. It was shown that despite the high temperatures needed for this type of treatment, the resulted sludge can be readily separated from liquid phase both in gravitational and magnetic fields, which implies a series of advantages, including shorter time for sludge sedimentation and separation and lesser volumes of sedimentation tanks, prevention of environmental pollution with heavy metals, and a variety of possible applications of magnetic sludge. Keywords




G. Duca
O. Covaliova (&)
L. Romanciuc Institute of Chemistry, Chisinau, Republic of Moldova e-mail: [email protected]




Ferritization Electrocoagulation Wastewater treatment Sediments particles





Ferrite

V. Covaliov Institute of Research and Innovation, Moldova State University, Chisinau, Republic of Moldova © The Author(s), under exclusive license to Springer Nature Switzerland AG 2023 E. Debik et al. (eds.), Wastewater Management and Technologies, Water and Wastewater Management, https://doi.org/10.1007/978-3-031-36298-9_9

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Introduction

Ferrite treatment of wastewaters has become attractive for researchers and water experts as it offers a possibility to obtain easily separable and dewatered ferromagnetic sediments with high chemical stability. In this way, a higher level of wastewaters purification from heavy metal ions and organic pollutants than conventional treatment methods can be reached. Although the ferritization processes as such have been known for a long time already [1], the interest in the development of ferrite technology of wastewater treatment re-appeared due to the obtaining the so-called “magnetic fluids” [2–4]. Magnetic fluids were among the technical innovations developed under the NASA “Apollo Program” in 1960s and were used for vacuum sealing of moving parts of astronauts’ suits [5, 6]. Magnetic fluids are composed of a liquid and nanoscale particles with the diameter of about 10 nm or less, made of magnetite (Fe3O4), hematite (Fe2O3) or other iron-containing components. However, these small-size particles are not evenly dispersed in the carrier liquid [7]. Since then, due to the research performed in different countries, a series of interesting physical properties of magnetic fluids have been discovered; thus, it was found that they can give better speaker sound quality, reduce the overall harmonic and dynamic distortion; some of them have shown rather high heat resistance, low volatility, and specific optical properties [8]. Because of these and other unique properties, magnetic fluids have found various applications, especially in medicine and biotechnology [9, 10], in optical devices [11], for vacuum dust sealing of motors, stepping the motor damper, lubrication [12, 13], wastewater treatment [14–21], oil spills remediation [22, 23], and other technical processes. Production of magnetic fluids includes synthesis of colloidal ferromagnetic particles with different compositions (mainly Fe3O4) distributed within the liquids. This technology is based on chemical processes like those of the ferrite treatment of wastewaters that can be expressed by the general type of reaction:

Fe2+ + 2Fe3+ + 8OH− ! Fe3O4 + 4H2O. In the presence of a mixture of heavy metal ions, a series of electrochemical, reduction–oxidation, and catalytic processes are running. As a result, the ferrites with overall formula are formed: (Me2k+Ok2−)m/2 (Fe23+O32−)n, where Me—metal, k—its valence, m and n—whole numbers. Purification of water is provided due to the enhanced sorption-active surface of ferrite structures now of their formation. The main factors influencing the ferrite transformation of the hydroxides Fe(OH)2 and Fe(OH)3 into Fe3O4 are temperature, pH, concentration of ironcontaining component, and other technological process parameters [24]. Practically, all ferritization water treatment methods involve the introduction of ironcontaining compounds into the treated water. Chemical technology is based on the introduction of water-soluble Fe(II) compounds, whereas the electrocoagulation or Galvano-coagulation processes involve the primary anodic dissolution of iron: Fe0 + 2 e ! Fe2+ under the application of external electric current, or during the internal electrolysis due to the contact of galvanic couples. Bivalent iron ions Fe2+ will be subsequently partially oxidized to the trivalent ions Fe3+. Ferrite wastewater treatment makes it possible to remove the heavy metal ions due to the formation of metal oxides and further formation of sediments with more chemically stable spinel structure. One of the main advantages of ferrite technology applied for the wastewater treatment is that it allows to obtain the sediments with the improved ferromagnetic, sedimentation, and filtration properties [25]. The ferrite particles thus formed can be easily separated in the gravitation and magnetic fields and are not hazardous for the environment. Due to the other useful properties, ferritized sludge can be utilized and commercialized. Ferrite technology is based on several technological directions and can involve both physico-chemical catalytic processes and electrocoagulation, galvano-coagulation, and other technologies (Fig. 1). These technologies are under permanent development. In the Republic of Moldova since

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Fig. 1 Classification of ferrite technologies for wastewaters treatment and ferrite sludge utilization (numbers in square brackets give the related references)

the late 1970s, different hydrothermal ferritization processes of concentrated sludge treatment, resulted from the electrocoagulation purification of multi-component wastewaters, have been implemented, for example, at the tractor factory, washing machines factory and fridge factory, where sufficient amounts of industrial wastewaters were formed containing such metals like Ni, Cr, Fe, Cu, etc., as well as surfactants, which made them appropriate for the electrocoagulation and ferrite treatment [26]. The sludge formed after the sedimentation and draining of purified water has been heated to 65–70 °C and subjected to ferrite treatment; then, it was easily dewatered and utilized. Barrado et al. [27] have suggested the mechanism of ferritization reactions with participation of heavy metal ions, following the general type reaction: xMen þ þ 3FeSO4 þ 4NaOH þ 1=2O2 þ ! Mex Fe3x O4 þ 3Na2 SO4 þ 3H2 O þ x½Fentotal ;

although this scheme does not completely reveal the complex series of all the running reactions. Practical usefulness and benefits of ferrite water treatment technology are, however, connected with the increased heat consumption for heating of the treated water along with the consumption of iron salts for the sludge ferritization process. Despite the power consumption for the treated solution heating, this disadvantage is compensated with the series of important advantages, among them rapid sedimentation and separation of the forming sludge, its possible applications in other technologies, higher water treatment degree as compared with the conventional water treatment methods. Although there exists another alternative of ferrite wastewater treatment at room temperature (the so-called low temperature ferritization), this study will be devoted to the ferritization under elevated temperatures. The task of this research was to explore the specifics and kinetics of the high-temperature purification (under 65–70 °C and higher) of

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metal-containing wastewater and to study the composition and phase-disperse transformations of ferritized sludge. Besides, it was necessary to improve the efficiency of separation and dewatering of ferromagnetic sludge, specifically, applying magnetic fields to promote their utilization and to resolve the other relevant technological issues.

2

Materials and Methods

In our experiments, the electrochemical reactor— an electrocoagulator was used with a volume of 5 L and conical bottom to evacuate the sludge and with plate iron electrodes 25  25 cm. Current density was 1.0–5.0 A/dm2, pH varied within the range 4.0–8.5, and the treatment time 0.1–3.0 h. The influence of the additives of a series of anions was studied: PO43, CO32−, NO3− as well as the cations Ni2+, Zn2+, Cu2+, introduced in the amounts of 1–100 mg/L into the model solution containing (mg/L): Cl−—100, SO42−—100, Cr (VI)—50. Anionic and cationic composition was selected proceeding from the typical composition of the wastewaters from the plating processes applied in the industry. Treatment processes were studied within the temperature interval of 60– 95 °C. The phase-dispersion analysis was performed by counting particles of a given size under microscope (MBI-1). Fairly accurate characterization of particles distribution by size was guaranteed due to the large number of measurements. Chemical composition of solution and particles was performed by standard chemical and photo-chemical methods, following well-known techniques. X-ray studies were carried out on DRON-3 diffractometer. To assess the effect of ferrite treatment conditions on sediment properties, its magnetic characteristics were studied by measuring the magnetizability of suspensions, following the method described in [28], based on the integration of modification of magnetic flux in the studied environment under the changes of magnetic field.

3

Results and Discussion

3.1 Effect of Temperature and Wastewater Composition on Ferrite Treatment As mentioned above, the ferrite treatment of wastewaters, based on the introduction of Fe(II) sulfate and alkaline component, can be carried out either with heating of treated water to 65–70 °C, or under the room temperature conditions with subsequent separation of sediment and its heat (hydrothermal) ferritization. In both cases, ferromagnetic sediments are formed. However, although dealing with higher temperatures applied, the first method ensures achievement of better treatment indicators. According to the dispersion analysis data (Fig. 2), the particles size of the formed sediment is increased with the temperature increase of electrochemical ferrite treatment. Heating the forming suspension to 65 °C and higher under pH  6.5 is leading to the formation of blackcolored sediment, mainly composed of the Fe3O4.

3.2 MKM: Particles Size is Expressed in Microns Studies of the effect of ferrite treatment conditions on sediment properties have shown that with the increase of pH value to 8–10, the efficiency of sludge ferritization (detected by the formation of sludge with clearly expressed magnetic properties) has been increased, but with further increase in pH it has decreased, which is apparently connected with the amphotericity of iron. Under pH  4–5, the formation of Fe3O4 does not proceed. Therefore, the main factor that causes the increase in both magnetic properties and size of forming sediments in these processes is the elevated temperature. This observation is in good correlation with the magnetic characteristics of the sediments obtained, which confirmed the running of ferritization process (Fig. 3).

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Fig. 2 Integral (a) and differential (b) functions of the sediment particles distribution by size, depending on the process temperature: 1–20; 2–70; 3–85; 4–95 °C

These regularities did not change, when the salts of other heavy metals (sulfates of Ni, Zn and Co) were introduced into the model treated solution containing the iron salts. Formation of the ferrite particles under the elevated temperatures can be explained by the specifics of composition and properties of iron hydroxides, which always bear an electric charge [29]. Their charge appears due to the hydrogen ions and water molecules adsorbed on the iron hydroxides surface, which form a stable hydrate shell, whose central nucleus is the iron hydroxide molecule. Usually, metal hydroxides in the colloidal form absorb H+ ions on their surface and are charged positively. Due to the electrical

charge on their surface, these particles are not sticking together and diverge again. Once the pH value of the colloidal solution is shifted to the alkaline area, the charges of colloidal particles are neutralized with OH− groups. When the charge on the hydroxide particle’s surface disappears, these tend to diminish their surface under the surface attraction forces. As the result, the particles stick together, can interact with each other, and undergo the structural phase transformations. The other factor triggering the destruction of hydrate shells of Fe(II) and Fe(III) hydroxides is temperature, which increases up to 65–70 °C and creates the conditions of their interaction with each other with the formation of Fe3O4 particles, following the general type reaction: FeðOHÞ2 þ 2FeðOHÞ3 ! Fe3 O4 þ 4H2 O: In the presence of other heavy metal ions, the composite spinel structures may be formed, following the reactions: nMe2 þ þ ð3  nÞFe3 þ þ 6OH ! Men Fe3n ðOHÞ6; Men Fe3n ðOHÞ6 þ 1=2O2 ! Men Fe3n O4 þ 3H2 O:

Fig. 3 Effect of the ferrite process temperature and pH on the degree of magnetizability of the sediment formed: pH 1–7.5; 2–7.0; 3–6.5; 4–6.3; 5–4.5

It could be expected that the composition of the treated wastewater affects the properties of the sludge formed, first the magnetic characteristics of the suspensions. As follows from the data obtained (Fig. 3), with the increase in temperature and current density, the magnetic properties of suspensions obtained by

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the electrochemical treatment of model solution are increasing. Therefore, a practical conclusion could be made that to produce an Fe3O4 suspension with sufficiently high magnetic properties, the electromagnetic treatment (EMT) of wastewater must be carried out under elevated current density (ia) and temperature, to ensure the balance amount of iron ions in the solution for the sludge ferritization process. The optimal conditions for this process are: ia—up to 2.5–3.0 A/dm2 and temperature 70–80 °C to avoid the overconsumption of heat and electrical energy. The effect of Ni and Zn ions on the magnetic properties of suspensions formed in the electromagnetic treatment is expressed by the curves passing through the maximum (Fig. 4a, curves 1 and 2). With the increase of the concentrations of these ions in the treated water up to 20–40 mg/L, magnetizability of sludge is increased, reaching a maximum, and then is reduced. Such a change in magnetizability can be associated with the formation of spinel particles or solid solutions. It is known, for example, that the introduction of small amounts of foreign metal inclusions into the crystal lattice of Fe3O4 is used in the technology of ferrites manufacturing to enhance their magnetic properties. At the same time, magnetic properties of the sediments in the presence of copper ions in the treated water are decreasing with temperature increase. Apparently, under the conditions of the EMT process, these ions are not included in the

crystal lattice of Fe3O4 due to the large radius of copper ions (0.96 Å) as compared to nickel and zinc ions (0.69 and 0.74 Å, respectively), and the purification effect from copper can be associated only with the sorption or occlusion of particles of copper hydroxide, which is formed during the alkalization of the near-cathode space. It can also be assumed that the influence of Ni2+, Zn2+, Cu2+ cations on the Fe3O4 formation process is also manifested through the catalytic mechanism. However, nickel and zinc ions at a certain concentration in water promote the reaction of Fe3O4 formation due to the d-configuration of electrons in their atoms, and with further increase in concentration, they inhibit some stages of Fe3O4 formation process. Under these conditions, copper ions have only an inhibitory effect. It should also be noted that the literature describes the possibility of cations influencing the formation of Fe3O4 by controlling the rate of oxidation by changing the degree of water structure ordering [30]. The effect of PO43− and CO32− ions on the formation of the ferromagnetic properties of suspensions can be explained by the consumption of a part of Fe3+ ions during EMT for the primary formation of insoluble phosphates and carbonates, which block the active surface of iron hydroxide particles involved in the formation of Fe3O4. Due to this, the degree of hydroxides’ transformation into Fe3O4 decreases and, accordingly, the proportion of the ferromagnetic

Fig. 4 Dependence of the magnetic properties of the sludge formed in the process of electromagnetic treatment of wastewater under the temperature of 70 0C in the

presence of the cations (a) and anions (b): 1—Ni2+; 2— Zn2+; 3—Cu2+; 4—PO43−; 5—CO32−; 6—NO3−. Magnetic field strength—1200 Oersted

Combined Ferrite Treatment of Multi-component Wastewaters …

solid phase decreases. It should be noted a slight increase in the magnetization of suspensions with an increase in the concentration of NO3– ions in the treated water, which is probably due to their oxidative effect on Fe2+ ions, which favors the hydroxides’ transformation into Fe3O4. Analyzing the effect of impurity ions on the ferromagnetic properties’ formation of the suspension, it should be emphasized that, despite the decrease in their magnetization, the magnetic properties can be controlled by increasing the total amount of iron in the solution and selecting the appropriate values of EMT process parameters. To carry out electrolysis with converting of Fe (III) hydroxide into Fe3O4, an optimal power consumption of 3–5 Coulombs per 1 mg of Fe (III) contained in the sediment is required. Studies of the magnetic characteristics of these suspensions have shown that their magnetization as well as that of suspensions formed during the electrochemical treatment of hot wastewater increases with the increase of the magnetic field strength, reaching saturation at 1300–1500 Oersted. The value of the specific magnetic susceptibility of sediment in the dry state is 2.0–3.0 cm3/ g. It was found that under the optimal treatment conditions, making current density ia—up to 2.5– 3.0 A/dm2, temperature 70–80 °C, pH 6.0, the metal ions’ removal from treated solution reached to 98–100%, according to the data of chemical and photo-colorimetric analyses. Thus, the optimal condition for the formation of ferromagnetic properties of sediments in the EMT process is the temperature elevated to 65 ° C or higher. As our further studies of the process showed, magnetic properties of Fe3O4 suspension formed in this case continue to increase for some time after electrolysis ends and are stabilized in 0.3–0.5 h.

4

Electrochemical Ferrite Treatment of Heated Industrial Wastewaters

For the treatment of the iron-containing sludge resulted from the electrocoagulation treatment of wastewaters, it was proposed to apply high-

131

frequency currents. The process was tested under flow-through conditions using the laboratory set up, which scheme is presented on Fig. 5 [31]. As source of the high-frequency current (HFC), the standard generator VCK-2–100/0,066 was used. As a result of electrocoagulation treatment of multi-component wastewaters, due to the anodic dissolving of iron electrodes and running redox processes, suspension of metal hydroxides is formed, which are discharged into the flow pipe with a linear flow rate 0.2– 0.5 m/min. The current with the frequency of 60–74 kHz affects the suspension, and its action is enhanced due to the internal heating of metal rod inserted inside the pipe. As a result of this treatment, the rapid ferritization of the hydroxide sludge occurs, with the formation of black-colored suspension having the ferromagnetic properties. The measured electrical susceptibility of the sludge made 2000– 2200 Oersted, whereas the index of hydraulic fineness of sludge particles has increased from 0.3 to 0.86 mm/s. Despite the encouraging results thus obtained, this method has not been further used because of the high energy consumption.

Fig. 5 Scheme of the laboratory setup for the ferrite treatment of electrocoagulation wastewaters containing heavy metal ions in high-frequency current field: 1— electrocoagulator; 2—current source; 3—generator of HFC; 4—flow-through pipe made of non-conducting material; 5—metal rod [31]

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Study of the Phase Composition of Ferrite Sediments

Based on the available data on the ferrite process mechanism, one should expect a rather complex phase composition of sediments, since the various modifications of iron oxides and oxyhydrates can be present, as well as the so-called superparamagnetic particles, which are ferromagnetic in nature, but due to their small size, manifest themselves as paramagnets. In this regard, a comprehensive analysis of the phase-dispersed state of wastewater sludge with the content of Fe3O4, obtained both under the conditions of electrochemical ferritization and by separate ferritization of sludge, was carried out. On the X-ray diffraction patterns of sediment samples (Table 1), obtained during the electrochemical treatment at a temperature of  65 °C and a current density of up to 1.5 A/dm2, i.e., in the regime of electrocoagulation treatment, only a “halo” is detected, typical for the fine or

amorphous particles. Along with the increase in temperature and anodic current density over these values, the peaks corresponding to the interplanar distances of crystalline particles of iron oxygen compounds are unambiguously detected on the X-ray diffraction patterns. On the diffractograms from sediment samples, phases of Fe3O4 of a cubic structure with lattice parameter a = 8.396 Å and ferromagnetic oxide c-Fe2O3 also of a cubic structure (a = 8.339 Å) are most clearly identified. In addition, there are individual peaks that can be attributed to the iron oxyhydrate phases, specifically, c-FeOOH and b-Fe2O3∙H2O. In this case, it can be assumed that, under EMT conditions, formation of ferrite-type spinel, for example, FeCr2O4, is possible, but the peaks of this compound are not detected in the diffractograms of the sample obtained in the presence of only Cr(VI) ions. The X-ray diffraction study of the phase composition of sediments obtained with the

Table 1 Interplanar distances and line intensities on the sediments diffractograms Line intensity

Interplanar distances, Å

Sample I

Experimentally detected l value

Sample II

Sample I 13

Value from literature

Not identified

7.14 17

8

The same 6.81

4.82

9

12

3.37

3.77

8

6

3.015

3.008

The same 4.85

111

Fe3O4

3.33

220

b-Fe2O3∙H2O

(

Not identified 2:95 2:9967

34

2.95

8

2.770

2.78

5

2.68

2.64

5

6

Identification

Sample II

7.91

9

100

Miller indices (hkl)

2.568 100

2.514

2.501

24

2.36

2.404

2.55 8 > < 2:513 2:51 > : 2:532 ( 2:362 2:408

220 220

c-Fe2O3 Fe3O4

221

c-Fe2O3 b-Fe2O3∙H2O b-Fe2O3∙H2O

331 331

NiFe2O4 c-Fe2O3 Fe3O4

0.31 221

c-FeOOH NiFe2O4

Note Conditions of the electro-magnetite treatment process: Sample I—ia = 2.0 A/dm2; T—80 °C; pH  6.0; C(Cr(VI)) = 50 mg/L; Sample II—the same + C (Ni2+)—25 mg/L

Combined Ferrite Treatment of Multi-component Wastewaters …

simultaneous presence of chromium and nickel ions in the initial solution (sample II) had also been of interest. Deciphering the interplanar spacings of the reflection lines and their intensities made it possible to assume the possibility of the formation under the experimental conditions, along with other noted phases, of the NiFe2O4 compound with an inverted spinel structure having a cubic lattice with a = 8.339 Å [32]. The X-ray study of sediments obtained under EMT regimes in the presence of other heavy metal ions in model wastewater solutions, such as Zn2+, Cu2+ and Sn2+, did not reveal the presence of other spinel structures of the MeFe2O4 type. At the same time, the lattice parameters of the magnetite phase are distorted, which may be due to the formation of interstitial or substitutional solid solutions. The nature of the change in relative size of crystallites of the compounds that make up the sediments and the quantitative contents of phases can be traced by the degree of broadening and intensity of the most characteristic lines within reflection angles of 15–22° (Fig. 6). As can be seen from the data presented, the Xray diffraction patterns of the precipitates

Fig. 6 Change in the line intensity of the reflection lines of Fe3O4 (400), (311), and (200) in the wastewater sediments in dependence on their formation conditions: a —effect of pH (pH): 1—8–65; 2—4.45; 3—3,2 (T = 75 ° C, ia = 2 A/dm2); b—effect of anodic current density, A/dm2: 4—4.2; 5—2.4; 6—0.96 (T = 75 °C, pH 6.0); c —effect of temperature (°C): 1—90; 2—80; 3—70 (pH— 6.0, ia = 2 A/dm2)

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obtained at a process temperature of 70 °C are characterized by line broadening associated with the dispersity of the Fe3O4 crystallites. As the temperature rises to 80 and 90 °C, the degree of line broadening decreases and their height increases significantly, which characterizes an increase in the crystallites size of this phase. At the same time, there is a sharp decrease in the peaks height related to the unidentified phase, which is apparently paramagnetic. The results obtained correlate with the previously described measurements of the effect of temperature on the magnetic characteristics of sediments. An increase in the current density leads to the increase in the amount of Fe2+ ions entering the solution during the anodic dissolution of the electrodes. At ia = 1 A/dm2, the peaks on the Xray diffraction patterns are weakly expressed, which is a consequence of the insignificant content of crystalline phases in the sediments. This may be due to the insufficient content of iron ions for the formation of Fe3O4. Under these conditions, the weak lines appear from highly dispersed particles with interplanar spacings close to the lepidocrocytic c-FeOOH [33] or b-Fe2O3 [34] phase, which are paramagnets. However, there are some deviations of the interplanar distances from the literature values for these compounds as well as from the values of samples specially synthesized in this study. It can be assumed that this is a consequence of the formation of complex crystalline interstitial phases with participation of Cr(III) hydroxide, which distort the crystal lattice. With an increase in the anodic current density, a decrease (curve 5) and then disappearance (curve 6) of the reflection lines from the above paramagnetic phases are observed, with simultaneous increase in the peak’s height related to ferromagnetic oxides (c-Fe2O3 and Fe3O4). Thus, the data of X-ray phase analysis give a general idea of the composition of sediments formed during EMT, which correlate with the measurement results of the magnetic characteristics of sediments. Owing to their oxide-ferrite structure, such sediments have a higher chemical resistance against leaching [35] and real disposal routes.

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Conclusion

The plant kingdom generates various compounds which justify their application in human alimentation, health. Some used traditionally are object of deep scientific research use in water disinfection. Plant-based natural agents suggest a promising future for this field of study by considering the great demand for green and local products to face sustainability. The application of new phyto-disinfectant agents for drinking water is recommended to operate with an efficient, easy, and speed extraction technique well known by women, a stability assessment, toxicological evaluation, and an optimum storage conditions protocol. However, several technological concerns exist that require further research before the application of these products can reach full-scale implementation. Specifically, future research efforts should focus on advanced technologies to increase their yields, lower production costs, low-thermal reducing temperature and time, no toxic solvent, such as water, or ethanol. For best application of research findings chemistry, biology and technology and “green” innovative techniques are promoted to offer environmental benefits or to minimize.

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6. Thomas, G. P. (2012). A guide to the theory, properties and applications of magnetic fluid. 7. R Rosensweig 1987 Magnetic fluids Annual Review of Fluid Mechanics 19 437 463 8. C Vasilescu M Latikka KD Knudsen VM Garamus V Socoliuc R Turcu E Tombacz D Susan-Resiga RHA Ras L Vekas 2018 High concentration aqueous magnetic fluids: Structure, colloidal stability, magnetic and flow properties Soft Matter 14 6648 6666 9. H Shokrollahi 2013 Structure, synthetic methods, magnetic properties and biomedical applications of ferrofluids Materials Science and Engineering: C 33 5 2476 2487 10. R Chaniyilparampu P Pande P Kopčanský R Mehta 2001 Application of magnetic fluids in medicine and biotechnology Indian Journal of Pure and Applied Physics 39 683 686 11. HE Horng C-Y Hong SY Yang HC Yang 2001 Novel properties and applications in magnetic fluids Journal of Physics and Chemistry of Solids 62 9–10 1749 1764 12. Scherer, C., Figueiredo Neto, A. M. (2005). Ferrofluids: properties and applications. Brazilian Journal of Physics, 35(3A), 718–727. 13. J Aswathy M Suresh 2014 Ferrofluids: Synthetic strategies, stabilization, physicochemical features, characterization, and applications ChemPlusChem —A Multidisciplinal Journal Centering on Chemistry 79 10 1382 14. M Erdem F Tumen 2004 Chromium removal from aqueous solution by the ferrite process Journal of Hazardous Materials 109 1–3 71 77 15. SJ Keny AG Kumbhar A Sanjukta S Pandey 2014 Antimony sorption and removal on carbon steel/ magnetite surfaces in relation to pressurized heavy water reactors Current Science 106 8 1094 1100 16. Chamoli, P., Shukla, R., Bezbaruah, A., Kar, K., & Raina, K. (2021). Ferrites for water purification and wastewater treatment. In Ferrites and multiferroics. Engineering Materials Series (117-127). Springer. 17. RR Kefeni BB Mamba TAM Msagati 2017 Application of spinel ferrite nanoparticles in water and wastewater treatment: A review Separation and Purification Technology 188 399 422 18. V Sharma H Singh S Guleria N Bhardwaj S Puri SK Arya M Khatri 2022 Application of superparamagnetic iron oxide nanoparticles (SPIONs) for heavy metal adsorption: A 10-year meta-analysis Environmental Nanotechnology, Monitoring & Management 18 100716 19. D Demirel O Yenigün M Bekbölet 1999 Removal of Cu, Ni and Zn from wastewaters by the ferrite process Environmental Technology 20 9 963 970 20. N Masunga O Kelebogile KK Kefeni BB Mamba 2019 Recent advances in copper ferrite nanoparticles and nanocomposites synthesis, magnetic properties and application in water treatment: Review Environmental Chemical Engineering 7 3 103179

Combined Ferrite Treatment of Multi-component Wastewaters … 21. H Singh N Bhardwaj SK Arya M Khatri 2020 Environmental impacts of oil spills and their remediation by magnetic nanomaterials Environmental Nanotechnology, Monitoring & Management 14 100305 22. Qiao, R., Tian, W., Bai, J., Wang, L., Zhao, J., Du, Zh., & Gong, X. (2019). Application of magnetic adsorbents based on iron oxide nanoparticles for oil spill remediation: A review. Journal of the Taiwan Institute of Chemical Engineers, 97, 227–236. 23. SK Dutta M Akhter J Ahmed MK Amin PK Dhar 2022 Synthesis and catalytic activity of spinel ferrites: a brief review Biointerface Research in Applied Chemistry 12 4 4399 4416 24. M Sugimoto 2004 The past, present, and future of ferrites Journal of American Ceramic Society 82 269 280 25. M Hua SJ Zhang BC Pan WM Zhang L Lv QX Zhang 2012 Heavy metal removal from water/ wastewater by nanosized metal oxides: A review Journal of Hazardous Materials 211–212 317 331 26. Covaliov, V. V., Caraghiaur, V. G. (1985). Intensification of electrocoagulation wastewater treatment/ Electromagnetite treatment. Chisinau, VINITI, 85 (18). 27. E Barrado F Prieto M Vega F Fernandez-Polanco 1998 Optimization of the operational variables of a

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Magnetically Separable and Reusable Mag-Mg/Al Layered Double Hydroxides for the Adsorption of Disperse Dyes of Navy Blue and Yellow F3G Sri Juari Santosa, Lutfia Isna Ardhayanti, Desy Permatasari, and Narsito hydrogen bonding played an important role. The adsorption kinetics of both dyes followed better the second order than the pseudo-first and pseudo-second orders with rate constants (k2) 52.20 and 44.40 (mol/L)−1 min−1 for navy blue and yellow F3G, respectively. Adsorption isotherm of both dyes fitted more closely to Langmuir than to Freundlich models with adsorption capacity 0.55
10−3 mol/g (146.30 mg/g) for navy blue and 0.39
10−3 mol/g (110 mg/g) for yellow F3G. Adsorption trial to a synthetic wastewater sample showed that the adsorption efficiency of Mag-Mg/Al LDHs increased from 30.86 to 76.63% with decreasing the concentration of the mixed dyes from 400 to 80 mg/L. These adsorption efficiencies were enhanced to approximately 1.27 times (39.11 to 97.76%) upon the use of a secondary adsorbent obtained by calcination of the used adsorbent at temperature 450 °C for 3 h.

Abstract

Dying of a polyester fabric by using a mixture of disperse dyes of navy blue and yellow F3G is now getting popular in Indonesia for the purpose of developing a military camouflage costume for anti-near infrared (NIR) detection device in jungle field. Since the efficiency of fabric dying is generally low, a considerable high concentration of navy blue and yellow F3G dyes will be released to wastewater, and they may give detrimental effect to the environment. To anticipate this undesired problem that may arise soon, a composite of magnetite and Mg/Al layered double hydroxides (Mag-Mg/Al LDHs) by a co-precipitation method was synthesized and then applied as adsorbent of navy blue and yellow F3G dyes in wastewater. Magnetite was incorporated onto Mg/Al LDHs for the purpose of facilitating an easy separation of the adsorbent from the adsorption system by simply using an external magnetic field. It was observed that pH 4.0 was the best acidity for the adsorption of navy blue and yellow F3G dyes, and

S. J. Santosa (&)  D. Permatasari  Narsito Department of Chemistry, Universitas Gadjah Mada, Yogyakarta, Indonesia e-mail: [email protected] L. I. Ardhayanti Department of Environmental Engineering, Islamic University of Indonesia, Yogyakarta, Indonesia

Keywords







Adsorption Magnetite Mg/Al LDHs Navy blue Yellow F3G



1

Introduction

Nearly all colorants used nowadays to color objects belong to synthetic dyes. They are applied in everything from metals to plastics, from wood

© The Author(s), under exclusive license to Springer Nature Switzerland AG 2023 E. Debik et al. (eds.), Wastewater Management and Technologies, Water and Wastewater Management, https://doi.org/10.1007/978-3-031-36298-9_10

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to food, and from paper to clothes [1]. This fast growth of synthetic dyes usage for every human need is caused by the fact that they are cheap to synthesize, easy to apply, can be provided in a vast range of colors, and give brighter and better properties to the dyed materials [2]. These advantages make the synthetic dyes replace natural dye in nearly all applications. Like dye, high-cost natural cotton-based synthetic fabric in the textile world is now steadily replaced by low-cost synthetic polyester-based fabric. Nowadays, in the textile industry, polyesters take a major part of dye consumption [3]. As a result, the contribution of disperse dye as the common type of dye used in coloring polyester is also forecasted to grow significantly. Among the disperse dyes, navy blue and yellow F3G are increasingly well known in Indonesia. It is because of their current implementation for coloring polyester in a textile industry to produce a military camouflage costume for antinear infrared (NIR) detection device in a jungle field. As typically 10 to 15% of the dye will be lost in a commonly dyeing process [4], the wastewater of a textile industry producing this camouflage costume contains an appreciable number of navy blue and yellow F3G dyes. Dyes in wastewater will at least degrade the aesthetics of the environment and more seriously obstruct the sunlight in penetrating aquatic reservoir and hence disturbing photosynthesis of aquatic plants. They are generally reactive and could react with dissolved metal ions to form chelates that are frequently toxic to aquatic organisms [5]. Furthermore, synthetic dyes in wastewater are hardly biodegraded and resistant to light, heat, and chemicals, and consequently, they are potentially bio-accumulated, enter the food chain, and eventually endanger the human health [6]. Therefore, efforts to treat the wastewater of textile industry before discharging to the environment are essentially important. To diminish the content of the dyes in the wastewater, a variety of methods were developed, such as ion exchange [7], solvent extraction [8], photocatalytic degradation [9], biosorption [10], and adsorption [11–13]. Compared to the others, adsorption possesses some

S. J. Santosa et al.

advantages because it is easily handled, simple in design, highly efficient, and able to be effectively operated in even very low concentrations of dye. Furthermore, adsorption is commonly known as a cheap method since it is possible to use lowcost adsorbents and to regenerate and reuse them. In a previous study of the authors, the possibility of Mg/Al layered double hydroxides (Mg/Al LDHs) as adsorbents for navy blue and yellow F3G dyes was investigated and found that the Mg/Al LDHs were effective to remove those two dyes from aqueous solution [14]. After being used, the Mg/Al LDHs could be reused by simple calcination. Further adsorption tests showed that the calcined Mg/Al LDHs even had higher adsorption ability in adsorbing navy blue and yellow F3G than the original Mg/Al LDHs [15]. In opposition to these advantages, there is a remaining drawback that the separation of Mg/Al LDHs from adsorption medium is still difficult, and it will hamper their large-scale application. To overcome this drawback, a combination of the Mg/Al LDHs with a magnetite was developed to obtain a composite material of magnetiteMg/Al LDHs (Mag-Mg/Al LDHs). Preliminary results showed that the Mag-Mg/Al LDHs had better performance than the Mg/Al LDHs in terms of adsorption ability and easiness to be separated from adsorption medium by simply using an external magnetic field [16]. In the present study, the performance of MagMg/Al LDHs as the adsorbent of navy blue and yellow F3G was further explored, adsorption kinetics and isotherms were determined, and the adsorption mechanism was elucidated. Moreover, trial test to adsorb the dyes from an artificial wastewater of a textile industry and to reuse the adsorbent after calcination was examined.

2

Materials and Methods

2.1 Materials All chemicals, i.e., Mg(NO3)2⋅6H2O, Al(NO3)3⋅ 9H2O, FeSO4⋅7H2O, FeCl3⋅6H2O, NaOH, and CH3COOH, were analytical grade, purchased from Merck Co. Inc. (Germany) and used

Magnetically Separable and Reusable Mag-Mg/Al Layered Double …

:O:

139 OH

: NH

N

:O:

O

: NH

Fig. 1 Chemical structure of navy blue (left) and yellow F3G (right) dyes

without further purification. Navy blue and yellow F3G dyes (Fig. 1) [14] were provided by CV. Maju Mapan, a Textile Industry in Tulungagung, East Java, Indonesia. Nitrogen gas was bought from CV. Perkasa, Yogyakarta, Indonesia.

2.2 Instrumentation

at a constant rate to the prepared mixed solution of FeCl3⋅6H2O and 1.789 g FeSO4 at 70 °C under continuous stirring until pH 10. After aging for 30 min, the formed magnetite was separated from the medium of reaction by pouring out the solution while retaining the magnetite inside the beaker glass using an external magnet. The separated magnetite was washed with distilled water until the filtrate was neutral. The separated magnetite was dried at 70 °C to constant weight, crushed and sieved to pass through a 200 mesh of sieve apparatus.

The main apparatuses used in this study were consisting of analytical balance (Mettler Toledo AL204), stirrer and hot plate (Nouva), electronic pH meter (Hanna Instrument 211), centrifuge (K (b) Mg/Al LDHs PLC series), oven (Fischer Scientific model 655 F), siever 100 and 200 mesh, external magnetic Mg/Al LDHs were prepared by a co-precipitation field, and shaker. Analytical instruments used technique as done in previous studies [17, 18]. included X-ray diffractometer (Shimadzu XRD- First, 12.8 g Mg(NO3)2⋅6H2O and 9.4 g Al 6000), Fourier transform Infrared spectrometer (NO3)3⋅9H2O were dissolved in 100 mL of dis(Shimadzu FT-IR Prestige 21), vibrating sample tilled water, followed by addition of NaOH magnetometer (type OXFORD VSM 1.2 H), and solution 0.5 M to pH 10.0 under N2 atmosphere. UV–visible spectrometer (UV-1700 Pharmaspec). The mixture was then treated hydrothermally at 120 °C for 5 h; the suspension was filtered, washed, and then dried at 70 °C to constant 2.3 Preparation of Materials weight. (a) Magnetite

(c) Mag-Mg/Al LDHs

Magnetite was prepared by first dissolving 3.254 g FeCl3⋅6H2O and 1.789 g FeSO4⋅7H2O in 100 mL of distilled water in a closed beaker glass, deoxygenating the beaker glass under N2 atmosphere for 30 min and maintaining the temperature of solution in the beaker glass at 70 °C. A NaOH solution 0.5 M was then added

Mag-Mg/Al LDHs were synthesized first by dispersing 1.3 g magnetite in 100 mL of distilled water. Second, a mixed solution of Mg and Al nitrates is prepared by dissolving 12.8 g Mg (NO3)2⋅6H2O and 9.4 g Al(NO3)3⋅9H2O in 100 mL of distilled water. Third, NaOH solution is prepared by dissolving 6.6 g NaOH in 100 mL

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of distilled water. The mixed solution of Mg and Al nitrates and the solution of NaOH were added at a constant rate to the stirred dispersion of the magnetite under N2 atmosphere to pH 10.0. The resulting suspension was hydrothermally treated at 120 °C for 5 h. After cooling down to room temperature, the suspension was filtered, washed, and then dried at 70 °C to constant weight.

3

Adsorption Experiment

3.1 Effect of Medium pH 25 mg Mg/Al LDHs was added into every 25 mL of navy blue solution (26.8 mg/L = 0.10 mmol/L) at initial pH 3.0, 4.0, 5.0, 6.0, and 7.0. CH3COOH solution was used to adjust the acidity of the navy blue solutions. The mixed solutions were then stirred for 120 min, filtered, and the remaining navy blue in the filtrate was determined using UV–vis spectrometer at the wavelength giving maximum absorbance (kmax = 602 nm). The same experiment was repeated as control, but the sample solutions contain no Mg/Al LDHs. As comparison, the same experiment was conducted, but Mag-Mg/Al LDHs were used to replace Mg/Al LDHs. The separation of MagMg/Al LDHs was done after filtrating the solution through a 0.45-lm membrane paper, at the same time with the help of an external magnet to retain the Mag-Mg/Al LDHs inside the reaction flask. In addition to navy blue, Mg/Al LDHs and Mag-Mg/Al LDHs were also employed to adsorb yellow F3G dye. The adsorption conditions used were the same as that for navy blue, but the quantification of yellow F3G was conducted at kmax = 410 nm.

3.2 Effect of Contact Time The effect of contact time was performed by first dispersing 25 mg adsorbent in 25 mL navy blue with the concentration of 0.10 mmol/L. The initial pH was adjusted at the optimum pH obtained from experiment 3.1. The mixture was

shaken at contact times of 5, 15, 30, 60, 90, 120, and 180 min. The suspension was filtered, while Mag-Mg/Al LDHs were retained inside the reaction flask by using an external magnetic field. The concentration of navy blue remaining in the filtrate was determined using UV–vis spectrometer at kmax = 602 nm. Every sample solution was accompanied by a blank solution containing no adsorbent. Like navy blue, the same experiment was conducted for yellow F3G dye. The concentration of yellow F3G remained in the filtrate was also quantified by using UV–vis spectrometer at kmax = 410 nm.

3.3 Effect of Initial Concentration The effect of initial concentration of dye was examined by contacting 25 mg Mag-Mg/Al LDHs with 25 mL navy blue at initial concentration of 10, 30, 60, 90, 120, 150, 180, 210, and 240 mg/L. The initial pH and contact time were adjusted at the optimum conditions obtained from experiments 3.1 and 3.2, respectively. The mixture was then filtered while retaining the Mag-Mg/Al LDHs by external magnetic field. The navy blue concentration remained in the filtrate was determined using UV–vis spectrometer at kmax = 602 nm. The same experiment for the sample containing no adsorbent was conducted as the control. As navy blue, the same experiment was conducted for yellow F3G dye. However, the concentration of yellow F3G remained in the filtrate was quantified by using UV–vis spectrometer at kmax = 410 nm.

3.4 Trial Test of Synthetic Wastewater of the Textile Industry Synthetic wastewater was prepared through a hydrothermal process of a 2
2 cm polyester fabric dying in 25 mL of mixed solution containing navy blue 0.267 mg/mL and yellow F3G 0.533 mg/mL for 45 min at 130 °C and pH 5.0. Total concentration of the two dyes was 800 mg/L or 0.80 mg/mL.

Magnetically Separable and Reusable Mag-Mg/Al Layered Double …

Adsorption of the remaining navy blue and yellow F3G in the synthetic wastewater by using Mag-Mg/Al LDHs was done at pH 5.0 for as long as 120 min. Before the adsorption was done, the synthetic wastewater was first diluted for 2, 4, 6, 8, 10, and 12 times. From every diluted wastewater, 25 mL were taken and 25 mg Mag-Mg/Al LDHs were added into this wastewater to perform the adsorption. After separating the filtrate from Mag-Mg/Al LDHs, concentration of the remaining dye in filtrate was determined using UV–vis spectrometer at kmax = 602 nm for navy blue and 410 for yellow F3G.

3.5 Regeneration and Reuse of Adsorbent After being used, the adsorbent was calcined at 450 °C for 3 h [16]. The calcined adsorbent was then reused to adsorb again navy blue and yellow F3G in diluted synthetic wastewater samples, and the concentrations of remaining navy blue and yellow F3G were determined using UV–vis spectrometer at kmax = 602 and 410 nm, respectively.

4

Results and Discussion

4.1 Characterization of Materials The preparation of magnetite and Mg/Al LDHs was done by using co-precipitation the method. As indicated by the Pourbaix diagram that the formation of magnetite started from pH 9 to pH 11 [19], the preparation of magnetite in this study was done at pH 10.0. Similarly, the Pourbaix diagram showed that the formation of Mg/Al LDHs from a mixture of Mg and Al nitrates at mole ratio 2:1 occurred from pH 9.5 to pH 14.0. Accordingly, the preparation of Mg/Al LDHs in this study was performed at pH 10.0. As it has been determined in author’s previous study, the general formula of the synthesized Mg/Al LDHs was Mg0.75Al0.25(OH)2(NO3)0.250.32H2O [18]. The preparation of Mag-Mg/Al LDHs was done

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by adding the mixed solution of Mg and Al nitrates at molar ratio 2:1 to the dispersion of magnetite and followed by the addition of NaOH 0.5 M at a constant rate and under continuous stirring until pH 10.0. Visual appearance of the synthesized materials is displayed in Fig. 2. Mg/Al LDHs are a white powder, while magnetite and Mag-Mg/Al LDHs are black and deep brown in color, respectively.

4.2 Structural Characterization The XRD pattern of Mg/Al LDHs showed main characteristic peaks at 2h 11.23, 22.79, and 34.55° (Fig. 3a), which corresponded to d003, d006, and d012 peaks, respectively. The intensity of d003 is approximately twofold higher than that of d006 (Fig. 3a) indicating that the Mg/Al LDHs have a layered structure [21] with the calculated basal spacing 7.89 Å. It is a closely resembled XRD pattern of Mg/Al LDHs, and relatively the same basal spacing had been found in previous studies [17, 20]. In addition to d012, other nonbasal peaks of Mg/Al LDHs were revealed by the appearance of d015, d018, d110, and d013 at 2h 38.31, 46.85, 60.33, and 62.34, respectively (Fig. 3a). The peak at 2h 60.33° (d110) indicates the lattice parameter a of LDH, which coincides with the closest M-M distance in the brucite-like layers [22]. A series of characteristic peaks at 2h 30.22, 35.59, 43.28, 56.88, and 62.21° have been displayed by the XRD pattern of magnetite (Fig. 3 b). According to the Joint Committee on Powder Diffraction Standards (JCPDS) card number 89.0691, those peaks correspond to d220, d311, d400, d511, and d440. Additionally, there was a small peak at 2h 31.87° according to the JCPDS card number 04-0755 that matches with the peak of a maghemite impurity. The XRD pattern of Mag-Mg/Al LDHs was apparently the result of simple combination between that of magnetite and Mg/Al LDHs (Fig. 3c). In the diffraction pattern of MagMg/Al LDHs, peaks originally from Mg/Al LDHs were still clearly observed and those originally from magnetite were also obviously

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Fig. 2 Visual appearance of the synthesized Mg/Al LDHs (left), magnetite (middle), and Mag-Mg/Al LDHs (right)

LDH

Mag-LDH Magnetite

Fig. 3 XRD diffractogram of a Mg/Al LDHs, b magnetite, and c Mag-Mg/Al LDHs

Magnetically Separable and Reusable Mag-Mg/Al Layered Double …

apparent concomitantly. Peak at 2h 34.55° from Mg/Al LDHs was partly overlapped with that at 2h 35.59° from magnetite. Similarly, peak at 2h around 63° from Mg/Al LDHs was also overlapped with that at 2h 62.21° from magnetite. Furthermore, the presence of magnetite affected insignificantly the basal peak of Mg/Al LDHs at 2h 11.23° indicating that the most probable position of magnetite was not in the interlayer of Mg/Al LDHs.

4.3 Functional Group Characterization Characterization the functional groups using FT-IR spectroscopy showed that the synthesized Mg/Al LDHs possessed main absorption bands at wavenumbers 420, 648, 1381, 1636, and 3464 cm−1 (Fig. 4a). Absorption bands at 420 and 648 cm−1 were contributed by the stretching vibration of Mg-O and Al-O, respectively. Sharp absorption band at 1381 cm−1 was attributed to the stretching vibration of N–O, indicating the presence of nitrate ion, instead of carbonate ion, in the interlayer of Mg/Al LDHs [18, 23]. When carbonate ion replaces nitrate ion as the interlayer ion in Mg/Al LDHs, a sharp absorption band at wavenumber of about 1360 cm−1 will appear due to the stretching vibration of C-O [24]. Absorption bands at 1636 and 3464 cm−1 corresponded to bending vibration of H-OH from the adsorbed water and stretching vibration of OH on the surface of Mg/Al LDHs. Magnetite possessed a distinctive absorption band at the wavenumber 586 cm–1, which attributed to the stretching vibration of Fe–O (Fig. 4b) [25]. Along with this distinctive band, there were weak absorption bands at wavenumbers of 1628 and 3426 cm–1. As observed for Mg/Al LDHs in Fig. 4a, these absorption bands were related to bending vibration of H-OH from the adsorbed water and stretching vibration of OH on the surface of magnetite. As described it previously in accordance with Pourbaix diagram, magnetite settles as a stable phase in alkaline aqueous medium, where the formed magnetite acquires negative charge on its

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surface. On the one hand, Mg/Al LDHs are a positively charged lamellar material with pHPZT 12.0 [26]. Therefore, magnetite and Mg/Al LDHs should be able to interact with each other to form a stable hybrid of Mag-Mg/Al LDHs. The most potential interaction between magnetite and Mg/Al LDHs occurs through electrostatic interaction between oxygen of magnetite (Fe3O4) and metal components (Mg and Al) of LDHs. If this interaction really occurs and is stable enough to be detected by FT-IR, their O–Mg and O–Al stretching vibrations will be potentially overlapped with the existing Mg–O and Al–O stretching vibrations originated from Mg/Al LDHs. In fact, FT-IR spectra of Mag-Mg/Al LDHs revealed the appearance of absorption bands at 427 and 648 cm−1 (Fig. 4c) that might be interpreted as the overlapping result of those stretching vibrations. The absorption band at 586 cm–1 originated from magnetite still appeared in the FT-IR spectra of Mag-Mg/Al LDHs. FT-IR spectra of Mag-Mg/Al LDHs displayed a sharp absorption band at 1381 cm–1 indicating that nitrate ion was still in the interlayer of Mg/Al LDHs [18]. Thus, it is likely that the most suitable position of magnetite is not in the interlayer of Mg/Al LDHs, and hence, it supports the above-mentioned prediction deduced from the result of XRD characterization.

4.4 Magnetic Characterization Magnetization curves of magnetite and MagMg/Al LDHs measured using VSM at room temperature showed that both materials were paramagnetic with zero coercivity (Fig. 5). The specific saturation magnetization (Ms) of magnetite, i.e., 74.6 emu/g (Fig. 5a), was significantly higher than that of Mag-Mg/Al LDHs (31.4 emu/g) (Fig. 5b), and therefore, the magnetite may be classified as superparamagnetic. It may also be deduced that the particle size of magnetite in its association with Mg/Al LDHs will be smaller than that in its pure form, since Ms value decreases with decreasing particle size due to the reduction in crystallinity of the magnetic domain [27].

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Fig. 4 FT–IR spectra of a Mg/Al LDHs, b magnetite, and c Mag-Mg/Al LDHs

586

1636

Transmitance (a.u)

427

(c)

648 1381

Fe-O)

586 ( 420 (

3464

1628 (

OH)

(b) 3426 (

Mg-O)

OH)

(a) 1636 ( 648 ( Al-O)

0

1000

1381 (

OH)

3464 (

N-O)

2000

3000

OH)

4000

Wavenumber (cm -1)

Fig. 5 Measured magnetization curves of a magnetite and b Mag-Mg/Al LDHs

4.5 Adsorption of Dye (a) Effect of Medium pH Medium pH is an important parameter in adsorption, since pH may affect the ionization degree of the functional groups of adsorbent and

adsorbate. The effect of this medium pH on the adsorption of navy blue and yellow F3G is presented in Fig. 6. The adsorption was not conducted at pH below 3.0 due to the instability of adsorbents toward dissolution in aqueous medium [17]. As displayed in Fig. 6a and b, the adsorption of navy blue and yellow F3G by both magnetite and Mag-Mg/Al LDHs was maximum at initial pH 4.0. It was observed that the pH value of the dye solutions always increased after the dye solution was contacted with the adsorbent. It means, there is a net flow of H+ from the solution to protonate the adsorbent. The similar evidence has also been observed previously by Ikhsani et al. [14]. The protonated sites of adsorbent should play an important role as an active site to interact with the dye adsorbate, and the interaction might be governed by hydrogen bonding to the remaining deprotonated oxygen atoms in dyes [16]. As the initial pH of dye solution increases to 4.0, the degree of protonation of adsorbent and adsorbate decreases to give a more appropriate condition to result in the most effective hydrogen

Magnetically Separable and Reusable Mag-Mg/Al Layered Double …

showed that navy blue was more adsorbed than yellow F3G by Mg/Al LDHs [14, 16]. The data in Fig. 7 were then evaluated by using kinetic models of pseudo-first order (Eq. 1) [28], pseudo-second order (Eq. 2) [29], and second order (Eq. 3) [30]. The evaluation was based on linearity (R2 value) of the plots of ln(1−q/qe) versus t from Eq. (1), t/qt versus   0 Ce Þ t from Eq. (2) and C1e ln CC0ðCðCC versus t from eÞ

Adsorbed Navy Blue (mmol/g)

0.6 0.5 0.4 0.3 0.2 Mg/Al LDHs 0.1

Mag-Mg/Al LDHs

0 0.5 Adsorbed Yellow F3G (mmol/g)

145

Eq. (3) for the pseudo-first order, pseudo-second order, and second-order kinetics models, respectively. As given in Fig. 8 and Table 1, the   0 Ce Þ plots of C1e ln CC0ðCðCC versus t (Eq. 3) gave Þ e

0.4

0.3

0.2 Mg/Al LDHs 0.1

better linearity than the two other plots.

Mag-Mg/Al LDHs

  qt ¼ k1p t ln 1  qe

0 2

3

4

5

6

7

8

Initial pH

t 1 1 ¼ þ t qt k2p q2e qe    1 CðC0  Ce Þ ln ¼ k2 t; Ce C0 ðC  Ce Þ

Fig. 6 Effect of medium acidity on the adsorption of navy blue (above) and yellow F3G (below) on Mg/Al LDHs and Mag-Mg/Al LDHs

bonding between adsorbent and adsorbate. As the pH increases further above 4.0, the degree of positive charge of the adsorbent and adsorbate further decreases and the availability of H+ also becomes more limited, so the interaction is more inhibited. Therefore, the adsorption decreases with increasing pH above 4.0. As Mag-Mg/Al LDHs consistently adsorb more navy blue and yellow F3G than Mg/Al LDHs itself, the further adsorption examinations were only be focused on Mag-Mg/Al LDHs.

ð1Þ ð2Þ ð3Þ

where qe and qt are the amount of dye adsorbed on adsorbent (mol/g) at equilibrium and at contact time t, respectively, C0, Ce, and Ct are the initial concentration of dye in solution, the concentration of dye remaining in solution at equilibrium, and the concentration of dye remaining in solution (each mol/L) at contact time t, respectively, and k1p, k2p, and k2 are pseudo-first

The adsorption profiles of navy blue and yellow F3G by Mag-Mg/Al LDHs were very similar (Fig. 7). The adsorption of both dyes was initially fast and started decelerating from the contact time of approximately 90 min and then reached relatively constant from contact time 120 min. Although the adsorption profiles were very similar, but Mag-Mg/Al LDHs had more ability to adsorb navy blue than yellow F3G at all contact times studied. Similarly, other studies had also

Adsorbed Amount (mmol/g)

0.6

(b) Effect of Interaction Time

0.45

0.3

Navy Blue

0.15

Yellow F3G 0 0

50

100

150

200

Contact Time (min)

Fig. 7 Effect of contact time on the adsorption of navy blue and yellow F3G on Mag-Mg/Al LDHs

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order, pseudo-second order, and second-order rate constants, respectively. The second order was formulated based on the remaining concentration of adsorbed species in the solution, while the two other kinetic models were formulated based on the number of adsorbed species on the adsorbent. Since the secondorder kinetic model was the best kinetic model in evaluating data in Fig. 7, the adsorptions of navy blue and yellow F3G were governed by two parameters, presumably the concentration of the dyes and the availability of active sites of the adsorbent. Additionally, since the second order was more linear than the pseudo-first and pseudo-second-order kinetic models, it means

ln(1-qt/qe)

0 -1 R² = 0.9735 -2 Navy Blue Yellow F3G

-3

R² = 0.9901

that all removed species from the solution are not completely attracted chemically on the surface of adsorbent. As described before, hydrogen bonding plays an important role on the adsorption of navy blue and yellow F3G on the Mag-Mg/Al LDHs, but the hydrogen bonding is not a strong interaction. Furthermore, Mag-Mg/Al LDHs are a lamellar material, which has interlayer and outer surfaces with different adsorption affinity toward an adsorbate.   0 Ce Þ From the slope of the plot of C1e ln CC0ðCðCC eÞ versus t from Eq. (3), shown in Fig. 8 (below), the obtained k2 was 52.20 and 44.40 (mol/L)–1 min–1 for navy blue and yellow F3G, respectively (Table 1). Thus, navy blue is more rapidly adsorbed than yellow F3G by Mag-Mg/Al LDHs. This may be, because yellow F3G has a relatively bigger molecular weight and size than navy blue. As simulated previously by using Gauss View 4, the molecular dimension of navy blue and yellow F3G was 10.73
9.80
3.90 (Å)3 and 13.28
7.78
5.12 (Å)3, respectively [14].

-4 400

t/qt

4.6 Effect of Concentration

R² = 0.9622

300 200

R² = 0.9922

100 0 8

R² = 0.9984

Y

6 4

R² = 0.9972 2 0 0

30

60

90

120

150

t (min) Fig. 8 Plots of ln(1−q/qe) versus t from the pseudo-first order (above), t/qt versus t from pseudo-second order (middle), and Y versus t from kinetics  second-order  models (below). Y denotes

1 Ce

ln

C ðC0 Ce Þ C0 ðCCe Þ

The adsorption profile of navy blue and yellow F3G on Mag-Mg/Al LDHs as a function of concentration of the dye at equilibrium (Ce) increased quickly at the early stage and changed to increase gradually, when the equilibrium concentration reached a value of approximately 0.12 and 0.11 mmol/L for navy blue and yellow F3G, respectively (Fig. 9). The application of the Langmuir (Eq. 4) and Freundlich (Eq. 5) isotherm models to the experimental data showed that the adsorption of both navy blue and yellow F3G on Mag-Mg/Al LDHs followed better the Langmuir than Freundlich isotherm model. This is an indication that Mag-Mg/Al LDHs have homogeneous surfaces with a uniform sorption energy. Adsorption capacities of Mag-Mg/Al LDHs for navy blue and yellow F3G according to the Langmuir isotherm model were 0.55 and 0.40 mmol/g, respectively. Before being composited with magnetite, Mg/Al LDHs had

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Table 1 Pseudo-first order (k1p), pseudo-second order (k2p), and second-order (k2) rate constants for the adsorption of navy blue and yellow F3G on Mag-Mg/Al LDHs Pseudo-first ordera

Dye

Pseudo-second orderb

Second orderc

k1p

R2

k2p

R2

k2

R2

Navy blue

0.028

0.9901

25.88

0.9922

52.20

0.9984

Yellow F3G

0.019

0.9735

8.96

0.9622

44.40

0.9972

Lagergren [28], k1p in min–1 b Ho and McKay [29], k2p in (mol/g)–1 min–1 c Santosa [30], k2 in (mol/L)–1 min–1

a

adsorption capacities of 5.23 and 1.64
10– 2 mmol/g for navy blue and yellow F3G, respectively [14]. It means that the presence of magnetite in the composite of Mag-Mg/Al LDHs improves the adsorption capacities of Mg/Al LDHs for navy blue by more than tenfold and for yellow F3G by more than 24-fold. m¼

bKCe Ce 1 Ce þ ¼ or 1 þ KCe m bK b

m ¼ BCe1=n or log m ¼ log B þ

1 log Ce ; n

ð4Þ ð5Þ

where Ce is the equilibrium concentration of dye (mol/L) in solution, b is the Langmuir’s adsorption capacity (mol/g), K is the equilibrium constant (L mol)–1, m is the adsorbed dye onto MagMg/Al LDHs (mol g–1), B is the Freundlich’s removal capacity (mol/g), and n is a constant.

4.7 Trial Test to a Synthetic Wastewater The synthetic wastewater initially contained 800 mg/L dyes, which consisted of 266.67 mg/L navy blue and 533.33 mg/L yellow F3G. Adsorption was done at pH 4.0 for as long as 120 min. Before the adsorption was conducted, the synthetic wastewater was first diluted for 2, 4, 6, 8, 10, and 12 times using CH3COOH solution 0.001 M at pH approximately 4.0, so the total initial concentration of the dyes in the sample of wastewater was 400, 200, 133, 100, and 80 mg/L, respectively. From each diluted wastewater was taken 25 mL, and 25 mg MagMg/Al LDHs was added to perform the

adsorption. After separating the filtrate from Mag-Mg/Al LDHs, concentration of the remaining dye in filtrate was determined using UV–Vis spectrometry at kmax = 430 nm. After the adsorption, the adsorbent was calcined at 450 °C for 3 h and then reused to adsorb again navy blue and yellow F3G in the diluted synthetic wastewater samples. As shown in Fig. 10, the calcined Mag-Mg/Al LDHs was always more powerful to adsorb the mixed dyes for all synthetic wastewater samples. The adsorption ability of Mag-Mg/Al LDHs and the calcined Mag-Mg/Al LDHs was 123.44 and 156.43 mg/g, respectively, when the concentration of the mixed dyes was 400 mg/L. These adsorption abilities decreased gradually with decreasing concentration of the mixed dyes in the synthetic wastewater sample and reached values of 61.30 and 77.89 mg/g, respectively, when the concentration of the mixed dyes was 80 mg/L. In average, the adsorption ability of the calcined Mag-Mg/Al LDHs was 1.27 times higher than that of Mag-Mg/AL LDHs. The efficiencies of the adsorption of mixed dyes by Mag-Mg/Al LDHs and calcined MagMg/Al LDHs increased with decreasing concentration of the dyes in the sample solution. At a total initial concentration of dyes 400 mg/g, adsorption efficiencies were 30.86 and 39.11% for Mag-Mg/Al LDHs and calcined Mag-Mg/Al LDHs, respectively. Adsorption efficiencies increased with decreasing concentration of mixed dyes in the synthetic wastewater sample and reached the value of 76.63 and 97.36% for MagMg/Al LDHs and calcined Mag-Mg/Al LDHs, respectively, when the adsorbents were applied to the sample containing 80 mg/L of dyes.

148

0.0009

Adsorbed Navy Blue (mol/g)

Fig. 9 Adsorption data of navy blue (above) and yellow F3G (below) on Mag-Mg/Al LDHs as a function of the remaining dye concentration at equilibrium (Ce) as well as the adsorption profiles modeled according to the Langmuir and Freundlich isotherm models

S. J. Santosa et al.

0.0006

0.0003

Experimental Data Langmuir Model Freundlich Model

0 0

0.0005

0.001

0.0015

0.002

Adsorbed Yellow F3G (mol/g)

0.0006

0.0004

Experimental Data

0.0002

Langmuir Model Freundlich Model 0 0

0.00025

0.0005

0.00075

0.001

Ce (mol/L) The consistently higher ability of the calcined Mag-Mg/Al LDHs than the original Mag-Mg/Al LDHs in adsorbing the mixed dyes may be caused by the memory effect owned by Mg/Al LDHs. After calcination, anions and water, which were initially adsorbed or intercalated in MagMg/Al LDHs, disappear and the lamellar structure of material collapses. Upon using the calcined product as adsorbent for the synthetic wastewater sample, a lamellar structure of MagMg/Al LDHs will be reconstructed and the dyes

in the sample are entrapped inside the interlayer of the reconstructed lamellar structure. Additionally, the dyes may also be adsorbed on the outer surface of the reconstructed Mag-Mg/Al LDHs. These entrapping and adsorption phenomena should be the reason why that the calcined Mag-Mg/Al LDHs always consistently possesses higher ability in removing the mixed dyes from the synthetic wastewater sample. The memory effect of LDHs to reconstruct to their initial structure has been frequently documented [31, 32].

Magnetically Separable and Reusable Mag-Mg/Al Layered Double …

Mag-Mg/Al LDHs Calcined Mag-Mg/Al LDHs Mag-Mg/Al LDHs Calcined Mag-Mg/Al LDHs

200

100

150

75

100

50

50

25

Adsorbed Dyes (%)

Adsorbed Dyes (mg/g)

Fig. 10 Adsorption ability and adsorption efficiency of Mag-Mg/Al LDHs and calcined Mag-Mg/Al LDHs at various initial concentrations of dyes

149

0

0 400

200

133

100

80

Initial Concentration of Dyes (mg/L)

5

Conclusion

A magnetically separable and reusable Mag-Mg/ Al LDH has been successfully prepared through the co-precipitation method using NaOH as precipitating agent. The results of structural and functional group characterizations indicated that magnetite particles were not distributed in the interlayer of Mg/Al LDHs. The application of Mag-Mg/Al LDHs as adsorbent showed that the adsorption of navy blue and yellow F3G had an optimum at pH 4.0. At this optimum pH, the adsorption kinetics followed the second-order model with the adsorption rate constant (k2) 52.20 (mol/L)–1 min–1 for navy blue and 44.40 (mol/L)–1 min–1 for yellow F3G. The adsorption isotherm matched the Langmuir model with the adsorption capacity 146.30 and 110 mg/g for navy blue and yellow F3G, respectively. Hydrogen bonding likely took an essential role for the adsorption of both dyes. Application to a synthetic wastewater sample containing 80 mg/L of the mixture of both dyes showed that 76.63% of those dyes was successfully adsorbed on Mag-Mg/Al LDHs. The adsorption efficiency increased to 97.36% when the adsorbent was reused after it was calcined at temperature

450 °C for 3 h. These efficiencies degraded, when the Mag-Mg/Al LDHs and calcined MagMg/Al LDHs were applied to higher concentrated synthetic wastewater samples. Eventually, this study shows the important advantage of Mag-Mg/Al LDHs that could be easily reused after calcining the adsorbent used.

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S. J. Santosa et al. 18. Santosa, S. J., Fitriani, D., Aprilita, N. H., & Rusdiarso, B. (2020). Gallic and salicylic acidfunctionalized Mg/Al hydrotalcite as highly effective materials for reductive adsorption of AuCl4−. Applied Surface Science, 507, 145115. 19. El-kharrag, R., Amin, A., & Greish, Y. E. (2012). Low temperature synthesis of monolithic mesoporous magnetite nanoparticles. Ceramics International, 38, 627–634. 20. Santosa, S. J., Kunarti, E. S., & Sudiono, S. (2010). Journal Ion Exchange, 21, 272–277. 21. Yang, K., Yan, L.-G., Yang, Y.-M., Yu, S.-J., Shan, R.-R., Yu, H.-Q., Zhu, B.-C., & Du, B. (2014). Adsorptive removal of phosphate by Mg-Al and ZnAl layered double hydroxides: Kinetics, isotherms and mechanisms. Separation and Purification Technology, 124, 36–42. 22. Mahjoubi, F. Z., Khalidi, A., Abdennouri, M., & Barka, N. (2015). Zn–Al layered double hydroxides intercalated with carbonate, nitrate, chloride and sulfate ions: Synthesis, characterization and dyes removal properties. Journal of Taibah University of Science, 11, 90–100. 23. Wan, D., Liu, H., Liu, R., Qu, J., Li, S., & Zhang, J. (2012). Adsorption of nitrate and nitrite from aqueous solution onto calcined (Mg/Al) hydrotalcite of different Mg/Al ratio. Chemical Engineering Journal, 195–196, 241–247. 24. Kloprogge, J. T., Wharton, D., Hickey, L., & Frost, R. L. (2002). Infrared and Raman study of interlayer anions CO32-, NO3-, SO42- and ClO4- in Mg/Al hydrotalcite. American Mineralogist, 87, 623–629. 25. Wojnicki, M., Rudnik, E., Luty-Błocho, M., Pacławski, K., & Fitzner, K. (2012). Kinetic studies of gold(III) chloride complex reduction and solid phase precipitation in acidic aqueous system using dimethylamine borane as reducing agent. Hydrometallurgy, 127–128, 43–53. 26. Han, S., Hou, W., Zhang, C., Sun, D., Huang, X., & Wang, G. (1998). Structure and the point of zero charge of magnesium aluminum hydroxide. Journal of the Chemical Society, Faraday Transactions, 94, 915–918. 27. Petcharoen, K., & Sirivat, A. (2012). Synthesis and characterization of magnetite nanoparticles via the chemical co-precipitation method. Materials Science and Engineering: B, 177, 421–427. 28. Lagergren, S. (1898). Zur theorie der sogenannten adsorption geloster Stoffe. Kungliga Svenska Vetenskapsakad Handlingar, 24, 1–39. 29. Ho, Y. S., & McKay, D. (1999). Pseudo-second order model for sorption processes. Process Biochemistry, 34, 451–465. 30. Santosa, S. J. (2014). Sorption kinetics of Cd(II) species on humic acid-based sorbent. Clean: Soil, Air, Water, 42, 760–766. 31. Yang, Y., Gao, N., Chu, W., Zhang, Y., & Ma, Y. (2012). Adsorption of perchlorate from aqueous solution by the calcination product of Mg(Al;-Fe)

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Heavy Metal Removal from Wastewater Using Different Cheap Adsorbents: Olive Cake, Moringa, Eucalyptus, and Pine Cone Mohammed Matouq, Zaid Al-Anbar, Mohammed Al-Anber, and Omar Al-Ayed

Abstract

The performance of four different biosorption materials for the removal of selected heavy metals from wastewater is introduced in this chapter. Four types of low-cost adsorbents Moringa, Eucalyptus, Pine cones, and Olive cake were examined on synthetic wastewater under different experimental conditions. Promising results were obtained by using these materials for the removal of heavy metal ions such as Nickel (Ni), Chromium (Cr), Copper (Cu), and Cadmium (Cd) that were fortified in typical wastewater concentrations. The removal efficiency at a Cd ion concentration of 100 ppm reached 66% at pH 6 and 28 °C using an olive pomace solid. The adsorption capacity decreased from 65.4 to 44.4 mg/g, when the temperature was increased from 28 to 45 °C. Grinded Moringa pods were tested for a mixture of all three heavy metal ions (Ni, Cu, and Cr) at concentrations between 25 and 100 ppm. An increase

M. Matouq (&)
Z. Al-Anbar
O. Al-Ayed Faculty of Engineering Technology, Chemical Engineering Department Amman, Al-Balqa Applied University, Amman, Jordan e-mail: [email protected] M. Al-Anber Department of Chemical Science, Faculty of Science, Industrial Inorganic Chemistry, Mu´Tah University, Mu´Tah, Jordan

in the Moringa dose resulted in a negative effect on removal efficiency. It was found to be lower than those of single metal solutions under the same conditions. The kinetic tests for three different adsorbents Eucalyptus, Moringa pods, and Pine cones demonstrated that the biosorption equilibrium is reached within 40 min for Cu, 30 min for Ni, and 40 min for Cr ions. The removal of Ni and Cr with Eucalyptus bark and Pine cones was tested at heavy metal concentrations of 400, 600, and 900 ppm. The removal efficiencies for Eucalyptus bark were 63–98% for Cr and 44–55% for Ni. Using Pine cones, removal efficiencies were 43–52% for Cr and 72–80% for Ni. In the case of all four biosorbents, the adsorption behavior of the tested heavy metal ions could be well described using the Langmuir and Freundlich isotherm models. The adsorption kinetics showed good agreement with first- or second-order kinetics depending on the ions and biosorbents used. It could be shown that with all tested biosorbents and heavy metal ions, removal from the wastewater up to 99% was possible. Keywords











Olive cake Eucalyptus bark Moringa pods Pine cones Isotherm models Heavy metal removal Biosorbents Wastewater










© The Author(s), under exclusive license to Springer Nature Switzerland AG 2023 E. Debik et al. (eds.), Wastewater Management and Technologies, Water and Wastewater Management, https://doi.org/10.1007/978-3-031-36298-9_11

153

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M. Matouq et al.

qt T t V

Nomenclature

bo bt Ce Cf Cp Co Ce Ci

Es K Kf KL

1/n qe t K

K1 K2 Kf Kid KT Lp m qe qe,m qm

Langmuir isotherm constant (L/mg) Constant related to the heat of sorption (J/mol) Equilibrium concentration of metal ions in solution (mg/L) Concentration of metal ion in the feed solution (mg/L) Concentration of metal ion in the permeate solution (mg/L) Initial concentration of metal ions in solution (mg/L) Final concentration of heavy metal in the solution sample at equilibrium (ppm) Initial concentration of heavy metal in the solution sample before treatment (ppm) Adsorption energy (KJ/mol) The rate constant of adsorption The Freundlich constant which related to the bonding energy (L/mg), Langmuir equilibrium constant which is related to the adsorption/desorption energy, the affinity of the binding sites. The heterogeneity factor and n (g/L) The mass of metal adsorbed at equilibrium (mg of ion/g of metal), The mixing time in minutes. Dubinin–Radushkevich constant related to mean free energy of adsorption Pseudo first-order rate for adsorption (1/min) Pseudo second-order rate constant for adsorption (g/mg. min) Freundlich constant related to the bonding energy (mg/g) (L/g)1/n Intraparticle diffusion rate constant (mg/g. min0.5) Temkin isotherm constant (L/g) Effective pore length (m) Dry weight adsorbent (g) Sorption capacity at equilibrium (mg/g) Measured adsorbate concentration at equilibrium (mg/g) Maximum adsorption capacity (mg/g)

1

Sorption capacity at time t (mg/g) Temperature (C) Time (min) Volume of aqueous solution (L)

Introduction

Using olive cake in different aspects of chemical processes has recently gained wide attention because of the favorable environmental effects of such materials [1–4]. Although olive cake itself is harmless initially, phenolic ingredients, such as oleuropein, hydroxytyrosol, and tyrosol, must be seen as crucial components from the ecological point of view [5]. The application of biomaterials as adsorbents aims at replacing expensive conventional activated carbon [6]. Activated carbon [7–10] is widely used as an effective adsorbent for the removal of Cd(II) ions from aqueous solutions. However, the costs for the production and preparation of activated carbon are still high. In view of this, it is more than justified to push forward research that focuses on cost-effective sorbents such as chitin [11], anaerobic sludge [12], apple residues [13], sawdust [14], rice plaster [15], clay [16], zeolite [17], chitosan [18], tea waste [19, 20], and algae [20, 21]. Historically, the evergreen characteristics of Pine trees have benefitted society by many aspects such as source of energy and their evergreen [22]. Recent studies on the use of Eucalyptus residues such as leaves, bark, and wood have identified them as of value as adsorbent materials [23]. Moringa trees in Jordan have also been cultivated nearby the Dead Sea, since warm weather is necessary to cultivate them. Several studies have tested Moringa pods, leaves, and barks as cheap adsorbents for wastewater treatment [24–28]. The present study introduces three plants that exist locally in Jordan and are cheap, namely Eucalyptus bark, Pine cones, and Moringa pods as well as Olive cake. The chapter will discuss

Heavy Metal Removal from Wastewater Using Different Cheap …

the removal efficiencies and comparison of different biosorbents used according to the adsorption capacity. It also gives examples of kinetic studies related to the adsorption process.

2

155

distilled water for two days by mixing with a magnetic stirrer until the water became clear. The clean particles were dried in an electrical oven overnight at 80 °C, for 12 h, and the dried particles were sieved to mesh sizes (1 mesh, 2 mesh, and 2.8 mesh) [29].

Material and Methods

2.1 Preparation of the Biosorbent Materials (a) Olive cake The sample used in this study for the olive cake was obtained from an olive mill company and kept in an open environment for natural drying under direct sun for around a year. The sample was then washed and sieved to remove sand and other impurities. The olive cake sample was first weighted and washed with hot deionized water, and when it was cooled, it was washed again with double distilled water. The sample was placed in an electrical oven at 100 °C until weight constancy. The sample was ground and sieved, and only a fraction of 0.1 mm size was used in this experiment. (b) Moringa pods The collected Moringa pods were crushed and then mixed in distilled water for ten hours in a glass beaker attached to a magnetic stirrer. The mixer was used to assure complete removal of dirt and color extracted from the pods. Water was removed and changed several times until it was clear. The cleaned sample was then dried in an electrical oven at 105 ± 2 °C overnight. The dried samples were sieved to 1 mesh, 2 mesh, and 2.8 mesh using an electric vibrating sieve.

2.2 Chemical Preparations Different heavy metals, Cr, Ni, and Cd, were used to test the ability of adsorption by these biosorbents and the concentration determined by their existence in real wastewater. The prepared solutions were assumed to simulate wastewater with each tested for certain metal ions. (a) Ni and Cr ıons’ preparation Two solutions for both ions were prepared by dissolving a specified amount of NiCl2.6H2O, and K2Cr2O3 in deionized water to get different concentrations of 400, 600, and 900 ppm. (b) Cd ions’ preparation A solution with 100 ppm of Cd was prepared by dissolving cadmium acetate in deionized water, and the required pH was adjusted using 1.0 N of sodium hydroxide and hydraulic acid. (c) Cu ions’ preparation A specified amount of CuSO4.5H2O was dissolved in deionized water to prepare a stock solution of 1000 ppm that can be re-diluted again to get the required concentration in each experimental work.

2.3 Kinetics and Isotherm Models (c) Eucalyptus barks and Pine cones Pine cones and Eucalyptus barks were collected from trees inside the university campus during the dry summer. The collected samples were crushed and sieved to get the required size for the experimental work. Particles were soaked in

To study the kinetics and removal capacity with different ion concentrations, the removal efficiency was calculated according to Eq. (1). Removal% ¼

ðCi  Ce Þ ; Ci

ð1Þ

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where Ci is the initial concentration of heavy metal ions in the solution at the initial time (ppm), Ce is the final concentration of heavy metal ions at equilibrium (ppm). (a) Isotherm models The isotherm is characteristic of a specific system at a particular temperature [30]. The term is related to the amount of solute adsorbed onto the solid and the equilibrium concentration of the solute in the solution at a certain temperature. The quantity of adsorptions can be calculated by Eq. (2): qe ¼

Co  Ce : m=V

ð2Þ

This equation can also give a good indication of the type of adsorption that occurs multilayer chemical or physical adsorption.

2.4 Langmuir Isotherm Model The Langmuir isotherm model describes the adsorption as a monolayer of the surface with several adsorption sites. The model indicates that when the site is filled, there will be no further adsorption to take place within the same site. This is due to the fact t surface will reach a saturation point where the maximum adsorption of the surface will be achieved. This model can be represented by Eqs. 3 and 4 [31]: qe ¼

qm b0 C e : 1 þ b0 Ce

ð3Þ

This equation can be re-arranged in a linear form to describe the Langmuir isotherm as: Ce 1 1 ¼ þ Ce : qe b0 qm qm

2.5 Freundlich Isotherm Model The Freundlich model is usually used to describe the adsorption energy, in which the energy exponentially decreases until the adsorption is completed [31]. The isotherm model can be described by the empirical Eq. 5: qe ¼ Kf Ce1=n ;

ð5Þ

where the Freundlich equilibrium constants can be calculated from the linear plot of ln qe and ln Ce as the following Eq. 6 indicates: ln qe ¼ ln Kf

1 lnCe : n

ð6Þ

The slope of 1/n usually lies between 0 and 1 and gives a measure of the adsorption intensity or surface heterogeneity. When the value of the slope reaches zero, the more heterogeneous process is dominant, whereas a value less than one indicates a chemisorption process (adsorption is a chemical process) when greater than one a cooperativeorption is dominant, which is a physical process [27].

2.6 Adsorption Kinetics Models The nature of the sorption process in the kinetic study usually depends on the physicochemical characterization of the adsorbent and system conditions, such as the temperature [27, 29, 31]. The quantity of metal ions adsorbed defined by qt, where t is the time, can be calculated from the following Eq. 7: qt ¼

C0  Ce : m=V

ð7Þ

ð4Þ

Values of Langmuir parameters qm and b0 are calculated from the slope and intercept of the linear plot Ce =qe versus Ce .

2.7 Adsorption Mechanism E kinetics for adsorption in porous adsorbents is controlled by three steps:

Heavy Metal Removal from Wastewater Using Different Cheap …

Step 1: Metal ions are transported from the bulk solution to the film boundary on the surface of the sorbent. Step 2: Mass transfer of the ions from the surface of the adsorbent to the intraparticle active sites. Step 3: The ions are then adsorbed on the exterior surface. The model can be mathematically defined as follows (Eq. 8): qt ¼ Kid t0:5 ;

ð8Þ

where the Kid can be calculated from the slope of plotting qt versus t0:5 . According to this model, two or more steps influence the adsorption process, and the plot will represent a multi-linear behavior. The first curve will relate to the boundary layer, while the second one will describe the gradual adsorption stage and the final one represents intraparticle diffusion [27, 29, 31].

2.8 Nonlinear Chi-Square Analysis ( v2 ) To test the good fit of the experimental data with the best fit for the isotherm models, a nonlinear Chi-square is used as an appropriate statistical tool to evaluate its fit. The Chi equation is expressed as Eq. 9 [27]: v2 ¼

3

 2 n X qe  qe;m : qe;m i¼1

ð9Þ

Results and Discussions

3.1 Removal of Cd by Olive Cake Adsorptions The previous study for the removal of Cd(II) showed a dependent relationship on the sorbents concentration (respective doses) [31]. The study used different amounts of olive cake as sorbents with 0.05, 0.1, 0.2, and 0.3 g with the change of the solution temperatures at 28, 35, and 45 °C. In the study, the pH was adjusted to 6. In this part of

157

Table 1 Removal efficiency of Cd(II) ions (100 ppm, at pH 6 and retention time: 24 h) with different doses of the olive cake [32] Olive cake (g)

Removal (%) 28 °C

35 °C

45 °C

0.05

12

10

8

0.10

28

23

18

0.20

52

28

32

0.30

68

62

50

the work, the Cd(II) concentration was set at 100 ppm when the olive cake was used. Table 1 summarizes the results. The removal efficiency can be seen as dependent on the amount of olive cake. Lower removal will be obtained, when the temperature increases from 28 to 45 °C as well as when the amount of cake is increased. This behavior is expected because when a higher amount of olive cake is used in the solution, a higher chance of exchanging sites will be available for the ions. The current results show that the best removal efficiency is obtained, when the olive cake is used at a temperature of 28 °C and 0.3 g. Adsorption preferably takes place at a lower temperature. This can be explained by the fact that when the temperature is increased, the ions bound to the surface lose their stability at the surface and escape into the main solution. Figure 1 shows the experimental data fitted by the linear plotting based on the Langmuir model, using 28, 35, and 45 °C temperatures. The sorption capacity is 65.36 mg/g, which is the maximum adsorption capacity (qmax), when the temperature is 28 °C, which can indicate also a complete monolayer coverage. The removal capacity decreases from 65.36 to 44.44 mg/g, when the temperature increases from 28 to 45 °C. The adsorption coefficient b in the Langmuir model, which is related to the adsorption energy, has decreased from 0.0710 to 0.0628 L/mg when the temperatures have increased from 28 to 45 ° C. The decrease in qmax and b values with temperature changes indicates that the cadmium ions are more favorable at lower temperatures when the olive cake is used. This also indicates that the adsorption process is exothermic (Table 2).

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Fig. 1 Langmuir adsorption isotherms fitting for Cd(II) ions 100 ppm [31]

3.2 Removal of Cu, Ni, and Cr by Moringa Pods This section presents the use of Moringa pods as another cheap adsorbent and uses different heavy metals than in the olive cake experiment with a different initial concentration. Moringa pods were used to examine the adsorption of copper, nickel, and chromium at different concentrations

Table 2 Langmuir and the Freundlich constants for Cd (II) ions adsorption on the olive cake [31]

Table 3 Using Moringa pods as adsorbents and models’ constants for Cu ions at temperatures 20, 40, and 70 °C, 1 g Moringa dose [27]

Temperature (°C)

and temperatures. The wastewater used in the study was again a simulated one. Tables 3, 4, and 5 show that the adsorption parameters and their values with the fitting to linear lines are represented by R2. The tables also demonstrate that the adsorption capacity decreases from 6.075 to 5.438 and 5.543 to 4.771 mg/g for copper and nickel, respectively, when the temperature has raised from 20 to 70 °C. This indicates that the adsorption process for Moringa pods is also an exothermic one. Moreover, Table 3 indicates an increase from 3.191 to 5.497 mg/g for Cr ions with the increase of the temperature from 20 to 70 °C which indicates an endothermic process. The Freundlich isotherm constant (n) at different temperatures is higher than one and it is between 1.94 and 3.624, indicating a good adsorption process for Moringa pods. The coefficient of determination (R2) value and Chi-square value (v2) was used to find out the best fit among the adsorption isotherm models. The large v2 value represents the variation from the experimental data. Among the three selected ions, the isotherm models were fitted with Langmuir

Langmuir constants

1/n

Kf

R2

0.987

0.237

19.9

0.793

60.6

0.989

0.228

19.1

0.815

44.4

0.991

0.195

15.7

0.875

b (L/mg)

qmax (mg/g)

28

0.071

65.4

35

0.070

45

0.063

Adsorption isotherm model

Freundlich constants R

2

Temperature (°C) 20

40

70

qm

6.075

5.935

5.438

b0

0.1989

0.1074

0.9487

R2

0.8545

0.7975

0.7061

v

0.6803

0.5890

0.7610

n

3.625

3.158

2.84

Kf

1.173

0.9336

0.9202

2

R

0.9227

0.9601

0.9071

v2

0.4095

0.2973

0.5840

Langmuir isotherm model

2

Freundlich isotherm model

Heavy Metal Removal from Wastewater Using Different Cheap … Table 4 Using Moringa pods as adsorbents and models constants for Ni ions at temperatures 20, 40, and 70 °C, 1 g Moringa dose [27]

Adsorption isotherm model

159

Temperature (°C) 20

40

70

qm

5.543

4.919

4.771

b0

0.1251

0.1359

0.0973

R2

0.9076

0.9184

0.8361

v

0.4213

0.4464

0.5789

n

2.334

2.557

2.711

Kf

0.9463

0.9606

0.9170

2

R

0.9223

0.8998

0.833

v2

0.3588

0.8998

0.6010

Adsorption isotherm model

Temperature (°C)

Langmuir isotherm model

2

Freundlich isotherm model

Table 5 Using Moringa pods as adsorbents and models constants for Cr ions at temperatures 20, 40, and 70 °C with 1 g Moringa dose [27]

20

40

70

qm

3.191

4.873

5.497

b0

0.6555

0.5857

0.3822

2

R

0.9775

0.9881

0.9744

v2

1.0586

0.2775

0.6186

n

3.624

3.158

2.842

Kf

1.390

1.512

1.508

R2

0.7004

0.7857

0.8086

v

1.0258

0.5402

0.6186

Langmuir isotherm model

Freundlich isotherm model

2

and Freundlich ones. For copper ions are the best fit by the Freundlich model at 40 °C with 96%. For Ni, both Langmuir and Freundlich gave a good correlation. The Langmuir isotherm provided a better correlation for chromium. In the previous section, the removal efficiency has been tested for a single ion placed in the simulated wastewater. However, in some cases, all ions as pollutants may exist together in the wastewater. For this reason, the study of mixing all ions (Cu, Ni, and Cr) together and placing them in the simulated wastewater has been examined to study the biosorbents’ behavior. Two different concentrations of 25 and 50 ppm

for as mixed ions with different Moringa doses (0.25, 0.5, 0.75, and 1 g) have been conducted. The results are shown in Table 6. This experiment was introduced to check the ability of the adsorbents to remove ions when there is more than one ion to remove. Table 6 indicates clearly that the removal efficiency declined when a mixture of ions was introduced. This can be explained by the active sites available during the adsorptions that are limited when there is more than one sort of ions present and finding a new place will be very competitive based on the sorption surface and ion radius size.

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M. Matouq et al.

Table. 6 Removal efficiency comparison when copper, nickel, and chromium are used in a single concentration and a mixture in a solution at a temperature 20 °C [27] Metal concentration

25 ppm

Metal ion Copper Nickel Chromium

50 ppm

Adsorbent dose of Moringa (g)

Adsorbent dose of Moringa (g)

0.25

0.25

0.5

0.75

1

0.5

0.75

1

Single (Cu)

61.6

68.3

80.5

90.3

41.7

51.4

57.9

65.2

Multi (Cu, Ni, Cr)

7.9

18.7

33.0

41.4

9.7

10.3

14.1

18.4

Single (Ni)

49.5

58.4

63.5

68.5

31.9

40.6

59.1

73.2

Multi (Cu, Ni, Cr)

4.7

9.8

12.416

16.9

9.4

11.7

13

15.3

Single (Cr)

23.3

45.1

66.7

91.3

9.1

39.8

51

67

Multi (Cu, Ni, Cr)

7.06

15.5

19.5

28.8

10.65

12.9

17.7

21.5

3.3 Adsorption of Ni and Cr by Pine Cones and Eucalyptus Bark In this section of the study, other types of adsorbents like Pine cones and Eucalyptus bark were tested according to the adsorption capacity of two different ions, Ni and Cr. The experimental work was also conducted on simulated wastewater that contains the expected concentrations of ions. The removal efficiency of metal ions is given in Table 7. Three different concentrations of Cr and Ni (400, 600, and 900 ppm) have been introduced in this part, and the removal capacity at different sorbents varied from 43 to 98%. The results indicate that Ni ions showed the lowest removal when Eucalyptus bark was used (44–55%). And when Pine cones were used, Cr ions gave the lowest removal efficiency (43– 52%). The equilibrium distribution of metal ions

Table. 7 Cr and Ni ions with the removal efficiency using Eucalyptus bark and Pine cones [29]

Initial concentration (ppm) Cr

Ni

between the adsorbents and the solution is important to determine the maximum adsorption capacity. Several isothermal models are available to describe equilibrium adsorption distribution. The models, Langmuir and Freundlich, were used to fit the experimental data, and the results are given in Table 8.

3.4 Kinetics Study for Removal of Ni and Cr by Moringa, Eucalyptus and Pine Figures 2 and 3 show an example of kinetics for the removal of ions using different biosorbents and how it fits the experimental work. The results indicate that Ni can have the best fit to the first order when Eucalyptus and Pines are used, while for Cr, the second order is the best for all three introduced adsorbents.

Eucalyptus bark (%)

Pine cones (%)

400

98

43

600

70

84

900

63

52

400

55

80

600

50

72

900

44

72

Heavy Metal Removal from Wastewater Using Different Cheap …

161

Table 8 Determined values of models fitting [29] Constants Langmuir

Freundlich

Ni

0.11

0.01

0.03

0.31

0.08

0.006

0.02

0.025

2

R

0.97

0.98

0.94

0.99

Kf

0.79

0.001

0.44

0.05

n

11.86

1.3

7

2.29

R2

0.83

0.82

0.59

0.99

Cr, Eucalyptus

Time (min)

0

60

90

-3

Ni, Moringa

-0.5

6

9

12

-1 ln(qe-qt)

y = -0.567x + 0.409 R² = 0.9014

-3

3

30

60

90

y = -0.448x - 0.2 R² = 0.9883

9

12

-3

0 120

6

-2

y = -1.195x + 0.415 R² = 0.9587

Cr, Pine

Time (min)

-1 -2

Time (min)

-5

150 ln(qe-qt)

ln (qe -qt)

0

120

y = -1.149x + 0.38 R² = 0.9144

-4

-3 Ni, Pine

90

-3

0

15

-1.5 -2.5

60

-2

Cr, Moringa

-1 -2

Time (min)

-5

Time (min) 3

30

-4

y = -1.625x + 0.2 R² = 0.9937

-5

0

-1 ln(qe-qt)

ln(qe-qt)

Cr

-2

-4

ln(qe-qt)

Ni

KL

30

-1

Eucalyptus bark

Cr aL

Ni, Eucalyptus

0

Pine cone

Time (min) 30

60

90

-2 -4

y = -0.994x - 0.185 R² = 0.8781

-6

Fig. 2 First-order kinetics fitting for three adsorbents using Cr and Ni ions as pollutants

120

162

M. Matouq et al. Ni, Eucalyptus

200 150

150 y = 29.335x + 87.697 R² = 0.8188

100

qt/t

qt/t

Cr, Eucalyptus

200

100

50

y = 29.754x + 40.805 R² = 0.9753

50

0

30

60

0

90

30

Ni, Moringa

20

qt/t

qt/t

8

2 0

3

6

9 12 Time (min)

15

0

5

150

Ni, Pine

250

y = 0.724x + 1.495 R² = 0.9958

6 4

y = 1.801x + 6.729 R² = 0.8844

5

200

Time (min)

10

15

Cr, Pine

100

150

y = 13.302x + 125.32 R² = 0.879

100

qt/t

qt/t

120

10

10

y = 29.548x + 15.67 R² = 0.997

50

50 0

90

Cr, Moringa

12

15

0

60

Time (min)

Time (min)

30

60

90 120 Time (min)

150

0 30

60 90 Time (min)

120

Fig. 3 Second-order kinetics fitting for three adsorbents using Cr and Ni ions as pollutants [29]

4

Conclusion

The four biosorbents, namely Moringa pods, olive cake, Pines cone, and Eucalyptus bark presented here, have suggested a promising potential to remove heavy metals from wastewater regardless of the concentration used in the study. The kinetics study shows that the biosorbents fit either the first- or second-order kinetics

models depending on the ions used in the experiments. The Langmuir and Freundlich models show a good fit for the adsorption. The results show that for all adsorbents used in the study, there is a good relationship between the adsorbent concentrations and the removal of ions. Higher removal efficiency is obtained when higher adsorbent concentrations are added to the solution. The Moringa pods show another behavior, namely that a decrease occurs in

Heavy Metal Removal from Wastewater Using Different Cheap …

removal efficiency when the concentrations are increased. The removal efficiency is better when single ions are used separately and not mixed. This conclusion may be attributed to the competition between mixed ions to be adsorbed on the surface and the available active sites as well as the radius of these ions. Using olive cake to remove Cd ions obtained a good removal efficiency of up to 98%. The experimental data are in good fit with the proposed isotherm models, such as Langmuir and Freundlich models. When Pine cones and Eucalyptus bark are used to remove Ni and Cr ions, there is a good removal from the solution with a higher initial concentration compared with olive cake and Moringa pods. If Moringa pods and olive cake are compared for their removal efficiencies, Moringa pods still give the highest efficiency. The Eucalyptus bark removes Ni and Cr with the lowest efficiency of 44–55%, while for Pine cones the efficiency is 43–52%, and for Moringa pods, the efficiency is 90–99%. When using Moringa pods, Pine and Eucalyptus bark, better correlations are obtained using the Langmuir model with correlation coefficients R2 of over 90%. The fitting gives a KL constant 0 < KL < 1 and aL 0 < aL < 1. The correlation coefficient (R2 = 0.94–0.99) for the adsorption of Cr and Ni shows that the isotherm best fits the Langmuir isotherm. In conclusion, all study results presented in this review demonstrate that the choice of cheap biosorbents is promising and can compete with other, more expensive adsorbent alternatives.

References 1. Al-Shaweesh, M., Mohammed, M., Al-Kabariti, D., Khamash, D., Al-Zawaidah, S., Hindiyeh, M., & Omar, W. (2018). Olive mill wastewater (OMW) treatment by using ferric oxide dephenolization and chemical oxygen demand removal. Global NEST Journal, 20(3), 558–563. 2. Al-Hmoud, L., Al-Saida, B., & Sandouqa, A. (2020). Olive mill wastewater treatment: A recent review. Jordanian Journal of Engineering and Chemical Industries (JJECI), 3(3), 91–106.

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3. Rawajfeh, K., Sandouqa, A., & Al-Rawajfeh, K. (2021). Precipitation of solid waste in olive mill wastewater by coagulation using calcium carbonate (CaCO3). Jordanian Journal of Engineering and Chemical Industries (JJECI), 4(3), 78–86. 4. Gueboudji, Z., Kadi, K., & Nagaz, K. (2022). Antiinflammatory activity of polyphenols from olive oil mill wastewaters. Jordanian Journal of Engineering and Chemical Industries, 5(1), 18–23. 5. Foti, P., Pino, A., Romeo, F. V., Vaccalluzzo, A., Caggia, C., & Randazzo, C. L. (2022). Olive pomace and pâté olive cake as suitable ingredients for food and feed. Microorganisms, 10, 237. 6. Babel, S., & Kurniawan, T. A. (2003). Low-cost adsorbents for heavy metals uptake from contaminated water: A review. Journal of Hazardous Materials, 97(1–3), 219–243. 7. Tilaki, D., & Ali, R. (2003). Study on removal of cadmium from water environment by adsorption on GAC, BAC and Biofilter. Ucd Dublin, 8, 35–39. 8. Marzal, P., Seco, A., Gabaldon, C., & Ferrer, J. (1996). Cadmium and zinc adsorption onto activated carbon influence of temperature pH and metal/carbon ratio. Journal of Chemical Technology and Biotechnology, 66(3), 279–285. 9. Seco, A., Marzal, P., Gabaldon, C., & Ferrer, J. (1997). Adsorption of heavy metals from aqueous solutions onto activated carbon in single Cu and Ni systems and in binary Cu–Ni, Cu–Cd and Cu–Zn systems. Journal of Chemical Technology and Biotechnology, 68(1), 23–30. 10. Madhava Rao, M., Ramesh, A., Purna Chandra Rao, G., & Seshaiah, K. (2006). Removal of copper and cadmium from the aqueous solutions by activated carbon derived from Ceiba pentandra hulls. Journal of Hazardous Materials, 129(1–3), 123–129. 11. Leyva-Ramos, R., Rangel-Mendez, J. R., MendozaBarron, J., Fuentes-Rubio, L., & Guerrero-Coronado, R. M. (1997). Adsorption of cadmium(II) from aqueous solution onto activated carbon. Water Science and Technology, 35(7), 205–211. 12. Benguella, B., & Benaissa, H. (2002). Cadmium removal from aqueous solutions by chitin: Kinetic and equilibrium studies. Water Research, 36(10), 2463–2474. 13. Ulmanu, M., Maranon, E., Fernandez, Y., Castrillon, L., Anger, I., & Dumitriu, D. (2003). Removal of copper and cadmium ions from diluted aqueous solutions by low cost and waste material adsorbents. Water, Air, and Soil Pollution, 142(1–4), 357–373. 14. Lee, S. H., Jung, C. H., Chung, H., Lee, M. Y., & Yang, J. (1998). Removal of heavy metals from aqueous solution by apple residues. Process Biochemistry, 33, 205–211. 15. Shukla, S. S., Yu, L. J., Dorris, K. L., & Shukla, A. (2005). Removal of nickel from aqueous solutions by sawdust. Journal of Hazardous Materials, 121, 243–246. 16. Singh, K. K., Rastogi, R., & Hasan, S. H. (2005). Removal of cadmium from wastewater using

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18.

19.

20.

21.

22.

23.

24. 25.

M. Matouq et al. agricultural waste ‘rice polish.’ Journal of Hazardous Materials, 121, 51–58. Farrah, H., & Pickering, W. F. (1977). The sorption of lead and cadmium species by clay minerals. Australian Journal of Chemistry, 30, 1417–1422. El-Kamash, A. M., Zaki, A. A., & Abed El Geleel, M. (2005). Modeling batch kinetics and thermodynamics of zinc and cadmium ions removal from waste solutions using synthetic zeolite A. Journal of Hazardous Materials, 127, 211–220. Evans, J. R., Davids, W. G., MacRae, J. D., & Amirbahman, A. (2002). Kinetics of cadmium uptake by chitosan-based crab shells. Water Research, 36, 3219–3226. Orhan, Y., & Buyukgungor, H. (1993). The removal of heavy metals by using agricultural wastes. Water Science and Technology, 28, 247–255. Da Costa, A. C. A., & De França, F. P. (1996). Cadmium uptake by biosorbent seaweeds: Adsorption isotherms and some process conditions. Separation Science and Technology, 31, 2373–2393. Al-Widyan, M. I., Al-Jalil, H. F., Abu-Zreig, M., & Abu-Hamdeh, N. H. (2002). Physical durability and stability of olive cake briquettes. Canadian Biosystems Engineering, 44(3), 41–45. Davidson, K., DiSalvo, A., Fukuda, J., Gersbach, J., Grotbo, J., & Ramse, J. (2016). Street tree inventory report: Hillsdale neighborhood. Portland Parks & Recreation. Sellers, C. H. (2018). Eucalyptus: Its history, growth, and utilization. Fb & c Limited. Aravind, J., Muthusamy, S., Sunderraj, S. H., Chandran, L., & Palanisamy, K. (2013). Pigeon pea (Cajanus cajan) pod as a novel eco-friendly biosorbent: A study on equilibrium and kinetics of Ni

26.

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31.

32.

(II) biosorption. International Journal of Industrial Chemistry, 4, 1–9. Vieira, A. M. S., Vieira, M. F., Silva, G. F., Araujo, A. A., Fagundes-Klen, M. R., Veit, M. T., & Bergamasco, R. (2009). Use of Moringa oleifera seed as a natural adsorbent for wastewater treatment. Water Air Soil Pollutant, 206, 273–281. Kalavathy, M. H., & Miranda, L. R. (2010). Moringa oleifera—A solid phase extracted for the removal of copper, nickel and zinc from aqeous solutions. Chemical Engineering Journal, 158, 188–199. Matouq, M., Jildeh, N., Qtaishat, M., Hindiyeh, M., & Al Syouf, M. (2015). The adsorption kinetics and modeling for heavy metals removal from wastewater by Moringa pods. Journal of Environmental Chemical Engineering, 3(2), 775–784. Matouq, M., Moatasem, S., Al-Ayed, O., El-Hasan, T., Yamada, H., & Tagawa, T. (2021). Biosorption of chromium and nickel from aqueous solution using pine cones, eucalyptus bark, and moringa pods: A comparative study. Water Practice and Technology, 16(1), 72–82. Harikishore Kumar Reddy, D., Seshaiah, K., Reddy, A. V. R., Madhava Rao, M., & Wang, M. C. (2010). Biosorption of Pb+2 from aqueous solutions by Moringa oleifera bark: Equilibrium and kinetics studies. Journal of Hazardous Materials, 174, 831– 838. Inglezakis, V. J., & Poulopoulos, S. G. (2006). Adsorption, ion exchange and catalysis: Design of operations and environmental applications. Elsevier. Al-Anber, Z., & Matouq, M. A. (2008). Batch adsorption of cadmium ions from aqueous solution by means of olive cake. Journal of Hazardous Materials, 151, 194–201.

Treatment Technologies in Developing Countries

Leachate Treatment in Brazil and Potential Technologies: A General Approach Mariana Islongo Canabarro, Siara Silvestri, Victor Alcaraz Gonzalez, and Elvis Carissimi

Brazil and several other countries, and future trends in this behalf.

Abstract

The global generation of municipal solid waste has been increasing every year, following population’s consumption profile. Currently, the most common and suitable way to dispose of this waste is landfilling. However, landfills are sources of two products considered highly polluting: landfill leachate and dump gas, both arising from the decomposition of solid waste in the landfill cells. Landfill leachate represents a high risk to the environment and may result in contamination of groundwater and/or surface water, leading to eutrophication of water bodies, if disposed of improperly, i.e., without prior treatment. With the development of technology over the years, several treatment techniques have emerged intending to improve the efficiency of the treatment and to adjust the leachate to the release standards for its subsequent disposal. Thus, this review addresses the generation and characteristics of landfill leachate, the techniques that can be applied in its treatment, and a general approach to leachate treatment in

M. I. Canabarro
S. Silvestri
E. Carissimi (&) Graduate Program in Civil Engineering, Federal University of Santa Maria, Santa Maria, Brazil e-mail: [email protected] V. A. Gonzalez Universidad de Guadalajara (UdeG), Guadalajara, México

Keywords

Landfill leachate technologies


Treatment
Treatment

Abbreviations

BOD COD AOPs MSW WTP CONAMA

1

Biological oxygen demand Chemical oxygen demand Advanced oxidation processes Municipal solid waste Water treatment plant Conselho Nacional do Meio Ambiente (National Environment Council)

Introduction

1.1 Generation and Characteristics of Landfill Leachates Landfill leachate refers to the liquid wastewater infiltrated by rainwater through waste disposed of in landfills, as well as the moisture present in the waste and its degradation products [1]. The amount of leachate produced is directly related to factors such as precipitation, evaporation, surface runoff, infiltration, groundwater intrusion into

© The Author(s), under exclusive license to Springer Nature Switzerland AG 2023 E. Debik et al. (eds.), Wastewater Management and Technologies, Water and Wastewater Management, https://doi.org/10.1007/978-3-031-36298-9_12

167

168

landfills, and the degree of compaction of garbage [2, 3]. The composition of landfill leachate can be very different and heterogeneous. However, it generally has the following components: high concentrations of dissolved organics (volatile fatty acids and refractory organic compounds such as humic and fulvic acids); inorganic macrocomponents (potassium, ammonium nitrogen, calcium, sodium, magnesium, iron, manganese, sulfates, chlorides, and hydrocarbons); heavy metals (nickel, lead, chromium, copper, cadmium, and zinc); and xenobiotic organic compounds (pesticides, phenols, and aromatics) [4, 5]. According to Kjeldsen et al. [6], the composition of the leachate varies widely depending on several factors including the composition of the incoming waste, climatic conditions, and the age of the landfill. The age of the landfill has a significant impact on the composition of the leachate, as some characteristic parameters also change as the landfill enters a stable phase. In general, the stabilization phases are (a) initial aerobic phase, (b) anaerobic acid phase, (c) initial methanogenic phase, and (d) stable methanogenic phase [7, 8]. According to Christensen et al. (2001) [8], during the initial phase of the landfill, known as the acid phase, the leachate has an acidic pH and high concentrations of easily biodegradable organic matter and volatile acids. For mature landfills that are in the methanogenic phase, there is a significant production of methane, the pH of the leachate is alkaline, and the organic material is present in the humic and fulvic fractions. Landfills with less than 5 years are considered new and those with more than 10 years are considered stabilized [7, 9]. In addition, the characteristics of landfill leachate are also evaluated through physical– chemical parameters. The most commonly used parameters are ammonia nitrogen and total nitrogen, biochemical oxygen demand (BOD), pH, chemical oxygen demand (COD), phosphorus, chloride, alkalinity, series of solids, and heavy metals [6, 10].

M. I. Canabarro et al.

1.2 Characteristics of Brazilian Landfill Leachates The degradation of landfilled waste involves different chemical, physical, and biological processes, which constantly affect the composition of the leachate generated, influencing its characteristics [11]. In Brazil, the process of waste degradation and leachate formation is favored due to the country’s tropical climate and its high temperatures and precipitation volumes [10]. According to Alfaia et al. [12], it is estimated that more than 50% of the total urban solid waste generated in Brazil is matter of organic origin. Table 1 shows the physical–chemical characteristics of leachates from some landfills in different regions of Brazil. In general, the ratio BOD5/COD indicates the biodegradability of the effluents. As can be seen from Table 1, this ratio for old landfills is quite low, indicating low biodegradability of organic compounds. For example, for the Belo Horizonte landfill, the BOD5/COD ratio is 0.029. This value is according to the research by Kjeldsen et al. [16], who observed that old landfills, which are in the action phase of methane-producing bacteria, have BOD5/COD proportion lower than 0.1. Another point to highlight is the high pH values of the analyzed leachates, which are indicative that the landfills are in the methane production phase. Concerning the values of ammonium nitrogen, the leachate from Bandeirantes City landfill has the highest value observed among the characterizations presented, with a concentration of 2,183 mg/L. The majority of ammonium nitrogen found in leachates comes from the degradation of proteins, which can consist of 0.5% of the dry mass of the residue. Ehrig's (1988) [17] research showed that old landfills contain a high concentration of ammoniacal nitrogen and remain virtually constant over time. Thus, leachates from old landfills have high concentration of ammonium nitrogen, which can make the place a strong source of pollution, making it urgent to use technologies for its elimination [16]. Heavy metals are present in leachate as a result of distinct wastes that are discarded in landfills,

Leachate Treatment in Brazil and Potential Technologies …

169

Table 1 Physical–chemical characteristics of leachates from some Brazilian landfills Landfill location (Brazilian cities) Parameters

Bandeirantes (State of São Paulo)

Florianópolis (State of Santa Catarina)

Belo Horizonte (State of Minas Gerais)

São Leopoldo (State of Rio Grande do Sul)

Landfill age (years)*

30

9

9

10

pH

8.1

8.1

8.3

8.4

BOD5 (mg/L)

2060

1683

68

335

COD (mg/L)

7373

3581

2354

2264

BOD5/COD

0.28

0.47

0.029

0.148

NH4+-N (mg/L)

2183

1419

1055

–

Cl− (mg/L)

–

–

2190

895

Alkalinity (mg CaCO3/L)

10,720

5863

5263

–

Total solids (mg/L)

–

–

8801

–

Size (ha)

140

–

133

135

Form of operation

Sanitary landfill

Sanitary landfill

Sanitary landfill

Sanitary landfill

References

[12]

[13]

[14]

[15]

(–) not measured * at the time of publication

such as light bulbs, paints, packaging used for chemicals and pharmaceuticals, and batteries. [17]. In accordance to Baun and Christensen [18], heavy metals are capable of forming complexes with organic molecules, salts, and alkalis, which are commonly found in leachates. As shown in Table 2, the concentrations of heavy metals in some leachates from Brazilian landfills are low, and this is due to the reduced solubility of metals at pH > 7. In the study of Kulikowska and Klimiuk [4], it was discovered that concentrations of heavy metals are higher in young landfills due to the acidic pH. As the landfill becomes stable, when it reaches a methane production phase, the pH becomes alkaline, and the metals become insoluble in this environment.

2

Advantages and Disadvantages of the Main Leachate Treatment Methods

The biological method of nitrification/ denitrification is probably the most effective and cheapest process to remove nitrogen from

the leachate. However, biological treatment is impaired by the presence of specific toxic substances, such as polyaromatic hydrocarbons (PAHs), adsorbable organic halogens (AOXs), and polychlorinated biphenyls (PCBs), or by the presence of biorefractory organics, such as humic acids or surfactants. The limited amount of biodegradable organics cause the denitrification efficacy to be reduced, particularly for already stabilized landfills (more than 10 years of operation) [22]. Among the main biological treatment processes are aerobic lagoons, in which one of the limiting factors is temperature control, because, in order to avoid slowing the bacterial growth, the temperature must be kept around 15 °C. Moreover, although this process guarantees a good removal of ammonium nitrogen, other treatments are necessary to meet the limit values of BOD5 and COD. On the one hand, the activated sludge process is especially efficient in the removal of organic matter, inorganic nutrients, and ammonium, but it has also some disadvantages concerning the treatment of landfill leachate, such as insufficient sedimentation of sludge, need for excessive aeration with the

170

M. I. Canabarro et al.

Table 2 Concentration of heavy metals in leachates from some Brazilian landfills Landfill location (Brazilian states)

Heavy metals concentration (mg/L) Ni

a

Cr

Cd

Zn

Cu

Pb

Fe

NM

0.23

0.04

0.10

NM

5.68

[14]

Santa Catarinaa

0.285

0.33

0.04

NM

1.16

0.380

7.77

Apud [19]

Pernambucob

0.47

ND

ND

0.28

0.11

0.24

7.67

[20]

Minas Gerais

< 0.02

References

ND not detected NM not measured a Sanitary landfill: According to the Diagnosis of Urban Solid Waste Management [21], it is defined as any facility with permanent technical and operational control in order to prevent waste and its liquid and gaseous effluents from causing damage to public health and/or to the environment b Controlled landfill: It refers to any installation with some care, mainly related to the safety of workers and the transit of people in the unit. It is characterized as the intermediate stage between open dumps and sanitary landfills

consequence of high-energy consumption, excess production of sludge and bacterial inhibition due to the concentration of ammonium [23]. On the other hand, membrane bioreactors have the advantages of low sensitivity to leachate changes, ease of controlling sludge age, lower sludge production, and need of limited space for operation. Even so, the main disadvantages that still limit its use are the high cost of membranes, cleaning, eventual replacement and maintenance due to clogging, and high-energy consumption [24, 25]. Conventional physicochemical methods of treating leachate, such as air stripping, coagulation, flocculation, and sedimentation, are often initially expensive because of plant equipment costs, energy requirements, and frequent use of additional chemicals. For example, coagulation– flocculation shows disadvantages, such as producing a consistent volume of sludge and increasing concentrations of aluminum or iron coagulants in the liquid phase [26]. Regarding the electrocoagulation process, several process variables can influence the treatment efficiency, such as the design of the reactor used, the electrode material, current density, pH, and conductivity [27]. Among the advantages of this process are the removal of fine colloidal materials, the short period for starting the process, the simple operation, the no addition of chemicals, the high capacity of pollutant removal, the easy collection of the sludge produced, and the easy control of the process [28]. The main disadvantage of the

process is the large consumption of the anode as well as the passivation of the electrode leading to a reduction of current and, consequently, a reduction in the efficiency of the process, the need to add some salts to some effluents that are not conducive, the composition of the sludge generated, likely present different materials, including chemicals that are not eco-friendly, if not managed properly [29]. Other methods such as reverse osmosis or activated carbon adsorption only transfer pollution from one phase to another and do not solve the environmental problems [30]. Among the advantages of reverse osmosis are the fact that it is a very effective process when applied in the leachate polishing stage. The disadvantages are the short lifetime of the membranes (around 5 years) due to clogging, the need for periodic cleaning, and the high-energy consumption [31]. Among the advanced oxidation processes (AOPs) are the Fenton process, which is very popular due to its advantages, such as the wide range of applications and the rapid degradation and mineralization of pollutants. However, at the same time, it has a major disadvantage related to the narrow working range for pH (between 2 and 3) [32]. The combined process of ozone (O3) with hydrogen peroxide (H2O2), known as the peroxone process, is limited by several factors such as the low water solubility of ozone, the energy consumption, and its sensitivity to some factors including pH, temperature, type of pollutant, and the occurrence of secondary reactions

Leachate Treatment in Brazil and Potential Technologies …

that also consume hydroxyl radicals (∙OH) [33, 34]. However, the main advantages of this process are related to the simplicity of operation and the fact that it has an excellent bactericidal activity. For these reasons, this method was developed as an essential disinfection step for drinking water treatment [35]. Currently, there is no single process capable of providing adequate treatment of landfill leachate. Conventional wastewater treatment processes usually are not capable of achieving a satisfactory level of degradation of the toxic substances present. Therefore, in Table 3, the advantages and disadvantages of the major methods of sanitary landfill leachate treatment are presented.

3

Leachate Treatment in Brazil

Monitoring the release of effluents began in Brazil in 1970. In 1986, the first classification and categorization of water bodies in Brazil was elaborated, through the formulation of Resolution No. 20 of CONAMA [64]. Resolution No. 430/2011 is the most recent version of this resolution [65]. This resolution addresses the conditions, parameters, norms, and guidelines for the management of effluent discharge into receiving water bodies [19]. Some detection limits for some parameters shown in Resolution No. 430/2011 CONAMA are presented in Table 4. The treatment of landfill leachates in Brazil through biological processes is still the most common approach. This can be justified by the fact that this method presents very effective in the removal of biodegradable organic matter, indicated by the parameter of BOD5, and because it is economically viable [66]. An example of a biological treatment process is the stabilization ponds, which are extensively used in rural areas due to the warm climate and the availability of large areas, which are necessary for this type of treatment [14, 68]. Table 5 shows some sanitary landfills in Brazilian cities, the treatment techniques that are applied to the leachate generated, information regarding the period of operation, and the amount of solid waste received per day.

171

It needs to be emphasized that in Brazil, the application of biological treatment to landfill leachate is a way to revive the model that was integrated in wastewater treatment plants already decades ago [79]. Due to the lack of knowledge about leachate properties, leachate treatment plants use the same process and design parameters as wastewater treatment plants. Many of these plants are still operating under the same conditions today, although biological processes are not the technology best suited and indicated for older and low BOD5/COD (biodegradable) leachates [20]. According to the investigation conducted by Filho [77], there is no reliable information regarding the technologies used in leachate treatment in Brazil. However, it is possible to notice a trend of combined treatment with sanitary sewage in Sewage Treatment Plants (STP). This alternative is chosen due to the low cost and being established in CONAMA Resolution No. 430 [65], not to be mandatory to comply with the removal of ammonium nitrogen, when leachate and sewage are treated together in the WWTP. However, recalcitrant organic matter present in leachate may not be treated in bioreactors, which are commonly used in sewage treatment. Thus, it is important to verify the feasibility of performing the joint treatment of both wastewaters, and if the leachate can just be diluted in the sanitary sewer. Table 6 presents some characteristic parameters of leachates from Brazilian landfills. An ideal strategy for the treatment of Brazilian landfill leachates would be to initiate it concomitantly with the beginning of the landfill operation. Thus, a biological process would be adequate and efficient to treat the leachate generated, because at the initial ages of the landfill, the leachate presents a higher number of biodegradable compounds. As this is not the reality of most Brazilian landfills, and in many places, the generated leachate is daily mixed with older leachates, some treatment strategies are possible, such as (i) to apply advanced oxidation processes such as ozonation, photocatalysis, and Fenton; or (ii) to operate membrane processes, such as reverse osmosis, nanofiltration,

X

X

X

X

X

X

X

X

X

Advanced oxidation processes (AOPs)

Fenton

Photo Fenton

Electrochemical oxidation

Electro-Fenton

X

Ultrafiltration (UF)

X

X

X

X

Membrane filtration

Ion exchange

X

Adsorption

X

X

Coagulation– flocculation

X

X

X

X

X

X

NH4+N

Air stripping

X

X

X

Electro coagulation

X

X

Chemical oxidation

X

X

Chemical precipitation

Evaporation

X

X

Recycling

X

X

Spray irrigation

X

X

COD

Target pollutants

BOD5

Co-treatment with sewage

Method

X

X

X

X

X

SS

X

X

X

X

Heavy metals

X

X

X

X

X

X

ROC

X

X

X

X

X

X

X

X

X

X

Simple operations/ process, easy maintenance

X

X

X

X

X

X

X

X

Low cost

Main advantages

X

X

X

Rise BOD/ COD ratio

X

X

X

X

X

X

X

X

X

X

X

Good quality effluent

Table 3 Advantages and disadvantages of various treatment methods applied to landfill leachate

X

X

X

X

X

X

X

X

X

X

X

X

X

X

X

X

X

X

X

X

X

X

X

X

Not economically attractive

Main disadvantages Generation of sludge or contaminant by products

X

X

X

X

X

High energy required

X

X

X

X

X

X

Operational issues (odor, gas, operation, maintenance, fouling)

(continued)

[3, 58, 59]

[19, 40, 44, 51]

[54, 57]

[39, 45, 54, 57]

[54–56]

[36, 41, 51–53]

[27, 48, 51]

[40, 49]

[27, 41, 44, 46, 50]

[40, 49]

[40]

[3, 44, 46–48]

[40, 41, 45

[3, 41, 45]

[16]

[12, 16, 39, 41, 43, 44]

[16]

[36–42]

References

172 M. I. Canabarro et al.

X

X

X

X

X

X

Activated sludge process

Anaerobic digestion

Phytoremediation system

Membrane bioreactor

X

X

Fungal treatment

X

X

X

X

X

X

X

X

X

NH4+N

X

X

X

X

X

X

SS

X

X

X

Heavy metals

X

X

ROC

X

X

X

X

X

X

X

X

X

X

X

X

X

Low cost

Main advantages Simple operations/ process, easy maintenance

= ammonium nitrogen; ROC = recalcitrant organic compounds SS = suspended solids; Source adapted with permission from Jagaba et al. [63]

NH4+-N

X

X

X

X

Anaerobic sequencing batch reactor (ASBR)

Upflow anaerobic sludge blanket

X

Rotating biological contractors

Trickling biofilters

X

Moving-bed, suspended carrier biofilm reactors

X

X

X

X

X

Aerated lagoons and stabilization ponds

X

Reverse osmosis (RO)

X

Distillation

X

COD

Target pollutants

BOD5

Nanofiltration (NF)

Method

Table 3 (continued)

X

Rise BOD/ COD ratio

X

X

X

X

X

X

X

X

X

X

X

Good quality effluent

X

X

X

X

X

X

X

X

X

X

X

Not economically attractive

Main disadvantages Generation of sludge or contaminant by products

X

X

X

X

X

High energy required

X

X

X

X

X

X

X

X

X

Operational issues (odor, gas, operation, maintenance, fouling)

[40]

[43, 44]

[3, 23, 41]

[62]

[23, 30, 41, 43, 55]

[3, 40, 41]

[45, 51]

[19, 23, 41, 43, 44]

[23, 30]

[43, 44]

[3, 19, 43, 44, 54]

[60, 61]

[3, 40, 57, 59]

[3, 59]

References

Leachate Treatment in Brazil and Potential Technologies … 173

174

M. I. Canabarro et al.

Table 4 Some detection limits of resolution Nº 430/2011 CONAMA Parameter

Limit

pH

5–9

BOD

Minimum removal of 60%

Ammonium nitrogen

20 (mg/L)

Settleable materials

1 (mg/L)

Fe

15 (mg/L)

Ni

2 (mg/L)

Cd

0.2 (mg/L)

Zn

5 (mg/L)

Cu

1 (mg/L)

Pb

0.5 (mg/L)

ultrafiltration, and MBR; or even (iii) to adopt a combination of processes, which sometimes provides better efficiency in the removal and/or reduction of pollutants. An example of a combined process applied in leachate treatment is the one investigated in the work of Chen et al. [78]. In this work, a combined coagulation-ozonation process was developed to treat concentrated leachates resulting from reverse osmosis treatment. As a result, they obtained a high removal of organic compounds in the coagulation process, and after applying ozone, the organic substances resistant to coagulation were removed. Another example of efficiency in the combination of processes can be seen in the work of Rohers et al. [79], in which leachate from a Brazilian landfill was used. This work was carried out in sand filters and activated carbon columns, as a pre-treatment alternative for landfill leachate and with the intention of improving the efficiency of the subsequent biological process. The results showed reductions above 74% for COD, 47% for BOD5, 93% for color, 90% for ammonium, and an increase of BOD5/COD ratio from 0.3 to 0.9 meaning higher biodegradability. In Fig. 1, some treatment possibilities for the leachate according to their BOD5/COD ratios have been depicted. It is important to emphasize that presently, there is not a single treatment process capable of treating landfill leachate with a maximum efficiency, fitting it to the standards for effluent

discharge. The choice of the appropriate treatment should take into consideration various factors, such as type and amount of solid waste received at the landfill, rainfall index and climate at the landfill site, humidity of the waste, and amount of leachate generated. Besides thinking of treatment alternatives for leachate, it is also necessary to look for ways not to generate it or to reduce the amount generated. One possibility for this is the incineration of municipal solid waste instead of its disposal in landfills. Incineration is a process for treating solid waste that involves the combustion of the organic substances present in the matter that compose it. During the incineration process, the waste presents a reduction in volume, weight, and initial hazardous characteristics through controlled combustion. In the work by Santos [80], comparison between the disposal of solid wastes in landfills and their incineration is presented. In the work of Machado [81], it is shown that the energy use of urban solid waste from the Bauru City/SP is possible through the incineration process and would be enough to supply approximately 39,000 residences in one month, or to collect about R$ 78,000.00 per day (US$ 15,100) of incinerator operation.

3.1 Comparison Between Leachate Treatment in Brazil and in Other Countries Investments in landfill leachate treatment technologies are directly related to the purchasing power of cities and/or countries. Therefore, when comparing developed countries such as in Europe and North America with developing countries such as in Asia and South America, significant differences in municipal waste management and leachate treatment systems can be observed. In Brazil, it is possible to observe these differences within each region. Researchers around the world have been seeking for more knowledge about wetlands as an alternative for the treatment of leachate from landfills, mainly in tropical, subtropical, and

Leachate Treatment in Brazil and Potential Technologies …

175

Table 5 Treatment processes applied to leachate in some Brazilian cities Landfill location

Final destination

Operating period

Amount (ton/day)

Treatment process

References

Osasco (São Paulo)

Sanitary landfill

1991– 2017

800

System consisting of biological treatment and bioreactor combined with ultrafiltration membranes (MBR)

[67]

Rio Claro (São Paulo)

Sanitary landfill

2001current

175

System composed of three lagoons and bioreactor combined with ceramic ultrafiltration membranes (MBR)

[68]

Dourados (Mato Grosso do Sul)

Sanitary landfill

2004current

180

Biological treatment (two anaerobic ponds, one maturation pond). When the volume of the maturation pond reaches its maximum, the leachate goes through a process of recirculation in the landfill

[69]

Santa Maria (Rio Grande do Sul)

Sanitary landfill

2008current

500

Combined system of nanofiltration and reverse osmosis membranes + air stripping tower

[70]

Minas do Leão (Rio Grande do Sul)

Sanitary landfill

2001current

3600

Combined system consisting of equalization tanks, physical– chemical treatment with calcium hydroxide, air stripping, anoxic pond, activated sludge, nanofiltration, and geobags

[71]

Victor Graeff (Rio Grande do Sul)

Sanitary landfill

2018current

400–450

Combined system with 1st stage of biological treatment, 2nd stage of treatment by electrocoagulation, and 3rd stage in an aerated lagoon, with final polishing in WTP*

[72]

Giruá (Rio Grande do Sul)

Sanitary landfill

2011current

500

Treatment with reverse osmosis membranes

[73]

Caieiras (São Paulo)

Sanitary landfill

2002current

10,500

Treatment performed externally

[74]

São Francisco do Conde (Bahia)

Sanitary and industrial landfill

2009current

400

Treatment performed externally

[75]

Marituba (Pará)

Sanitary landfill

2015current

1300

Treatment with reverse osmosis membranes

[76]

*

WTP = Water Treatment Plant

temperate regions. There are records of over 10,000 wetlands in North America [19]. The wrong disposal of urban solid waste in open-air dumps is still a reality in Brazilian territory, representing about 17,5% of the waste

generated [82]. Unfortunately, this incorrect disposal action is also quite common in Asian countries. According to the study by Yadav and Samadder [83], 60% of the total waste collected in India goes to open-air dumps, 50% in

176

M. I. Canabarro et al.

Table 6 Real parameters of some leachates from Brazilian cities Conductivity (µS/cm)

pH

Total alkalinity (mg/L)

BOD5 (mgO2 /L)

COD (mgO2 /L)

BOD/ COD

Ammonium nitrogen (mg/L)

Santa Maria (RS) raw leachate

32,700

7.9

8876

2516

7673

0.33

2583

Minas do Leão (RS) raw leachate

27,600

7.9

15,500

2200

12,068

0.18

4060

São Francisco do Conde (BA) raw leachate

13,720

8.9

–

860

2217

0.39

92

Marituba (PA) raw leachate

36,040

7.7

–

2365

–

–

2472

Fig. 1 Flowchart with treatment possibilities for leachate according to its biodegradability (BOD/COD)

Bangladesh, 85% in Sri Lanka, 80% in Pakistan, and 40% in Indonesia. Based on these figures, it can be highlighted that Brazil, even if slowly, is investing in the right disposal of waste, which can be corroborated by the enlargement of the coverage of collection services, the closing of open-air dumps, and the development of research in the area [19]. Regarding the development of leachate treatment technologies, landfills in metropolitan areas of Brazil have opted to outsource their services, ultimately favoring the use of modern wastewater treatment technologies such as the use of membranes (MBR, reverse osmosis, ultrafiltration, and nanofiltration) [20]. Membrane filtration technology is one of the high-value treatment methods used in North America for municipal and industrial wastewater and sanitary landfill

leachate [88, 89]. An example of membrane application for leachate treatment is at Seneca Meadows Landfill, which is the largest active landfill situated in New York, USA. In agreement with Robinson, European countries have been using membrane processes such as MBR systems to treat leachate since 1990. An example is the leachate treatment of the Hannover Landfill in Germany, which consists of an MBR system with oxygen injection followed by membrane ultrafiltration. Finally, it is possible to highlight the lack of available data on leachate treatment systems in Brazilian landfills. This can be attributed to the lack of information available to companies managing large landfills and a gap in government demand for and management of this information to make it public [20].

Leachate Treatment in Brazil and Potential Technologies …

4

The Future of Leachate Treatment: Studies in Brazil

In accordance with Contrera et al. [64], a variety of Brazilian cities have been struggling with the treatment of leachate from sanitary landfills. In fact, the application only of stabilization ponds (aerobic and anaerobic) has proven not to be the most suitable treatment for ancient leachates, which have recalcitrant organic compounds in their composition (humic and fulvic acid, for example). Ideally, leachate treatment should be initiated at the beginning of the landfill operation, since the biological treatment is more efficient in removing biodegradable organic matter in young landfills. However, the older the landfill, the more difficult it is to remove the refractory compounds, and the concentration of the pollutants can exceed the limits of the legislation for the disposal of the final effluent, which consequently results in the need for higher costs with advanced treatments for the leachate. Thus, there is a need to combine biological processes with other complementary treatments. Therefore, for the removal of organic compounds and ammonium nitrogen, both characteristics present in abundance in sanitary landfill leachates, the treatments that employ activated carbon present a great efficiency when used to treat sanitary landfill leachates. The application of activated carbon associated with the activated sludge system has been studied recently by Brazilian researchers as a pre- or post-treatment step of the leachate [20] (Fig. 1). A study on activated carbon powder treatment system that associates biological oxidation in activated sludge system with adsorption on activated carbon powder was conducted by Fernandez Bou et al., removing turbidity and associated with the combined treatment of landfill leachate with domestic sewage. In this study, the use of activated carbon was favorable for the removal of color and dissolved solids, although not significant for the removal of turbidity.

177

After several kinds of research, many advances have been obtained regarding advanced oxidation processes (AOPs), and those, in which the Fenton process is applied, are considered the most economical processes by Deng and Englehardt [84]. An alternative for treating effluents with large amounts of recalcitrant compounds, which is usually associated with leachate from old or closed landfills, is the application of ozone oxidation processes as a treatment step [57]. Scandelai et al. [85] analyzed the applicability of ozonation process combined with photocatalytic metal oxides (O3/TiO2 and O3/ZnO) for the treatment of leachate from the Maringá landfill (Paraná State). In this research, COD was reduced by 63% and color by 98.2% when using ZnO combined with O3. Biological membrane reactor systems (MBRs) are also largely investigated for wastewater treatment in Brazil. In accordance with Ahmed and Lan [86], MBR technology is promising, as it is an advanced biological process that uses a membrane separation unit, and unlike the activated sludge system, it does not use a conventional sedimentation unit. Membrane separation processes, mostly reverse osmosis, have been considered as a proper approach for the treatment of sanitary landfill leachates, because it results in a treated effluent that meets the emission standards of the legislations [87]. However, for this process to be successfully applied in a treatment plant, a suitable pre-treatment needs to be applied to reduce fouling of the membranes [3]. Besides the processes cited, it has been investigated in Brazil recently the possibility of treating the sanitary landfill leachates through the process of evaporation to reduce mainly ammonia and COD. Thus, one of the investigated areas is solar distillation, which uses solar radiation (direct and diffuse) to promote heating and phase change of the liquid to be treated in a simple and low-cost system [88]. In the study conducted by Maroneze et al. [88] using a solar distiller, it was possible to remove BOD5, COD, and color higher than 99%.

178

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Conclusion

The leachates from sanitary landfills are effluents that present a high potential for pollution due to their composition, which is highly variable and heterogeneous. The differences observed between different leachate compositions are due to the composition of the waste that is received at the sanitary landfills, the age of landfill, and the climate at the region. Therefore, due to this large variation, it is difficult to suggest just one treatment technique capable of being applied efficiently to all leachates. Biological treatment processes are widely used because they are simple to operate and are generally cost-effective. However, it is important to emphasize that this type of treatment only becomes effective when applied to sanitary landfill leachates with a high BOD5/COD ratio that is valid for new leachates with good biodegradability. Physical–chemical processes, such as coagulation–flocculation and electrocoagulation, are also processes that show high efficiency when applied to the treatment of more stabilized leachates, that is, leachates from older landfills (  9 years) having a low BOD5/COD ratio. However, this type of treatment has the major disadvantage of using coagulants, which makes the process more expensive and usually requires another treatment step to remove the residual coagulant. Several studies are reporting the efficiency of advanced oxidation processes (AOPs) for the treatment of landfill leachates, as presented throughout this work. Besides the biological, physicochemical, and AOPs processes, distillation emerges as a sustainable alternative for leachate treatment, because it does not require the addition of chemicals or pH adjustment of the leachate and not generate excess sludge at the end of the process. In Brazil, landfill leachates are treated mostly by biological processes, such as aerobic and anaerobic ponds, biological filters, activated sludge, and wetlands. In some landfills, some more advanced treatment technologies are applied, however, involving high costs of

installation, maintenance, and operation, as is the case of reverse osmosis technology. Because leachate is a type of wastewater, it needs to be treated before it can be discharged into receiving water bodies. Having this in mind and the fact that Brazil, in its great majority, uses only biological processes for the treatment of leachates, one realizes the need for investments in more promising and efficient treatment techniques that ensure a good quality of the final effluent.

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Sanitation Context and Technological Challenges to Municipal Wastewater Management in Africa Patrick Ogola Onyango

are needed for sustainable wastewater treatment in Africa, a continent that is suffering from water-stress and high burden of sanitation-related diseases. Existing evidence shows that chemically enhanced primary treatment (CEPT) is an ideal first step technology for wastewater treatment in the developing world because of its low cost, efficiency, and ease of implementation. This chapter presents the complex sanitation context in Africa. It summarizes the performance of commonly used wastewater treatment systems. In addition, it revisits the proposition that CEPT is an efficient, low cost and adaptable wastewater treatment technology that offers great promise to address wastewater treatment challenges facing municipalities in the developing world including Africa.

Abstract

The debate on the municipal wastewater problem in the developing world is not new. However, the inaction on the part of governments or municipal urgencies is worrying in the face of persistent health burden arising from release of partially treated or untreated wastewater into the environment. The inaction derives in part from high costs and expertise constraints that are needed to set up and operate advanced wastewater treatment technologies. In part, the inaction even in the face of mounting threat to public health may be due to a deep-rooted perception that wastewater is not a resource or is inherently bad and must thus be disposed. Consequently, investment in safe disposal or in technologies for recycling and reuse of wastes remains low. In addition, population growth in many municipalities continues to increase and so is the production of high volumes of wastewater and other wastes. Furthermore, the situation is exacerbated by emerging pollutants of concern that cannot be removed from wastewater by existing technologies. Technologies that can overcome these constraints and challenges

P. O. Onyango (&) Department of Zoology, Maseno University, Private Bag, Kisumu, Kenya e-mail: [email protected]

Keywords







Africa Sanitation Wastewater treatment Chemically enhanced primary treatment CEPT

1







Introduction

Population growth in African cities and towns continues to over-stretch the existing waste and wastewater management infrastructure. Although the proportion of the world population in urban areas continues to rapidly rise, projected to reach

© The Author(s), under exclusive license to Springer Nature Switzerland AG 2023 E. Debik et al. (eds.), Wastewater Management and Technologies, Water and Wastewater Management, https://doi.org/10.1007/978-3-031-36298-9_13

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68% by 2050 [1], the pace at which capacity for wastewater treatment in urban areas is improved to deal with these changes continues to lag. This is particularly the case in the developing world. Yet more than 2 billion of the world population, 52% of which are in Africa, do not have access to safe water for domestic use [2]. Municipal wastewater results from a wide range of sources including household wastes, institutional discharges (e.g., from hospitals, hotels, schools), and commercial and industrial effluents. According to the United Nations Department of Economic and Social Affairs [1], it is likely that 80% of wastewater in Africa is released into the environment untreated; Jones et al. [3] estimates it to be 48%. Indeed, many urban areas discharge untreated wastewater into adjacent rivers, shallow basins or coastal waters resulting in serious threats to human and wildlife health. The consequence is rising opportunities for exposure to raw or untreated wastewater. These exposures are associated with 842,000 deaths that occur annually from diarrheal diseases [2]. Children in particular bear a disproportionate burden of diarrheal diseases that derive primarily from polluted waters. In the current dispensation, the burden of diseases attributable to water pollution are expected to increase. Sustainable Development Goal 6—ensure access to water and sanitation for all—is indeed alive to the public health and environmental challenges arising from water pollution. However, for this goal to be met by 2030, municipalities in Africa must put in place steps to manage municipal wastewater and other wastes. The challenge municipal and other urban waste management agencies face is that waste is diverse. For example, hospital wastes are likely to be rich in pharmaceutical products than say household waste. On the other hand, composition of industrial wastes is variable and depends, for the most part, on the type of industry although they are mostly rich in acids, metals and salts. Differences in municipal wastes imply that different technologies are needed. In megacities across the developing world, however, wastewater from different sources converge into a single pool. The consequence then is that very

P. O. Onyango

sophisticated and costly technologies are needed to treat such wastewater before it is released into natural waterways or before it is recycled and reused. Unfortunately, the traditional wastewater treatment technologies that are common in most developing countries are not designed to deal with some of the emerging pollutants such as pharmaceuticals, cosmetics and personal care products [4]. In addition to the challenges of treating wastewater in municipalities in much of the developing world, there is also a deep-rooted perception, partly informed by societal norms or culture, that wastewater and associated biosolids are inherently bad [5]. However, there is a large body of evidence that human excreta—also known as night soil—has been and continues to be used as fertilizer for both terrestrial agricultural production and aquaculture. The problem facing wastewater management and its threat to public health increasingly manifests in livestock production in urban as well as urban areas encroach into peri-urban and in some cases rural areas. The consequence is the presence of livestock in urban areas. Livestock in urban areas scavenge in refuse sites and drink water from open storm drainages, temporary water pools and open lagoons (Fig. 1). The consequence is that any contaminants or pollutants in such water make into the food chain, where they potentially threaten to human health. This chapter presents the complex African sanitation context and reviews the performance of the predominant wastewater treatment systems in the continent. In addition, the chapter revisits the utility of an old yet versatile wastewater treatment technology, chemically enhanced primary treatment (CEPT) to tackle the sanitation challenges in developing countries including those in Sub-Saharan Africa.

2

The African Sanitation Context

Much of Africa has a complicated sanitation picture that is characterized by a high burden of sanitation-related diseases arising from poor sanitation coverage. However, the nature and

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Fig. 1 Open lagoon in Nyalenda, Kisumu City, Kenya. Note livestock grazing around the lagoon. Exposure of livestock to partially treated wastewater has serious public health consequences

distribution of the burden of the wastewater management problem is not the same across the continent. For instance, countries in North Africa appear to be far ahead of their counterparts in the rest of Africa but especially those in Sub-Saharan Africa where the problem is exacerbated by several interconnected factors. First, the human population in the continent is growing at a rate of 2.4% and that in Sub-Saharan Africa currently at 3.9% annually, which is much higher than the global average rate of 1.5% [1]. For instance, between 1990 and 2008, the urban population in Africa increased from 88 to 175 million. Most of the urban population in Africa lives in slums and other informal settlements. Rapid increase in human population does not only overstretch the existing wastewater treatment infrastructure but also constrains both the opportunities for expanding existing wastewater and other waste management systems, and those for establishing new ones, as available spaces are converted into human settlements or for socioeconomic activities to support the growing population. Yet, although the urban population has been growing, the municipal wastewater treatment capacity continues to lag behind and unable to cope with the ever-increasing large volumes of urban wastes. In parts, this explains why 2017 nearly 800 million people in Africa did not have access to basic sanitation services, which is defined as the use of improved sanitation facilities that are not shared with other households [6]. The consequence is that a large proportion of the

population is exposed to sanitation-related diseases such as diarrheal ones that account for a significant cause of morbidity and mortality in children [7, 8]. The lag between population growth and its concomitant demand for sanitation services and capacity of municipal waste management is mostly attributed to high capital, operational and management costs associated with advanced wastewater treatment technologies as well as to expertise constraints needed for such technologies. In most African countries, wastewater remains an unwanted product that must be disposed of, but whose disposal municipalities are unwilling to invest in. Second, either independent of or because of the high rate of population growth, the continent has been witnessing high rate of growth in the industrial, manufacturing, construction and mining sectors [6]. Expectedly, the consequence of these economic activities has been a fast pace of urbanization and production of large volumes of waste beyond the capacity of the existing waste management infrastructure. In addition, the economic activities complicate any efforts at waste and wastewater management because of the nature of the wastes that they generate. Of particular concern is the production of contaminants and pollutants of emerging concern, whose removal from water require advanced and costly waste management technologies. Third, available data suggest that Africa is the second driest continent after Australia [9]. In the

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context of climate change and its effects including flooding and prolonged droughts, the continent is likely to find itself in a vicious cycle of food insecurity, increased levels of school dropouts as school-aged children quit school to join the family labor force, deforestation as people resort to illicit or illegal activities such as charcoal burning among others. In Africa, for example, droughts are associated with increased prevalence of anemia, scabies, dengue, and increased incidence of disabilities in children [10]. When faced with such challenges, governments are likely to redirect limited financial resources to overcome such challenges that ultimately investment in sanitation and wastewater treatment will fall even lower than the current levels. Yet, ironically, wastewater treatment may be the solution to some of these challenges associated with water scarcity. For instance, Hristov et al. [11] used agro-economic modeling to demonstrate that treated water reuse is a potential alternative to water abstraction for agricultural production in the European Union. Fourth, the continent suffers the twinproblems of low capital capacity for waste and wastewater management and the perception that wastewater is inherently bad and has no value.

P. O. Onyango

Previous research has shown that a country’s income level, ceteris paribus, is a good predictor of its capacity for wastewater management [12]. These disparities ultimately reflect themselves on the amount of wastewater that is collected, treated, and reused. For instance, only 61.1% the generated wastewater, 25.7% of which is treated, out of which only 21.5% is reused, respectively, in Sub-Saharan Africa, compared with 100%, 100% and 90.0% in North America, and 99.8, 99.8 and 42.5% in Europe [3]; (Fig. 2). Although there is a persistent perception against the wastewater as a recourse, evidence from elsewhere demonstrate its untapped potential. For example, in China and South America, human excreta have long been used either in treated or untreated form to fertilize farms. Recent estimates indicate that 4–20 million hectares are irrigated with municipal wastewater [13] particularly in urban areas, where 26–90% of vegetables consumed by the urban population is produced in farms irrigated using wastewater [13]. Recycling and reuse of wastewater and resultant sludge will help to reduce the pressure on existing landfills and demand for establishing new ones thereby reducing the ecological footprint of wastewater treatment. However,

Fig. 2 Regional disparities in wastewater production, collection, treatment and reuse [3]

Sanitation Context and Technological Challenges to Municipal …

recycling and reuse of wastewater particularly for agricultural production must be regulated so as to protect farmers and consumers from products from farms irrigated or fertilized using treated wastewater or associated biosolids [5, 14]. The increasing water stress may force many municipal agencies to increase investment in processes that increase opportunities for use of treated wastewater.

3

Technological Constraints to Wastewater Management in Africa

Urban wastewater treatment in municipalities in the developing world face a wide range of challenges including high capital investment for wastewater treatment. Consequently, the predominant wastewater treatment systems in Africa include constructed wetlands (CWs), and waste stabilization ponds (WSPs) that are low-cost in capital and operational investment. In the proceeding section, the performance of CWs and WSPs is summarized for some African countries and selected developed nations. The goal here is not to compare the waste removal efficiencies of the wastewater treatment systems between Africa and the developed world. Instead, the goal is to appreciate differences in the

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wastewater management approaches even with the same wastewater treatment system or technology.

3.1 Constructed Wetlands The United States Environmental Protection Agency defines constructed wetlands (CWs) as treatment systems that use natural processes involving wetland vegetation, soils, and their associated assemblages to improve water quality. Expectedly, the performance of CWs varies as a function of several factors including the type of the CW (Free surface water (FSW), horizontal sub-surface flow (HSSF), vertical sub-surface flow (VSSF)), the types of vegetation (macrophytes), the source of wastewater [15] and the climate [16]. For example, Cyperus papyrus has been demonstrated to perform better than Phragmites mauritianus in the removal of COD and BOD [15]. However, even a given macrophyte has been shown to exhibit different removal efficiencies at different CWs reinforcing the evidence that the performance of CWs is influenced by several factors. Table 1 summarizes removal efficiencies for organic matter, nutrients, and suspended solids in some African countries, where data on performance of CWs are available [15]. Performance of CWs appear to vary widely across African

Table 1 Organic matter, nutrients and suspended solids removal efficiencies for constructed wetlands in Africa and Ireland [15, 16] Removal efficiency (%) Type of CW

COD

BOD5

Ammonia

TN

TP

TSS

Selected African countries FWS

68.2

69.7

–

–

–

55.9

HSSF

36.5–97.8

47.8–97.7

23.0–88.5

54.5–74.8

12.6–39.4

56.0–83.8

FWS

82.2 ± 96.9

28.4 ± 49.4

6.7 ± 8.3

20.5 ± 9.3

3.8 ± 2.1

46.4 ± 171.3

HSSF

46.5 ± 65.0

11.8 ± 28.2

10.4 ± 15.5

20.3 ± 19.5

4.6 ± 4.6

19.4 ± 39.0

Ireland

Data reported for CWs that were at operational stage. For data from Africa, ranges for removal efficiency provided for cases, where there are multiple studies on the same type of constructed wetlands. For Ireland, removal efficiencies are presented as arithmetic means ± standard deviations. CW = constructed wetland, FWS = Free water surface, HSSF = horizontal sub-surface flow, VSSF = vertical sub-surface flow; BOD5 = biochemical oxygen demand, COD = chemical oxygen demand, TN = total nitrogen, TP = total phosphorous

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countries and in a developed country such as Ireland (Table 1). Typically, in many developing nations including those in Africa, where CWs and stabilization ponds, as discussed in the subsequent section, are used for wastewater treatment as the main and only wastewater treatment facility. There are, however, exceptions to this general observation. In contrast, in many developed nations, CWs are used primarily for secondary or tertiary wastewater treatment and for specific pollution streams or sources or pollutants and clearly defined emissions criteria [16].

3.2 Stabilization Ponds According to Kayombo et al. [17], stabilization ponds (WSPs) are large shallow basins, in which raw sewage is treated entirely by natural processes involving both algae and bacteria. They are effective in the removal of bacteria including pathogenic species domestic and municipal wastes. Their design is such that ponds are often arranged as series of anaerobic and facultative ponds. Edokpayi et al. [18] demonstrated average reduction efficiencies of different trace metals in WSPs as follows: Al (31.4–82.9%), Cr (12– 30.8%), Cu (13.1–77.1%), Fe (17.7–63.5%), Mn (23.0–80.6%), Pb (0–100%) and Zn (5.5–94.8%) in Vhembe District, South Africa. They also reported COD removal efficiency of WSP in Tanzania (52.4–71.2%), South Sudan (− 22.8%), Ghana (75.5%), Nigeria (− 25.2%) and South Africa (− 10.94%). Regardless of the type of wastewater treatment technology used, municipalities in many African countries appear to be asking the question: Do we have a wastewater treatment facility or system? On the other hand, agencies in developed countries appear to be asking the question: What wastewater treatment system is appropriate for treatment of X type of wastewater? The two questions may be motivated by differences in details in both policies and policing. In Ireland, for example, wastewater treatment is regulated by both national and EU Water

P. O. Onyango

Framework Directive [12], which increases compliance to set standards.

4

Way Forward for Wastewater Management for Municipalities in Africa—Chemically Enhanced Primary Treatment (CEPT)

The complex sanitation context in Africa requires an efficient and low-cost wastewater treatment technique to achieve a minimum level of wastewater treatment needed to protect public health and the environment. Further urgency is also brought to bear by the increasing burden of waterborne diseases, which although preventable still account for a large burden of morbidity and mortality in the developing world including Africa [5, 8]. In recent years, there has been increased interest on CEPT as a technology with the key characteristics needed to overcome the wastewater treatment needs of the developing world [19]. At a basic level, CEPT refers to the process of adding chemicals to primary influent to promote coagulation and flocculation—facilitating aggregation of solids into flocs—so as to enhance sedimentation of suspended and dissolved solids and their associated carbon, phosphorus and other nutrients from wastewater [20]. It can be implemented in several ways including from the outset, a CEPT design with a dedicated CEPT tank; retrofitting conventional primary treatment plants or stabilization ponds. CEPT’s versatility makes it an excellent candidate for a wide range of urban or locational challenges as illustrated in several case studies, where it has been shown to be efficient in removing carbon, phosphorus, heavy metals, inorganic compounds, and eggs of helminths such as Schistosoma mansoni [21–23]. In this section, key case studies on the utility and suitability of CEPT for the developing world are summarized. Focus is given particularly to CEPT experiences in Mexico, California, Hong Kong, and Brazil, which faced challenges comparable to those presently experienced in the developing world: financial constraints to setting

Sanitation Context and Technological Challenges to Municipal …

up and to operate wastewater treatment plants, and to serve large human population and thus large waste volume in municipalities.

4.1 The Challenge of High Financial Cost Experiences from both Point Loma, San Diego in California and Hong Kong have demonstrated the low cost that is associated with setting up and operating a CEPT plant. At Point Loma, for example, the use of small amounts (* 25 mg/L) of a low-cost coagulant, ferric chloride, in combination with very small amounts of a flocculant, approximately 0.2 mg/L of anionic polymer, resulted in a substantial reduction in operation cost of wastewater treatment. The reductions in cost of plant operation were accompanied by significant improvements in treatment efficiencies. For example, removal of suspended solids (TSS) and BOD increased to 85% and 55%, respectively. Similarly, in Hong Kong, adoption of CEPT increased the removal of suspended solids and BOD to approximately 85% and 74%, respectively, achieved under high production capacity of 40 m3/s [24]. These costs are further reduced by the fact that CEPT is relatively easy to implement in terms of equipment and expertise [25].

4.2 The Challenge of Adaptability Three case studies have been used to illustrate the adaptability of CEPT. First, overhaul of existing wastewater treatment infrastructure may be associated with both logistical challenges and prohibitive costs. Consequently, technologies that are versatile to alterations or modifications Table 2 Removal efficiencies before and after modification of wastewater treatment lagoons in Riviera de Sao Lourenco, Brazil

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are preferable and would tackle the challenges the developing world experience in wastewater management. The Point Loma experience illustrates the utility of CEPT in this regard. In 1985, southern cities in California were faced with the problems of keeping pace with high flowrate that outpaced wastewater treatment plant capacity by a factor of more than two of original design capacity and state regulation, California Ocean Protection Plan that required wastewater treatment plants to achieve suspended solid removal efficiency of at least 75%. Addition of trivalent metal salts solved both problems and did so at a lower cost compared to what would have been spent were the city of San Diego to overhaul its wastewater treatment plant. Second, experiences in Riviera de Sao Lourenco, Brazil, demonstrate, how CEPT fits well into existing wastewater treatment infrastructure, where it was used to improve the performance of existing lagoons, both for in-pond and pre-pond set-ups, and substantial cost reductions were realized [26]; Table 2. Third, although the use of wastewater for agricultural production had improved local farmers’ livelihoods and provided vegetables to its growing population, Mexico City faced a public health threat from the high density of helminth eggs up to 250 eggs/L that were present in the sludge from its primary wastewater treatment plants [21, 27]. This threat was tackled by the use of CEPT in combination with polished sand filters, which together saw a substantial reduction in the density of eggs in the wastewater effluent that was destined for the irrigated farms [28]. The consequence was a substantial increase in agricultural productivity attributable to the high levels of phosphorus and nitrates in the resultant biosolids from the city’s After

Before

In-pond CEPT

Pre-pond CEPT

71.8

NA

85.0

COD (%)

50.5

67.5

70.0

TSS (%)

–

79.5

–

BOD (%)

BOD = biological oxygen demand; COD = chemical oxygen demand; TSS = total suspended solids; NA = data not available

190

wastewater [29]. However, it is important to note that the CEPT sludge still contained helminth eggs that would require further treatment through processes such as disinfection and irradiation.

4.3 The Challenge of High Waste Volume Declining opportunities for setting aside new landfills as old ones get filled present a major challenge to municipal and national wastewater treatment experts. The waste paradox, therefore, is that although the volume of generated wastes continues to rise, spaces for dumping even treated wastes have declined in size. Proponents of CEPT argue that the technology is quite robust with respect to handling large amounts of suspended solids. For example, Point Loma experience illustrates, how small adjustment was employed to deal with a rapidly growing population and its concomitant production of large volumes of wastewater. In sum, the aforementioned CEPT experiences present a strong case for its utility in wastewater treatment in the developing world because of its affordability, efficiency, ease of implementation and flexibility to be used with other technologies. However, CEPT has been shown to have several limitations. First, CEPT is ineffective in removing heavy metals and emerging pollutants such as pharmaceutical products, microplastics, persistent organic pollutants, cosmetics and personal care products from wastewater require more advanced technologies. Second, the use of chemicals, which is at the core of CEPT, is associated with toxicity and health hazards. Third, CEPT produces a large volume of sludge, hence a secondary pollutant, which requires additional treatment [30]. Despite these limitations, urgency is needed to address the public health burden associated with the common practice of releasing untreated or partially treated wastewater into the environment. Existing wastewater treatment plants in Africa can be modified or retrofitted with CEPT improve their wastewater treatment capacity.

P. O. Onyango

5

Conclusion and Recommendation

Sustainable access to safe drinking water and public health are the major water-related concerns in urban areas of Africa. In many places in the continent, available water for domestic use is highly contaminated. CEPT fits well with the predominant wastewater treatment in the continent and presents an effective and low-cost technique. For many municipalities in Africa, CEPT remains the most cost-effective first step and one that can always be followed at a later stage by more advanced biological treatment processes as needed. Additionally, municipal agencies in Africa must embrace wastewater as a source for fertilizer, water, and energy. Investment in behavioral change from the persistent sense of disgust toward wastewater and any products from is needed. Indeed, such behavior change may ultimately attract investment for wastewater management. Lastly, municipal agencies and individuals or enterprises that privately treat wastewater they generate must adopt a targeted approach to wastewater management to link wastewater treatment technology to the nature of pollutants in the wastewater to ultimately safeguard the environment and public health.

References 1. UNDESA. (2019). Millenium development goals report 2009. 2. WHO. (2014). Preventing diarrhoea through better water, sanitation and hygiene: Exposures and impacts in low- and middle-income countries. 3. Jones, E. R., van Vliet, M. T. H., Qadir, M., & Bierkens, M. F. P. (2021). Country-level and gridded estimates of wastewater production, collection, treatment and reuse. Earth System Science Data, 13, 237–254. 4. Pruden, A., Larsson, D. G., Amézquita, A., Collignon, P., Brandt, K. K., Graham, D. W., Lazorchak, J. M., Suzuki, S., Silley, P., Snape, J. R., Topp, E., Zhang, T., & Zhu, Y. G. (2013). Management options for reducing the release of antibiotics and antibiotic resistance genes to the environment. Environmental Health Perspectives, 121, 878–885. 5. UNHSP. (2008). Global atlas of excreta, wastewater sludge, and biosolids management: Moving forward

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the sustainable and welcome uses of a global resource. World Health Organization. Global Resource Information Database. Sanitation and wastewater atlas of Africa. GBD. (2016). Mortality collaborators, 2017. Global, regional, and national age-sex specific mortality for 264 causes of death, 1980–2016: A systematic analysis for the Global Burden of Disease Study 2016. The Lancet, 390, 1151–1210. Prüss-Ustün, A., Bartram, J., Clasen, T., Colford, J. M. J., Cumming, O., Curtis, V., Bonjour, S., Dangour, A. D., De France, J., Fewtrell, L., Freeman, M. C., Gordon, B., Hunter, P. R., Johnston, R. B., Mathers, C., Mäusezahl, D., Medlicott, K., Neira, M., Stocks, M., … Cairncross, S. (2014). Burden of disease from inadequate water, sanitation and hygiene in low- and middle-income settings: A retrospective analysis of data from 145 countries. Tropical Medicine and International Health, 19, 894–905. United Nations Environment Programme. (2010). Africa Water Atlas. Asmall, T., Abrams, A., Röösli, M., Cissé, G., Carden, K., Dalvie, M. A. (2021). The adverse health effects associated with drought in Africa. Science of the Total Environment, 793. Hristov, J., Barreiro-Hurle, J., Salputra, G., Blanco, M., Witzke, P. (2021). Reuse of treated water in European agriculture: Potential to address water scarcity under climate change. Agricultural Water Management, 251. WWAP (United Nations World Water Assessment Programme). (2017). The United Nations World Water Development Report 2017. Jiménez, B., Drechsel, P., Koné, D., Bahri, A. (2010). Wastewater , sludge and excreta use in developing countries : An overview. In P. Drechsel, C. A. Scott, L. Raschid-Sally, M. Redwood, & A. Bahri (Eds.), Wastewater, irrigation and health: Assessing and mitigating risk in low-income countries (pp. 3–27). International Water Management. WWAP (United Nations World Water Assessment Programme). (2017). Wastewater: The untapped resource. Mekonnen, A., Leta, S., & Njau, K. N. (2015). Wastewater treatment performance efficiency of constructed wetlands in African countries: A review. Water Science and Technology, 71, 1–8. Hickey, A., Arnscheidt, J., Joyce, E., O'Toole, J., Galvin, G., O’ Callaghan, M., Conroy, K., Killian, D., Shryane, T., Hughes, F., Walsh, K., Kavanagh, E. (2018). An assessment of the performance of municipal constructed wetlands in Ireland. Journal of Environmental Management, 210, 263–272. Kayombo, S., Ladegaard, N. (2004). Waste stabilization ponds and constructed wetlands design manual. 1–59 Preprint at (2004).

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18. Edokpayi, J. N., Odiyo, J. O., Popoola, O. E., & Msagati, T. A. M. (2021). Evaluation of contaminants removal by waste stabilization ponds : A case study of Siloam WSPs in Vhembe District. South Africa. Heliyon, 7, e06207. 19. Libhaber, M., Oroczo-Jaramillo, A. (2012). Sustainable treatment and reuse of municipal wastewater. IWA Publishing. 20. Neupane, D. R., Riffat, R., Murthy, S. N., Peric, M. R., & Wilson, T. E. (2008). Influence of source characteristics, chemicals, and flocculation on chemically enhanced primary treatment. Water and Environmental Research, 80, 331–338. 21. Jimenez, B., & Chavez, A. (1998). Removal of helminth eggs in an advanced primary treatment with sludge blanket. Environmental Technology, 19, 1061–1071. 22. Johnson, P. D., Girinathannair, P., Ohlinger, K. N., Ritchie, S., Teuber, L., & Kirby, J. (1998). Enhanced removal of heavy metals in primary treatment using coagulation and flocculation. Water (Basel), 80, 472– 479. 23. Haydar, S., & Aziz, J. A. (2009). Characterization and treatability studies of tannery wastewater using chemically enhanced primary treatment (CEPT)-A case study of Saddiq Leather Works. Journal of Hazardous Materials, 163, 1076–1083. 24. Morrissey, S. P., Harleman, D. R. F. (1992). Retrofitting conventional primary treatment plants for chemically enhanced primary treatment in the USA. In R. Klute, & H. Hahn (Eds.), Chemical water and wastewater treatment II: Proceedings of the 5th gothenburg symposium (pp. 401–416). Springer. 25. Chagnon, F., Harleman, D. R. F. (1992). An introduction to chemically enhanced primary treatment efficiency of CEPT. 1–5. 26. Cabral, C., Chagnon, F., Gotovac, D., Harleman, D. R. F., Murcott, S. (1999). Design of a chemically enhanced wastewater treatment lagoon in Brazil. 27. Jimenez, B., Maya, C., & Galvan, M. (2007). Helminth ova removal from wastewater for agriculture and aquaculture reuse. Water Science and Technology, 56, 43–51. 28. Murcott, S., Dunn, A., Harleman, D. R. F. (1996). Chemically enhanced wastewater treatment for agricultural reuse in Mexico. International Association of Water Quality: Biennial Conference. 29. Landa, H., Capella, A., & Jimenez, B. (1997). Particle size distribution in an effluent from advanced primary treatment and its removal during filtration. Water Science and Technology, 36, 159–165. 30. Teh, C. Y., Budiman, P. M., Shak, K. P. Y., & Wu, T. Y. (2016). Recent advancement of coagulationflocculation and its application in wastewater treatment. Industrial and Engineering Chemistry Research, 55, 4363–4389.

Management of Water and Wastewater in Morocco and Arab Countries Abdelmalek Dahchour and Souad El Hajjaji

Wastewater reuses vary within Arab countries, Jordan and Abu Dhabi who stand to be the leaders in this issue with 100% of the available/collected wastewater used. The level of the treatment and reuse depends on the existing sanitation infrastructure. The estimated irrigated area with wastewater or polluted water fells between 4 and 6 million ha. Most of the technologies applied in Arabic world for wastewater treatment are based on primary, secondary, and tertiary processes. The choice of the technology is based on the availability of lands, financial budget, and the technical capacities for maintenance. Usage of wastewater is also facing lack of nonsufficient legislation to meet safe usage.

Abstract

Morocco and Arab countries are located in a very large geographic surface area covering 12.7 million km2 of arid and semi-arid lands dispersed in North Africa and Middle East, with population exceeding 414 million inhabitants. Precipitations range from 150 to 600 mm per year, and the ratio of water is varying from 172 m3/cap⋅year to 1278 m3/capyear. These figures make some countries in situation of scarcity. Some resources are native in certain countries, while some countries are depending on their neighbors. The water scarcity is also aggravated by the level of pollution that reduces its eventual usage. To cope with water scarcity, tendency toward treated wastewater is increasing in most of Arab countries. In 2013, the amount of treated wastewater produced in Arab countries was estimated at 10.9 BCM (billion m3), 50% of which being used for different purposes. Oil-rich countries have relied mainly on desalinization of seawater and brackish water for producing freshwater.

A. Dahchour (&) Agronomy and Veterinary Institute, Rabat, Morocco e-mail: [email protected] S. El Hajjaji Faculty of Science, Mohammed V University in Rabat, CERNE2D Rabat, Morocco

Keywords










Water Wastewater Management Treatment Water scarcity Arab countries

1







Introduction

Morocco and the 21 Arab countries are dispersed over a large geographic surface estimated at 12.7 million km2 representing 10.2% of total world area, covering North Africa and a part of the Middle East, with increasing demographic growth, varying climate and economy and status of sanitation (Fig. 1). According to 2017 UN

© The Author(s), under exclusive license to Springer Nature Switzerland AG 2023 E. Debik et al. (eds.), Wastewater Management and Technologies, Water and Wastewater Management, https://doi.org/10.1007/978-3-031-36298-9_14

193

194

A. Dahchour and S. El Hajjaji

Fig. 1 Map of Arab countries

estimation, the population was about 415 million inhabitants [1]. The region is facing a population-expanding rate of 2.6%, associated with increased economic growth, urbanization, industrialization, and an increase in food demand. To tackle these needs, efforts are deployed to improve the agricultural production and the quality of life of the population. Total irrigated agriculture lands represent 17% of the world’s total arable land and generates about 34% of total agricultural production globally. Nevertheless, Arab countries import more than 50% of the food requirements [2]. These efforts tend to overexploit freshwater resources that are already affected by climate changes, the pollution by different pollutants, and the intrusion of seawater. Rainfalls vary from 150 to 600 mm annually, representing 2.6% of the world precipitation and 0.3% of renewable water resources. In terms of ratio per capita, the average for Arab countries is 512 m3/cap⋅year against 7243 m3/cap⋅year worldwide. This ratio varies from 172 m3/cap⋅year to 1278 m3/cap⋅year from Arabic peninsula to North Africa, respectively. However, some countries depend on their renewable resources on water from their

neighbors such as Kuwait (100%), Egypt (96%), and Syria (80%). This makes the region as the most water-scarce region of the world [2, 3]. The scarcity is aggravated by the low levels of the quality, availability, usage, and sanitation in Arab countries, where exploitation is mainly focused on groundwater such as is the case in Palestine, Libya, and Saudi Arabia, where the exploitation level could reach 100% [1]. The use of treated wastewater is being considered as an alternative resource. In this respect, only few countries like Lebanon have sufficient renewable water sources; most of other countries consider wastewater reuse as a priority issue. About 10.9 BCM (billion m3) of treated wastewater (TWW) was produced in 2013, 50% of which is used in different sectors and purposes [2]. The prevailing scarcity in most Arab countries has forced authorities to find alternatives for freshwater resources. Focus has been made by oil-rich countries on desalinization of seawater and brackish water. Wastewater reuse varies within the Arab countries. It could reach 100% of the available/collected wastewater like in Jordan or Abu Dhabi. The level and the quality of

Management of Water and Wastewater in Morocco and Arab Countries

treatment rely on performance of the existing sanitation infrastructure [2]. Though information on the irrigated land with wastewater and treated wastewater are scarce, some publications estimated irrigated area with wastewater or polluted water around 4–6 million ha [4]. Most of the technologies applied in Arabic world for wastewater treatment are based on primary, secondary, and tertiary processes. The applied technology is based on the availability of lands, financial budget, and the technical capacities for maintenance.

2

Water and Wastewater Management in Morocco

2.1 Geographical Location Morocco is located mainly in semi-arid to arid region on North Africa. Historically, repetitive waves of drought and water-scarcity events occurred in Morocco with negative impacts on its socio-economic development. This has oriented decision makers to adopt the national strategy for securing water supply of this vital commodity. Population of Morocco has culminated at 35 million inhabitants in 2018, distributed on 710,850 km2 (Fig. 2).

2.2 Potential of Water in Morocco Morocco has an important network of rivers providing 17 BCM surface water and about 32 deep aquifers and more than 46 inventoried shallow resources totalizing 5 BCM groundwater [5]. Rainfall varies between the north (up to 1500 mm/year) and the south (100 mm/year). Rainfall records indicate a decreasing tendency since 1970s (Fig. 3) [6]. Mean ratio per capita has decreased from 1000 m3/(cap⋅year) in 1960 to 550 m3/cap⋅year in 2016. It is expected to become below 500 m3/(cap⋅year) by 2030, as it is shown in Fig. 4. This mean figure hides the dramatic repartition between the north (1353 m3/(cap⋅year)) and south (362 m3/(cap⋅year) [7, 8]. Considering

195

this situation, a national strategy of sustainable management of water resources has been assessed. It aimed to increase the water storage capacity. Morocco totals about 149 dams with a storage capacity of 17 BCM. This potential is regulated by the law 36–15. Management is supervised by 10 basin agencies, 3 private associations for distribution of drinking water, and 9 OMRVA (Office de Mise en Valeur Agricola) for irrigation. Deep structuration of the environmental field has generated adoption of the National Act of the Environment Law 99–12 that is implemented through different strategies including strategy of the protection of environment and sustainable development (SNPEDD), national action plan for the environment (PANE), national plan to control climate changes (PNCC), and national plan of waste management (PNDM). In general, national strategies aimed at achieving the following goals: • • • •

Efficiency, Improvement of the supply, Protection of water resources, Reducing vulnerability to climate change by: – Regulatory and institutional reforms – Infrastructure and capacity building.

2.3 Potential of Wastewater Resources in Morocco Total wastewater produced in Morocco was estimated at 600 million m3 (MCM) with different loaded pollutants, 48% of which is discharged into the rivers or applied to the land, the remaining reaches the sea. Projections indicate that this amount would reach 900 MCM by 2020 mainly from agriculture and industrial activities (Fig. 5) [8]. This issue is tackled by the liquid sanitization program (PAN) that aims at: • Achievement of 80% of sewer connection in urban zone by 2020 • Achievement of 60% of treated wastewater by 2020.

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A. Dahchour and S. El Hajjaji

Fig. 2 Administrative map of Morocco

This is based on rehabilitation and extension of sewerage network and the building of wastewater treatment plant (WWTP) in 330 cities and urban centers totalizing more than 10 million inhabitants. Officially about 160 WWTP, which serve 3.38 million m2 of the country’s surface, are operated. Total of 45 MCM has been treated, and 23 MCM has been reused for industrial purposes and to irrigate green parks. The program tends to reuse 100 MCM each year. This will reduce the pollution generated by wastewater by more than 80% [9].

2.4 Quality of Raw Wastewaters The main pollutants in the raw municipal wastewater include organic pollutants (131,715 tons), nitrogen (42,131 tons), and phosphorus

(6.23 tons). These amounts depend on the size of urban center [10] (Table 1). Usage of wastewater in agriculture offers an interesting alternative to cope with shortness of conventional sources in some areas. Beside the efforts deployed by authorities through PNAL (national plan of liquid sanitization), tendency was expressed toward usage of wastewater in agriculture. Prediction tends to assess that the reused volume of wastewaters will not surpass 4.2% of water resources in Morocco by 2020. This volume cannot be totally mobilized for the following reasons [9]: • Lack of irrigable sites downstream of the discharges in numerous centers, mainly in coastal cities, • The high cost of the water conveyance that involves costs for pumping and channeling, • The availability of conventional waters.

Management of Water and Wastewater in Morocco and Arab Countries

Fig. 3 Geographic distribution of hydraulic basins in Morocco

Fig. 4 Annual ration of water in Morocco (in m3/cap⋅year)

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Fig. 5 Total volume of wastewater produced in Morocco

Though the raw wastewater has low quality, it is used illegally in some areas due to the closeness of the irrigated land to this source and the increasing demand for food by urban areas [2, 4]. Based on an average of 30–60 m3/cap⋅year of wastewater production, a population of 1 million inhabitants and a rate of irrigation of 10,000 m3/ha, the amount of wastewater produced would irrigate 3000–6000 ha of agricultural land [11]. On the one hand, TWW is reused in different places such as Marrakech’s Palm Grove, contributing to saving 33 MCM/year of freshwater and developing extensive golf courses. Other 20,000 ha is irrigated with TWW, mainly in the highlands for different crops. In the coastal areas, TWW serves to stabilize active sand dunes. On the other hand, illegal practices of wastewater reuse are noticed all over the country. About 7235 ha were irrigated with raw wastewater mainly in Marrakech (2000 ha), Meknes (1400 ha), and Oujda (1175 ha) regions [11]. To improve the quality of wastewater treated, Plan National d’Assainissement Liquide (PNAL) was launched in 2005. Ten million inhabitants were targeted in 338 urban centers within a period

of 10 years (2005–2015). It was aimed at improving sewerage collection (80% by 2020), treating of both industrial and domestic wastewater, reducing the pollution by 60% until 2020, and moving to its reuse in agriculture. This action would provide an alternative water source and protect the groundwater from overuse in some vulnerable areas such as Houz and Souss regions, where groundwater layer is predicted to last no longer than 12 to 16 years. However, despite of the existence of WWTP in several main cities, their functioning rate faces difficulties related to maintenance and spare part availability [9, 11].

3

Type of Wastewater Treatment Technologies

3.1 Physical and Chemical Treatment These include flotation, precipitation, oxidation, solvent extraction, evaporation, carbon adsorption, ion exchange, membrane filtration, and electrochemistry. In practice, a combination of several techniques is adopted [12]. Other techniques were introduced since 1950s for small urban wastewater treatment such as activated sludge, trickling filter, and biodisk. Generally, lack of maintenance and high cost of energy are the main constraints facing the functioning of these plants.

3.2 Biological Wastewater Treatment Biological treatment processes include aerobic and anaerobic treatment, phytoremediation of wastewater, which includes constructed wetland based on filtration, degradation, phyto-degradation, phytoaccumulation, phyto-transformation, and myco-

Table 1 Quality of wastewater discharged in different urban centers [9] Parameters

Small centers < 20,000 inhabitants

Average center 20,000– 100,000 inhabitants

Large city > 100,000 inhabitants

National average inhabitants

350

300

350

BOD5

400

COD

1000

950

850

900

TSS

500

400

300

400

Management of Water and Wastewater in Morocco and Arab Countries

remediation of wastewater. Operation parameters like pH, temperature, and the sludge residence time could influence the growth of the microorganisms and the clogging of the membranes in membrane bioreactors (MBRs) [12]. Activated sludge process: The process is based on aerobic treatment of wastewater in stirred bioreactors. It is a high air consumer to meet the oxygen requirement of microorganisms. Two stirred tanks are needed for carbon removal in the first one, and subsequently denitrification in the second one. The only constraint is the cost of air injection [12]. Marrakech WWTP uses activated sludge, and a tertiary treatment can treat up to 150,000 m3/day of wastewater and generate biogas [8]. Similar plants are operated in Agadir, recording good results of environmental parameters that are below the limits recommended by Moroccan standard (Fig. 6) [13]. Lagooning: The wastewater is allowed to settle in a consecutive artificial basin during a long residence time. The transfer from one basin to another is performed by gravitation. Pollutants are eliminated by populations of microorganism (bacteria, fungi, protozoa, metazoans, algae, fish, plants, etc.) [14]. This type of natural lagooning systems is more attractive for their economical cost and efficiency [15, 16]. In

Fig. 6 WWTP of Marrakech city

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Morocco, the most biotechnological wastewater treatment processes are adopted natural lagoons (Figs. 7). Morocco is adopting more WWTP based on biological treatment process. Such a project will be built in Tafoya at the south of Morocco; another plant will protect the lagoon and the beach of Moulay Bousselham, a fishing village and tourist destination on Morocco’s north coast, from discharged wastewater that will be used to recharge the water table.

4

Management of Water and Wastewater in Arab Countries

4.1 Water Resources and Situation of Water Scarcity Arab countries account to 6% of world population but only 1.3% of world’s renewable fresh water. 90% of the land is in arid and semi-arid zones. Precipitation range is between 150 and 600 mm/year, with seven Arab countries among the top 10 water scarce countries globally. Due to limited resources, overexploitation and pollution of water resources occur. Scientific predictions assess for 2050, the region is expected to experience 12% decrease of water supplies [3].

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Fig. 7 Lagoon WWTP

Annual water amounts in Arab countries tend to decrease and could be under 1000 m3/cap⋅year that represents the scarcity indicator. For instance, in Egypt, available water resources were 860 m3/cap in 2000 and were expected to decrease to 720 m3/cap by the year 2017. Beside of this, the surface of main lakes in Egypt has decreased dramatically, as, e.g., for Lake Manzala and Lake Edko. Their surfaces areas have decreased by 80% and 50%, respectively [2]. Remediation measures to water crisis could be insured by non-conventional resources such as seawater and brackish water desalination, and reclaimed effluents of wastewater [17].

Annual treated wastewater across the Arab countries ranges from less than 2% to about 98%; the average is 3–10% annually. The predicted produced wastewater in the region is about 36 BCM/year by 2030 [2]. Treatment of wastewater differs within Arab countries (Table 2). An average of 60% of the household wastewater is treated, while 40% of wastewater is discharged with little or no treatment. The major treatment consists of the stabilization ponds technology because of the availability of sufficient land in Egypt, Yemen, Morocco, and Syria. Other technologies (e.g., activated sludge) are preferred in other countries, while quality control is still not generalized [18].

4.2 Potential of Water and Wastewater in Arab Countries

4.3 Reuse of Wastewater in Arab Countries

The estimated amount of wastewater produced in the region ranged from 30 m3/cap⋅year to 90 m3/cap⋅year. In Arab countries, the total estimated wastewater volume was about 12.2 BCM/year in 2013 [18]. A total of 6.13 of BCM/year (51%) is discharged untreated, the remaining 49% of the total generated wastewater volumes are treated, and 77% of treated wastewater is re-utilized (Table 2) [2].

The reuse of wastewater for irrigation is recently being considered as a technical solution to alleviate the pressure on freshwater resources. National regulation limits tend to prevent adverse effects of treated or untreated wastewater on soil and plants [2]. Different options are adopted depending on environmental and economic objectives of the country. The chosen option is dictated by the water quality needed.

Management of Water and Wastewater in Morocco and Arab Countries

201

Table 2 Wastewater (WW) production in the Arab region [2] Country

Produced WW/year Million m3

Treated WW/year Million m3

Reused WW/year Million m3

Algeria

820

326

132

Bahrain

73

62

16.5

Egypt

3760

2971

2971

Iraq

450

230

120

Jordan

82

107

83

Kuwait

244

250

78

Lebanon

310

4.0

2.0

Libya

546

40

40

Mauritania

7.5

0.7

0.7

Morocco

650

40

0.1

Oman

90

37

37

Tunisia

187

–

21

Qatar

55

–

43

Syria

1364

–

215

Saudi Arabia

730

547

58

UAE

650

289

550

Yemen

74

46

6.0

In most countries, WWTP effluents are allowed for restricted agriculture as for the irrigation of trees, forage, and green spaces. In some countries, these regulations are ignored. The main usage of TWW is oriented to the main sectors consuming fresh water that could be supplemented. Targeted sectors are [2]: • Agriculture: TWW could help meeting agricultural demand for expanding populations. This usage remains very limited compared with other regions. The acceptance of TWW by farmers is another challenge to face. In contrary, in some cases as in Tunisia and Morocco, farmers would be ready to pay for TWW for irrigating their crops. • Environmental and recreational uses: This includes landscapes, park areas, fountains, dust control, fire protection, and constructed wetland in countries such as UAE, Kuwait, and Qatar.

• Industry: TWW could be reused for cooling system, boiler-feed water, wash down uses, road maintenance, and construction projects. • District cooling: This consumes a huge amount of fresh water in many Arab countries specially Gulf Cooperation Council (GCC) countries. • Groundwater aquifer recharge: Groundwater recharge could protect the aquifer systems from dryness [18].

5

Situation in Selected Arab Countries

Egypt has a surface area of 1 million km2 and an estimated population of 94 million in 2017, 95% of which lives in about 5% of the country land along the Nile River. The weather is dry in summer and warm in winter; the average rainfall is 50 mm/year [18].

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Nile River provides a total of 55.5 BCM of freshwater annually. Additional water resources including non-conventional water total to 22 BCM/year in 2015. Another 1.3 BCM comes from water recycling and about 7 BCM from groundwater [19]. Potential renewable water resources are estimated at around 57 BCM/year, 97% of which comes from the Nile River. Available water capacity is 1000 m3/cap⋅year, and it tends to decrease to 600 m3/cap⋅year by 2025. Water demand is estimated at 72 BCM/year, over 80% of which is used for agriculture [20]. Total wastewater collected during 2014–2015 amounted to 5.05 BCM, 74.4% of which is treated. The primary, secondary, and tertiary treatment processes represented 16.8%, 81.4%, and 1.8%, respectively [20]. TWW is used in farming of rice and aquaculture and to recharge the aquifer to stop seawater intrusion [2]. One of the main uses of raw wastewater or TWW is wood production. Afforestation and green belt programs have been allocated 2.4 BCM. Furthermore, evaluations of the safe use of TWW for irrigation of different crops such as jatropha, jojoba, sorghum, flax, and flowers have been conducted in Luxor [18]. Jordan has a surface area of 88,945 km2, 9% of which receives more than 200 mm of rain annually. Total annual surface water resources amount to approximately 400 MCM, while groundwater resources total to 500 MCM. Part of this resources (21%) comes from transboundary waters [2]. The average annual water supply decreased from 3600 m3/cap in 1946 to less than 145 m3/cap in 2008 [21]. Reported survey on water usage indicated that on average 22% is for municipal use, 3% for industrial use, and 75% for agriculture use [22]. The overexploitation of freshwater is partially compensated by groundwater [19]. Wastewater treatment plants are operated with a total influent amount of 112 MCM and effluent of 86.5 MCM/year, 25% of this amount is lost during the treatment process. The amount of treated wastewater was expected to reach 250 MCM/year by 2020 to compensate about 30–40% of future water demand in the

A. Dahchour and S. El Hajjaji

agriculture [2, 22]. Other option for wastewater reuse consists of groundwater recharge to prevent sea water intrusion [18]. Tunisia receives on average annual rainfall of 230 mm/year representing 36 BCM. This potential varies between min. 11 BCM in dry season and max. 90 BCM for wet season. Surface water resources are estimated at 2700 MCM provided by the northern part of the country (82%). Groundwater resource is confined in 212 shallow aquifers with a capacity of 719 MCM as well as 267 MCM of capacity in deep aquifers. The increasing demand and the impact of climate change generate space and time deficit in water supply. The projected volumes for the year 2030 indicate a deep aquifer. The increasing demand and the impact of climate change real water stress deficit of 389 MCM that makes it necessary to reuse reclaimed water [23]. Tunisia has started usage of wastewater for agriculture since 1975. Even, over 80% of TWW is consumed by agricultural sector, due to increasing trend in water consumption in general, Tunisia would face the problem of water scarcity by 2025. The total of collected wastewater is around 247 MCM/year. TWW consists of 90% of household wastewater (5% of the mobilized water resource). In 2014, 108 WWTPs have treated around 240 MCM/year. In 2015, 60 MCM/year of the treated wastewater has been reused mainly by agriculture and tourism sectors. A total of 9746 ha of golf and green areas were irrigated with treated sewage water, and approximately 3000 m3/day are used by the industrial sector. The amount of reused wastewater is about 24% of the total collected. Part of this (53.8%) flowed indirectly into the water reserves, and 46.2% were used directly for irrigation, parks, and golf courses. The target objective is to achieve 50% of reuse by 2020. This will be insured by 38 new WWTPs (of which 9 are industry-oriented). In Saudi Arabia groundwater depletion threatened the unique agro-ecosystem of Al Hassa

Management of Water and Wastewater in Morocco and Arab Countries

Oasis. A program of collection, treatment, and transport of TWW from neighboring cities has contributed to irrigate about 16,000 ha. By 2016, the whole area would be irrigated with TWW. Treated wastewater amount is 547 MCM, while only 166 MCM is reused [2]. Desalination plants and groundwater stand for the main sources of freshwater in Saudi Arabia. The United Arab Emirates are in the southeastern part of the Arabian Peninsula, covering a surface area of about 77,700 km2, and the population is about 10 million. Rainfall ranges from 700 to 1480 MCM/year. The potential of existing conventional water resources M includes 125, 3, 22, 20, 109 MCM/year from seasonal floods, permanent springs, seasonal springs, Falaj, and aquifer recharge, respectively [24]. Non-conventional water resources include 475 MCM/year of desalinated water and 150 MCM of reclaimed water [25]. Statistics of different emirates covering the period from 2000 to 2006, reported that rainfall received has varied from 2.8 mm in 2001 to 153 mm in 2006. The estimated potential of surface water in the UAE was about 150 MCM/year [26]. Abu Dhabi Emirate currently uses TWW to recharge aquifers to reduce seawater intrusions. The production of TWW has registered an increase of 10% annually, and this is provided by 19 WWTPs in Abu Dhabi with a total capacity of 884 MCM/year [25, 26]. In Oman potential water resources consist of groundwater that constitute 96% of water supply. Potential needs of water are estimated at 1735 MCM, while potential renewable water resources are estimated as 1000 MCM/year. The ratio per capita amounts to 388 m3/(cap⋅year) nonconventional resources such as desalinated seawater (from 43 plants, providing 3% of water supply) and treated wastewater (which contributes 1% to water supply) are used. The number of WWTP totals to 262, with an overall capacity of 52,000 m3/day, over 50% of which are located around the capital, Muscat [2]. Annual volume of untreated wastewater is estimated at 296,000 m3/day. Municipal TWW was expected to reach 445,000 m3/day (162.5

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MCM/year) by the year 2010. Most of TWW (26 MCM/year) is reused for landscape and agricultural irrigation, while nearly 25% (7.2 MCM/year) is used for aquifer recharge, and 0.1 MCM/year is reused for industrial purposes. Kuwait has around 2.6 million population, concentrated mainly in five urban areas. The rapid growth of population had generated an increased water demand in various sectors. Due to the arid condition of the country, average annual rainfall is less than 125 mm, and there are no surface water sources. Desalination provides 95% of potable water. The other resource is brackish groundwater. Agriculture consumes 52% of water resources, while the remaining 48% is dedicated to the domestic usage. Groundwater provides 60% of irrigation water, while the remaining 40% comes from recycled water. About 90% of the urban population is connected to a sewerage system. Kuwait has practiced water reclamation and reuse for the past few decades to save freshwater supply. Four main municipal plants (Ardiya, Riqqa, Jahra, and Um Al Haimam) ensure treatment of wastewater that varies between 206 and 254 MCM/year [2]. Yemen has the actual consumption of wastewater that is equal to 30 MCM/year, used to stabilize active sand in the coastal areas and dunes. Potential TWW usage can save 7–10% of the water used to irrigate 20,000 ha mainly for forage and cereals crop production. Syrian Arab Republic is characterized by a semidesert or desert conditions. It covers 185,180 km2 with a population approaching 22 million inhabitants, most of them concentrated in the Euphrates valley. Potential of water relies on 7 river basins. This potential is experiencing deficit since the renewable water resources of 15,208 MCM/year are exceeded by a total water consumption of 17,669 MCM in 2004. Total wastewater produced in Syria amounted to 1200 MCM/year, of which on average 75% is collected. The amount of treated wastewater varies between 273 and 550 MCM/year, corresponding to a treatment rate between 23 and 46% of generated domestic and

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Specific strategies and policies related to TWW reuse are adopted. Standards and regulations are adopted to sustainable reuse of TWW, to build user and public acceptance and ensure safe production.

was improved in 2017 with an additional 1.7 BCM [29]. This increase remains below the expectations, as the Delta Region alone generates more than 2 BCM/year, mainly from Cairo and Alexandria. WWTPs have served nearly 55% of the population in towns and cities in 2008. There is a need to build more WWTPs [30]. Since 1984 law (12/1984), the use of wastewater is prohibited, unless it meets regulated limits. Some efforts are made for phytoremediation as in El-Bahw Drain near Mansoura in the Nile Delta that takes a total discharge of 4000 m3/day, including agricultural and domestic wastewaters. An engineered wetland was constructed before entering Lake Manzala. The facility treats 25,000 m3/day with 90% of BOD removal. TWW is used for raising healthy fish suitable for human consumption. These methods offered as cheaper alternatives to conventional treatment methods [31]. Absence of quality control and monitoring facilities for wastewater before and after treatment is among the constraints of wastewater reuse in Egypt [32]. The improvement of water quality faces other constraints such as lack of public awareness, reliance on manual systems, lack of coordination between different water and wastewater authorities, overstaffed unskilled labors, inadequate operation and maintenance plans, lack of accurate and reliable data, and incomprehensive master plan. Most of the facilities have overloaded, inefficient water and wastewater networks, small capacities, and lack of new technologies [33].

In Egypt about 5.05 BCM of wastewater was collected between 2014 and 2015. The total TWW represents 74.4% of the collected wastewater. About 1.3 BCM/year of wastewater comes from industrial effluents. About 3.5 BCM/year of municipal wastewater is discharged into the Nile River and the Mediterranean Sea. The number of WWTP was reported to be 372 municipals in 2012, dealing with an average of 10.1 MCM/day wastewater. In 2004, the amount of wastewater treated represented 35% of the total wastewater released into the Nile River as 3.8 BCM/year [22]. The capacity of treatment

In Jordan stabilization ponds are responsible for 88% of the treated effluents. Beside of this, activated sludge, conventional or extended aeration, and trickling filters are used. WWTPs provide effluent with standards of the World Health Organization (WHO) for the use of treated wastewater for irrigation. The WWTP Al-Samar is one of the largest waste stabilization pond systems in the world. The facility treats 76% of all wastewaters in the country. Simultaneously, other mechanical treatment plants and activated sludge plants are under construction for Amman city. Currently,

industrial wastewater. While more than 40 WWTPs are operating in Syria, only 3% of TWW is used for irrigation [27]. In Lebanon the annual mean rainfall fells between 600 and 900 mm along the cost and 1400 mm in the hills. Annual available resources are estimated at 2.7 BCM including 2.2 BCM as surface water and 500 MCM as groundwater. This was expected to be lower than the consumption needs by 2015 before severe shortage in 2030. According to the World Bank, an estimated 248 MCM/year of wastewater was generated in 2010. Data on the real estimation of wastewater produced is scarce. Based on daily generation that is about 120 L/cap and an increase rate of 1.5%, total production of wastewater is estimated at 240 L/cap⋅day by 2040. This gives an annual production of 305 MCM in 2015 and 642 MCM by 2040. A national emergency program was launched to tackle the issue of wastewater. Wastewater treatment plant is scheduled for cities of more than 10,000 inhabitants [28].

6

Technologies Used in Arab Countries

Management of Water and Wastewater in Morocco and Arab Countries

98 WWTPs with primary and secondary treatment stages are in operation. Tertiary treatment is only provided in 5 WWTPs [18]. As-Samra as the largest WWTP was built in 2008 with a mechanical and a secondary biological treatment. It covers 181 ha, a capacity of 68,000 m3/day, and serves 2.2 million inhabitants in Greater Amman Zarqa area. The treated wastewater follows to Zarqa River with a low pollution load when it reaches King Talal Dam [34]. In Saudi Arabia activated carbon and reverse osmosis were used to remove pollutants from wastewater after the traditional treatment. Usage of wastewater for irrigation purposes has contributed to reduce the freshwater demand by 70% [2]. Oman has approximately 262 wastewater treatment plants, with an overall capacity of 52,000 m3/day, and tertiary treatment type is applied for 15,000 m3/day of this capacity. The main WWTPs are Darsait in Muscat, Al Ansab, and Shatti al Qurm, where they provide treated effluents of 11,500 m3/day, 5400 m3/day, and 750 m3/day, respectively [2]. In Kuwait four main municipal treatment plants (Ardiya, Riqqa, Jahra, and Um Al Haimam) using secondary and tertiary processes with high levels of disinfection with an annual treatment capacity of over 255 MCM are in process. A new treatment plant at Sulaibiya has been started with a treatment capacity of 375,000 m3/day, and the current treatment capacity will be extended to 600,000 m3/day. Advanced treatment processes are applied, namely ultrafiltration and reverse osmosis (RO) to remove biological nutrients, residual pollutants, and dissolved solids, and to lower the salinity of the municipal wastewater.

7

Conclusion

Arab countries are under arid and semi-arid conditions. These prevailing conditions make them under alarming water stress and scarcity. Potential of water resources is decreasing

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steadily, and treated wastewater is adopted or under investigation to be adopted for different sectors such as agriculture, recreation areas, green parks, and cooling. Quality parameters are set in most of the Arab countries; however, only few have the capacity and means to meet these standards. Jordan can be considered as the most advanced country regarding quality control and safety schemes for reuse. However, this scheme is currently limited to the Jordan Valley and requires further national upscaling. Other countries like Egypt or Tunisia have set very strict reuse standards, limiting reuse to forestry, green spaces, and industrial crops.

References 1. Dabour, M. (2006). Water resources and their use in Arab countries. Journal of Economic Cooperation, 27(1), 1–38. 2. Dawoud, M. A. (2017). Treated wastewater reuse for food production in the in the Arab region. Water Council Journal, 8(1), 55. 3. Conference on: The Use of Treated Wastewater in the Agricultural Production in the Arab World: Current Status and Future Prospective 14–16 January 2014 Dubai, United Arab Emirates Prepared by: International Center for Biosaline Agriculture (ICBA) 4. Jiménez, B., & Asano, T. (2008). Water reclamation and reuse around the world. In B. Jiménez & T. Asano (Eds.), Water reuse: An international survey of current practice (p. 648). IWA Publishing. 5. Ait Kadi, M., & Abdeslam, Z. (2018). Integrated water resources management in Morocco. In World Water Council Editor Global Water Security Lessons Learnt and Long-Term Implications. Springer. 6. Choukr-Allah, R. (2011). Comparative study between Moroccan water strategies. In S. Junier, M. El Moujabber, G. Trisorio-Liuzzi, S. Tigrek, M. Serneguet, R. Choukr-Allah, M. Shatanawi, & R. Rodríguez (Eds.), Dialogues on Mediterranean water challenges: Rational water use, water price versus value and lessons learned from the European Water Framework Directive (pp. 181–188) Bari: CIHEAM. (Options Méditerranéennes: Série A. Séminaires Méditerranéens; n. 98). 7. Ouhssain, M. (2008). La gestion sociale de l’eau au Maroc de AZERF à la loi sur l’eau. Revue HTE, 2008, 141. 8. Mandi, L. (2000). Review of wastewater situation in Morocco. In I. Chorus, U. Ringelband, G. Schlag, & O. Schmoll (Eds.), Water (pp. 383–387). Water Series, IWA Publishing, London, UK.

206 9. Salama, Y., Chennaoui, M., Sylla, A., Mountadar, M., Rihani, M., & Assobhei, O. (2014). Review of wastewater treatment and reuse in The Morocco: aspects and perspectives. International Journal of Environment and Pollution Research, 2(1), 9–25. 10. Choukr-Allah, R., & Hamdy, A. (2005). Wastewater treatment and reuse as potential water resource for irrigation. In H. Atef (Ed.), The use of nonconventional water resources Proceedings of the international Workshop Alger (pp. 101–124) Algeria, 12–14 June 2005. Options Méditerranéennes, Séries A n. 66. 11. Choukr-Allah, R. (2005). Wastewater treatment and reuse in Morocco: Situation and perspectives. In: A. Hamdy, F. El Gamal, N. Lamaddalena, C. Bogliotti, & R. Guelloubi, (Eds.), Non-conventional Water Use- WASAMED Project, CIHEAM. 12. Pell, M., & Wörman, A. (2011). Biological wastewater treatment systems. Comprehensive Biotechnology (Second Edition), 6, 275–290. 13. Mansir, I., Oertlé, E., & Choukr-Allah, R. (2021). Evaluation of the Performance and Quality of Wastewater Treated by M’zar Plant in Agadir, Morocco. Water, 13, 954. https://doi.org/10.3390/ w13070954 14. Narayanann, M., & Narayan, V. (2019). Biological wastewater treatment and bioreactor design: A review. Sustainable Environment Research, 29, Article number: 33. 15. Seidl, M., & Mouchel, J. M. (2003). Valorisation des eaux usées par lagunage dans les pays en voie de développement. Centre d’enseignement et de recherche Eau Ville Environnement, centre conjoint de l’ENGREF, de l’ENPC et l’UPVM. 7p. 16. Ruochuan, G., & Heinz, G. S. (1995). Stratification dynamics in wastewater stabilization ponds. Water Research, 29(8), 1909–1923. 17. Vasel, J. L., & Jupsin, H. (2007). Etat de l’art et perspectives des techniques extensives d’épuration des eaux usées domestiques sous climat aride (pp. 57–72). Congrès international Eau et Déchets. Université Mohammed Premier. 18. Toshio Sato, T., Qadir, M., Yamamoto, S., Endo, T., & Zahoor, A. (2013). Global, regional, and country level need for data on wastewater generation, treatment, and use. Agricultural Water Management, 130, 1–13. 19. ACWUA Working Group on Wastewater Reuse, March 2010. Wastewater Reuse in Arab Countries Comparative Compilation of Information and Reference ACWUA – Wastewater Reuse in Arab Countries List. 20. Abdelhamid, H. F., Abdel Daem, M. M., El-Hohary, E. H., & Abou Elnaga, Z. (2017). Safe reuse of treated wastewater for agriculture in Egypt. In S. Elhajjaji, A. Dahchour, & N. Dichtl (Eds.), Water perspectives in emerging countries. Water use in MENA countries 2017, November 03-08-Marrakech, Morocco.

A. Dahchour and S. El Hajjaji 21. Fawzi, K., & Boufaroua, M. (2014). Strategies to augment agricultural water shortages: Case studies of grey and treated waste water from Jordan and Egypt. The Use of TWW in the Agricultural Production in the Arab World: Current Status and Future Prospective. United Arab Emirates, Dubai. 15 Jan. 2014 (Presentation). 22. Shaalan, N. S. (2001). Egypt country paper on wastewater reuse. In Joint FAO/WHO Consultation for Launching the Regional Network on Wastewater Reuse, Amman, Jordan. 23. Benabdallah, S. (2007). The water resources and water management Regimes in Tunisia. In National Academies of Sciences, Engineering, and Medicine. Agricultural Water Management: Proceedings of a Workshop in Tunisia Washington, DC: The National Academies Press. https://doi.org/10.17226/ 11880 24. Zein, S., & Aboulrahmane, S. (2003). Water resources in the United Arab Emirates. Developments in Water Science, 50, 245–264. 25. Al-Rashed, M., & Sherif, M. M. (2000). Water resources in the GCC countries: An overview. Water Resour Manag., 14, 59–75. 26. Murad, A. (2010). An overview of conventional and non-conventional water resources in Arid Region: Assessment and constrains of the United Arab Emirates (UAE). Journal Water Resource and Protection, 2, 181–190. 27. Syria National Report on The actual statues of wastewater and Reuse for agriculture in Syria AHT GROUP AG, RG/2008-01/FTF - Identification and removal of bottlenecks for extended use of wastewater for irrigation or for other purposes, Draft Country Report , 2008, Eurpean Investment Bank. 28. Karam, F., Mouneimne, A., El-Ali, F., Mordovanaki, G., & Rouphael, Y. (2013). Wastewater management and reuse in Lebanon. Journal of Applied Sciences Research, 9(4), 2868–2879. 29. Tawfic Ahmed, M. (2008). Wastewater, challenges and opportunities—An Egyptian Perspective. In INNOVA– MED workshop, April 2008, Agadir, Morocco. 30. Qasim, S. R., Lim, S. W. D., Motley, E. M., & Heung, K. G. (1992). Estimating costs for treatment plant construction. Journal of the American Water Works Association, 84(8), 56–62. 31. Nasr, M., & Zahran, H. F. (2015). Assessment of agricultural drainage water quality for safe reuse in irrigation applications—A case study in Borg ElArab Alexandria. Journal Coastal Life Med, 3(3), 241–244. 32. Hassan, A. A. (2013). Wastewater treatment and projection of the northern lakes of Egypt. DAADEXCEED Regional Workshop on Wastewater Treatment and Reuse, 3–6 June 2013, Konya, Turkey. 33. Nazih, A. M. (2014). Wastewater operation and maintenance in Egypt (Specific Challenges and Current Responses). International Journal of Sciences: Basic and Applied Research, 18(2), 125–142.

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New Concerns on Treatment Technology

Paradigm Shift in Domestic Wastewater Treatment: Toward Energy Minimization, Greenhouse Gas Emission Reduction, and Resources Recovery Bilge Alpaslan Kocamemi, Sümeyye Çelik, Abdullah Bugra Senol, Halil Kurt, and Esra Erken Novel wastewater treatment technologies operating with a minimum energy requirement, low sludge production, and insignificant greenhouse gas emissions are discussed together with potentially recoverable compounds from these processes. The benefits of using molecular techniques under the new context of biological treatment is explained. Finally, integration of nanobiotechnology offering game changing breakthroughs in both treatment and recovery schemes is discussed.

Abstract

Water–Energy–Food (WEF) Nexus has gained a tremendous momentum to ensure optimal global use of water, energy, and food resources under changing climate conditions. Agenda 2030 for Sustainable Development with 17 Goals (SDGs) together with Paris Agreement have caused a paradigm shift in the wastewater treatment concept. This new concept favors innovative technologies allowing to capture organic matter from wastewater, while handling nutrients with a minimum of energy and greenhouse gas emissions and achieving a maximum recovery of all value-added compounds. This chapter aims to describe the paradigm shift in wastewater management linked to WEF Nexus Perspective. First, the major milestones in wastewater treatment and WEF Nexus are described.

B. Alpaslan Kocamemi (&)
S. Çelik
A. B. Senol
E. Erken Environmental Engineering Department, Marmara University, Istanbul, Turkey e-mail: [email protected] S. Çelik
A. B. Senol Kuzu Group Technopark Branch, Marmara University Istanbul, Istanbul, Turkey H. Kurt Department of Medical Biology, International Faculty of Medicine, University of Health Sciences, Istanbul, Turkey

Keywords




Biological wastewater treatment Energy minimization Novel technologies Resources recovery Sewage SDGs Water–Energy–Food Nexus




1













Introduction

Wastewater treatment history (Fig. 1) can be viewed under six major periods: (i) Early history, (ii) Roman period, (iii) Sanitary Dark age, (iv) Industrial age, (v) Environmental standards age, and (vi) Energy Minimization and Recovery Age. In early historic times, until first civilization, waste management was carried out by burying the wastes in the ground near the settlements [1]. Mesopotamian Empire (3500–2500 BC), which is known as the first advanced civilization in the history, used the first drainage system to remove

© The Author(s), under exclusive license to Springer Nature Switzerland AG 2023 E. Debik et al. (eds.), Wastewater Management and Technologies, Water and Wastewater Management, https://doi.org/10.1007/978-3-031-36298-9_15

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Fig. 1 History of domestic wastewater treatment

wastewater from the city [2]. In Indus civilization period (early 2500 BC), first wastewater treatment system was discovered in Rakhigarhi [3]. The houses in the city were connected to a sewer system. They used sumps between drainage pipes to remove settleable matter from wastewater [4]. In the ancient Greek civilizations (300 BC to 500 AD), wastewater and storm water pipelines in the cities were separated and connected to larger open channels that reached to storage tanks outside the city. The collected water was used for irrigation [5]. The Romans are known as the first civilization to adopt an integrated wastewater system including wastewater collection and sewer systems for the disposal of wastewater and storm water. Furthermore, the Romans reused the wastewater from baths/spas centers in the public toilets. Roman sewer systems were complex. They used manholes in the cities for the maintenance of pipelines [6]. The first physical adsorption method for wastewater treatment was used in Pompei by using porous subsoil of lava [7]. In the Middle age, starting with collapse of the Roman Empire, until nineteenth century, cities did not have wastewater distribution lines. Wastewater was mostly collected in cesspools/ cesspits and/or directly discharged into surface water storages. Inadequate wastewater management polluted the drinking water sources (surface

and underground), and this caused public health problems periodically in medieval cities [7]. With the advent of the industrial age, mankind realized the necessity of a proper waste disposal. Sewer systems were constructed, and wastewater was collected in the cities. However, collected wastewater was pumped into rivers or the irrigation fields far from the cities in Europe and the USA. The Eighth Report (1912) of the Royal Commission on Sewage Disposal is a milestone for wastewater treatment. In this report, biochemical oxygen demand (BOD), established standards, and tests to be applied to sewage and sewage effluents were declared. Before First World War (WWI), countries started to mandate wastewater treatment, but WWI delayed the construction of treatment facilities in Europe [8, 9]. The Second World War also led to delays in wastewater treatment applications, and this resulted in increased water pollution [1]. By 1950, with increasing water pollution problems, water quality standards were set and new wastewater treatment processes, especially biological ones, were developed. Wastewater Management (Fig. 2) started with physical primary treatment techniques. The oldest method is the sedimentation in trenches and pits. These were replaced with sedimentation tanks in ancient Roman civilizations [10]. In the Industrial Age after the industrial revolution

Paradigm Shift in Domestic Wastewater Treatment …

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Fig. 2 Milestones in domestic wastewater treatment

(Fig. 1), wastewater treatment activities gained importance. In 1860, L.H. Mouras designed cesspit “Fosses Mouras” [1]. This was followed by septic tank design by Donald Cameron in 1895 [1] and Imhoff tank design by Karl Imhoff in 1906 [11]. Secondary biological treatment was officially started in 1893 in England with trickling filters [1]. Later, Rotating Biological Contactors (RBCs) were invented in 1916 [10]. The Eighth Report of the Royal Commission on Sewage Disposal is a milestone in wastewater treatment opening a new age of stringent environmental standards to protect water bodies (Figs. 1 and 2). In 1913, Edward Arden and Lockett [1, 12] invented suspended growthactivated sludge process. Biological nitrification and denitrification processes were discovered in the years 1877 and 1885 [13]. The first attempts for biological nitrification and denitrification processes were done in 1962 by Wuhrmann as a post-anoxic configuration and by Lutdzack Ettinger as a pre-anoxic configuration [1, 14, 15]. In 1970, first patented biological phosphorus removal process, Phostrip, was investigated [15]. In the same period, membrane treatment

technologies had started to receive attention and the use of membrane bioreactors started in fullscale treatment plants, initially in the USA, Japan, and Europe [1]. In late 1970s, with the onset of the energy crisis, aerobic wastewater treatment had started to shift anaerobic ones. In 1980, upflow anaerobic sludge blanket reactors (UASB) were developed [15]. In 1972, Clean Water Act was introduced in the USA, and the secondary treatment became mandatory for wastewater treatment and nutrient removal from wastewater that gained more importance. In the year 1973, James Barnard modified this process to increase the efficiency of nitrogen removal [1, 15]. The first combined biological nitrogen and phosphorous removal in a single sludge system (A/O process) was developed by James Barnard in 1974 and patented under the name Phoredox in 1975 [14, 15]. In 1975, Nicholls applied enhanced biological phosphorous removal at Johannesburg’s Alexandra Wastewater Treatment Plant (WWTP) [14, 16]. Later, in 1975, Barnard developed the A2/O and Bardenpho (BARnard Denitrification and PHOsphorus) removal processes [15, 17]. The

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first Bardenpho process was built in Palmetto Florida, USA, in 1979 [18]. Besides suspended-growth processes, in 1990s biofilm configurations, which are efficient for slow biological processes (e.g., nitrification) with less sludge generation were applied in WWTPs [19]. In 1994, Odegard et al. invented a new type of attached growth process called as moving bed biofilm reactor (MBBR) system. This new biofilm configuration system has received great attention in the application of novel nitrogen removal technologies especially in cold climate countries [20]. In 1990s, with the development of the Water– Energy–Food (WEF) Nexus concept [21] a new wastewater treatment age called as energy minimization and recovery age has been started. In the same period, development of molecular techniques accelerated the development of novel technologies. Among those new technologies, ANaerobic AMMonia OXidation (Anammox) process, discovered in 1995 [43], is a milestone in the start of a new perspective in nitrogen removal from centrate and mainstream. Later, various novel technologies (e.g., Denitrifying Ammonium Methane Oxidation (DAMO), Complete Ammonium Oxidation (Comammox), and Dissimilatory Nitrate Reduction to Ammonium (DNRA)) were established between 2006 and 2021. In the context of Water–Energy, and Food (WEF) Security Nexus, with the discovery of these new processes, a paradigm shift and a new period in wastewater treatment aiming at energy-neutral or energy-positive treatment together with resource recovery approaches has occurred in wastewater treatment.

2

Water–Energy–Food (WEF) Nexus Perspective on Domestic Wastewater Management

Due to global rapid population and economic growth in combination with accelerated urbanization and changing lifestyles, the demand for water, energy, and food is constantly increasing under limited natural resources conditions. In 2011, the Federal German Government

organized an international conference known as “Bonn 2011: The Water, Energy, and Food (WEF) Security Nexus—Solutions for the Green Economy” to develop new solutions [22]. WEF approach is based on an understanding of the synergies and regulated negotiation of fair tradeoffs between competing uses of water, land, and energy-related resources [21]. It contributes not only to sustainable development but also to the United Nations (UN) overall developmental and global commitments, namely the Agenda 2030 (Fig. 3) with its Sustainable Development Goals (SDGs). In September 2015, the United Nations adopted the Agenda 2030 for Sustainable Development that includes 17 Goals (SDGs), and this was followed by Paris Agreement on climate change in April 2016 (UN, 2015, 2016). 17 SDGs are integrated with social, economic, and environmental perspectives [23]. Among SDGs, especially SDG 6 “Clean water and sanitation,” SDG 7 “Affordable and clean energy,” SDG 11 “Sustainable cities and communities,” and SDG 13 “Climate actions” are related to sustainable wastewater treatment in some points [24]. In the year 2019, European Green Deal was presented by the European Commission, committing to climate neutrality by 2050. Green Deal provides a roadmap to boost the efficient use of resources by moving to a clean, circular economy and stopping climate change, reverting biodiversity loss, and achieving net-zero greenhouse gas emissions. It outlines the investments needed and financial tools available, and it explains how to ensure a just and inclusive transition [25]. The 26th UN Climate Change Conference of the Parties (COP 26) hosted in Glasgow, UK on October 31–November 13, 2021 [26] brought representatives of the 197 attending parties together to accelerate actions toward the goals of the Paris Agreement and the UN Framework Convention on Climate Change. In this context, wastewater has become a valuable resource in circular economy, and its energy self-sufficiency with minimized greenhouse gas emissions has received great attention. More effort has been dedicated to replacing high

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Fig. 3 Milestones in Water– Energy–Food (WEF) Nexus perspective (GHG: German Federal Government, UN: United Nations, EU: European Union, EC: European Commission)

energy consuming, large amounts of waste sludge and greenhouse gas generating conventional activated sludge (CAS)-based biological nutrient removal (BNR) facilities with emerging and innovative technologies allowing to capture organic matter from wastewater while handling nutrients with a minimum energy and maximum recovery approach. Under this new context, WWTPs have been started to call as resource recovery facilities. This paradigm shift of energynegative to energy-neutral or positive operation of WWTPs has changed wastewater treatment concept as discussed in the following sections.

3

New Insights on Nutrient Removal

Domestic wastewater is mainly comprised of water together with suspended solids, organic matter, nitrogen, and phosphorus. Organic matter in domestic wastewater is commonly quantified with chemical oxygen demand (COD) and fractionated as biodegradable and non-biodegradable COD. Conventional biological WWTPs (Fig. 4a) are mainly designed to remove not only readily biodegradable fraction of COD (rbCOD), but also slowly biodegradable ones, i.e., particulate

COD (sbCOD) by the application of aerobic heterotrophic biological process. As being heterotrophic, the biological removal of rbCOD and sbCOD generates huge amount of waste activated sludge together with greenhouse gas (CO2) emissions and hence prevents recovery of organic matter as energy source. Nitrogen in domestic wastewater, which exists in the form of organic nitrogen and ammonium, is converted to N2 gas by the application of conventional nitrification and denitrification technologies (Fig. 4a). Nitrification process, which oxidizes ammonium to nitrate in two steps by aerobic autotrophic ammonium and nitrite oxidizing species, consumes significant amount of energy for oxygen supply, which is necessary for the oxidation process. Following the nitrification process, denitrification process reduces nitrate to N2 gas by facultative heterotrophic bacteria by wasting valuable organic matter as sludge and generating thereby the potent greenhouse gas nitrous oxide (N2O). In conventional WWTPs (Fig. 4a), both organic matter and nitrogen are efficiently removed with the well-proven technologies as explained above. However, in the view of the Green Deal, wastewater has become a valuable resource in the circular economy and its energy

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Fig. 4 Paradigm shift in domestic wastewater treatment: a conventional approach and b novel approach

self-sufficiency with minimized greenhouse gas emissions. Development of more sustainable and circular economy-based wastewater treatment systems has received increasing attention. In this context, research efforts have been dedicated to innovation and integration of novel technologies to conventional treatment plants. The main idea behind these novel technologies is (i) to enhance organic carbon capture from wastewater to be diverted to energy recovery processes (e.g., digestion), (ii) to minimize energy consumption due to oxygen supply, (iii) to decrease operational costs due to excess sludge treatment and disposal, (iv) to eliminate greenhouse gas emissions, and (v) to recover and to valorize valueadded compounds. In this framework, a paradigm shift has been occurring in domestic wastewater treatment in such a way that high energy consuming, large amounts of waste sludge and greenhouse gas generating conventional activated sludge processes (Fig. 4a) will be replaced by emerging and innovative technologies (Fig. 4b) allowing to capture organic matter from wastewater, while handling nutrients with a minimum energy and maximum recovery. Novel

nitrogen/nutrient removal technologies such as high-rate activated sludge (HRAS) for A-stage (Fig. 4b), Anaerobic Ammonium Oxidation (Anammox), Complete Ammonium Oxidation (Comammox), and Dissimilatory Nitrate Reduction by Anammox (DNRA) processes for BStage (Fig. 4b) have gained great attention under this new context.

3.1 High-Rate Activated Sludge (HRAS) Process (A Process) The high-rate activated sludge (HRAS) process, which is also called as A (Adsorption) process in wastewater sector, is an old technology discovered by Busswell and Long in 1923 [27]. The process removes and redirects organics from wastewater to energy generating anaerobic digestion or incineration process in an efficient manner [28]. It is usually operated under short hydraulic retention time (HRT; 0.25–4 h) and sludge retention time (SRT; 0.1–4 d) conditions [28–30]. The main mechanism of the process is based on the biological adsorption, also called as

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biosorption [31]. The mechanism of the biosorption could be explained in three processes (Fig. 5) [32]: (i) bio-flocculation step, both particulate and soluble organic compounds are physicochemically adhered to the floc, (ii) absorption step, compounds absorbed into the cell are later used for anabolism and catabolism processes, (iii) hydrolysis step, adsorbed compounds are hydrolyzed to smaller soluble compounds by enzymes or extracellular polymeric substances (EPS). The HRAS systems can be designed and operated under three different strategies: (i) to meet secondary effluent standards (e.g., 30 mg/L BOD5 and 30 mg/L TSS limits) without aiming at nitrogen removal, (ii) to provide capture of organic matter prior to second stage (B stage) to be designed for nitrogen removal with conventional nitrification/denitrification processes, and (iii) to provide a maximum capture of organic matter prior to second stage (B stage) to be designed for nitrogen removal with novel autotrophic nitrogen removal processes (e.g., Anammox). The last two operation strategies are called as A-stage and operate notably different than the first one under very low SRT, HRT, and dissolved oxygen (DO) values (< 1 day SRT, * 30 min HRT, < 1 mgO2/L DO). Although A-stage HRAS separation can be adjusted theoretically

to remove only a portion of the organic carbon so that organic carbon requirement of denitrification in B stage can still be met, C deficiency in B stage can be a problem in practice. Additionally, full/partial nitrification risk under first operational strategy (especially during summer months) should be considered in the design stage. Most of the existing A/B plants [33] were designed according to the second operation strategy explained above. However, the recent trend is shifting to the third strategy by integrating the Anammox process to mainstream treatment by overcoming known challenges (e.g., partial nitritation). All over the world, there are only two full-scale A/B processes including Anammox, which were retrofitted from the existing Changi WWTP in Singapore (Step Feed Activated Sludge SFAS) and Strass WWTP in Austria (pre-denitrification, nitrification) [34, 35].

Fig. 5 Mechanism of the biosorption concept in HRAS

3.2 Partial NitrificationDenitrification Via Nitrite The first attempts for energy minimization in conventional nitrification and denitrification applications were accomplished by evaluating shortcut nitrification–denitrification process (Fig. 6).

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Fig. 6 Partial nitrification–denitrification via nitrite

In this approach, NH4+–N is oxidized to NO2−–N but not to NO3−–N, and denitrification occurs via NO2−–N. This process allows not only 25% O2 saving with respect to conventional nitrification–denitrification process, but also reduces the external carbon addition approximately by 40%, necessary for the denitrification of wastewater containing low C/N ratio [37, 38]. Additionally, the sludge production in the partial nitrification–denitrification via nitrite process is about 30% less than that of conventional nitrification–denitrification process [39]. However, under mainstream conditions partial nitrification is difficult to achieve due to low NH4+ content of domestic wastewater. Growth rate of nitrite oxidizing bacteria (µNOB) is greater than the growth rate of ammonia oxidizing bacteria (µAOB) at temperatures below 25 °C, which is typical for domestic wastewater. Therefore, NO2− accumulation cannot be achieved, and NO2− is converted to NO3− instantaneously [40]. The difficulty of achieving steady and sustainable partial nitritation of sewage was demonstrated in various studies, even under pilot-plant conditions [41, 42].

3.3 Anaerobic Ammonium Oxidation (Anammox) Process The Anaerobic Ammonium Oxidation (Anammox) process, which has been discovered in late 1990’s [43], oxidizes ammonium to nitrogen gas using NO2− as an electron acceptor in the

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absence of dissolved oxygen (Fig. 7). The strictly anaerobic, autotrophic, distinct red-color Anammox bacteria are classified under the phylum Planctomycete and have seven identified species up to now, namely Brocadia, Kuenenia, Jettenia, Scalindua, Anammoxoglobus, Anammoximicroium, and Loosdrechtii [44–47]. As being an anaerobic chemolithoautotrophic biological conversion process, the anaerobic ammonium oxidation (Anammox) process, is a very cost-effective and hence attractive alternative to conventional nitrification and denitrification processes (Fig. 8). It allows saving oxygen and organic matter, minimizing sludge production, and preventing greenhouse gas emissions. However, the NH4+/NO2 ratio requirement of the Anammox stoichiometry [48] requires the use of partial nitritation (PN) process prior to Anammox. In partial nitritation, about half of the NH4+ must be oxidized to NO2, with a fraction of NH4+ remaining unconverted. In past studies, different strategies and approaches have been evaluated to achieve a stable PN process by adjusting dissolved oxygen (DO), temperature, sludge age, pH, and free ammonia (FA) [41, 42, 49–51]. However, for wastewater with low ammonium concentrations, e.g., sewage, it is very difficult to obtain stable and sustainable nitrite accumulation with the control of only one of these parameters. This is one of the major challenges of replacement Anammox process with conventional nitrification and denitrification process in domestic WWTPs. Today, Anammox process is still under investigation for mainstream applications [42, 50–52]; however, it is a well proven and applicable process for the treatment of sludge digester supernatant, which can be partially nitrified effectively by Sharon process due to its unique temperature and alkalinity characteristics [53]. The other challenges that limit mainstream Anammox applications are slow growth rate of Anammox bacteria and unavailability of local Anammox seed. These obstacles cause a long start-up period in newly started plants [54]. However, there are inspiring attempts all around the world for the enrichment of local Anammox seed from activated sludge [54]. In these

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Fig. 7 Anammox pathway in overall biological nitrogen removal processes

Greenhouse gases

Fig. 8 Conventional and novel nitrogen removal approaches in wastewater treatment correction

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attempts, real-time monitoring of process dynamics is quite important [55–57]. Integration of nanotechnology into Anammox applications (especially nanoscale zero valent iron nZVI particles) enhances quick Anammox enrichment activities from local activated sludges [58, 59]. The operational factors and wastewater constituents that can inhibit the slow growth rate of Anammox species are also under deep investigation [60, 61].

3.4 Deammonification Applications Deammonification is a common term describing partial nitritation followed by Anammox process. Deammonification systems can be designed as two-stage or one-stage systems (Fig. 9) using suspended-growth sequencing batch reactor (SBR) [62], granular upflow anaerobic sludge blanket (UASB) systems [63], attached growth moving bed biofilm reactor (MBBR) [64], or hybrid suspended and attached growth integrated fixed film activated sludge (IFAS) reactor [65] configurations (Fig. 10). As being a fully autotrophic process, deammonification process including partial nitritation followed by Anammox appeared as good candidate for “B stage” coming after “A-stage,” which enhances organic carbon capture from wastewater to be diverted to energy recovery processes

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(Sect. 3.1) [66, 67] However, as explained in Sect. 3.3, partial nitritation, which has to end up with ammonium to nitrite ratio of 1.3–1.5 for Anammox, is still a limiting factor for the application of deammonification on mainstream. Hence, resilient mainstream nitrite-centered deammonification systems are still not applicable in practice. This challenge can be overcome by a paradigm shift from nitrite-dependent deammonification systems to the systems that use versatile metabolism of Anammox bacteria and interactions and synergies among newly discovered low-abundant species like Anammox Ca. Loosdrechti (Sect. 3.5) and Comammox (Sect. 3.6) species. Among deammonification reactor configurations, the MBBR system, which was established in late 1980s [20], has received great attention due to its high treatment efficiency; low capital, operational, maintenance and replacement cost; single reliable and robust operation [68]. Contrary to most biofilm reactors, it is a noncloggable biofilm reactor with no need for backwashing, low head loss, and a high specific biofilm surface area [69]. Today, the number of full-scale deammonification MBBR systems is very limited, and the common approach to start a new deammonification MBBR system is to use pre-colonized media to shorten start-up period due to very slow growth rate of Anammox species [53, 70].

Fig. 9 Deammonification applications: a two stage, b one stage [52]

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Fig. 10 Deammonification reactor configurations: a suspended-growth, b Granular, c attached-growth [42, 50, 52]

However, pre-colonized media is not always available locally. Additionally, not all wastewater resource recovery facilities include digester supernatant although the current strategy for deammonification MBBR enrichment is mainly dependent on side stream. Hence, further research is necessary to develop an optimum Anammox enrichment strategy for deammonification MBBR systems without pre-seeding and feeding with mainstream wastewater. Being a biofilm process, the performance of MBBR is hardly predictable due to complex behavior of microbial consortia and various effects of physical and chemical factors in the bulk liquid and biofilm. The mass balance and substrate fluxes in MBBR systems are fundamentally different compared to conventional kinetics in activated sludge models. The substrate concentration varies through the entire thickness of the biofilm due to diffusion limitation through diffusion boundary layer and biofilm. Attachment/detachment of biofilm has also significant impacts on the process. The mathematical models can help fast forward the processes and put to the test hypotheses difficult to practically test by making so many calculations, interpolations, and extrapolations simultaneously. To accelerate the progress and to receive the results from the mathematical models, commercially available computer-based simulation programs can be used. The study of Uzkurt [71] is a good example

of demonstrating the benefits of using computerbased simulation models to identify the most optimal operation scheme favoring Anammox growth in one-stage mainstream Anammox systems.

3.5 Dissimilatory Nitrate Reduction to Ammonium (DNRA) It was recently discovered that an organotrophic Anammox genera (Loosdrechtii) can perform dissimilatory NO3− reduction to NH4+ (DNRA) with NO2− as an intermediate using volatile fatty acids (VFA) as electron donors (Fig. 7) [72]. With this discovery, DNRA has appeared to replace partial nitritation process, which is required prior to the Anammox process for nitrite generation and has many challenges in mainstream applications. DNRA can be integrated to Anammox process in two ways. First, with partial DNRA, nitrate produced through nitrification and Anammox processes can be reduced to nitrite, which can further be utilized as an electron acceptor in Anammox process. In the second pathway, full DNRA, nitrate can be reduced to ammonium, which can be utilized by Anammox bacteria [47]. DNRA and/or partial DNRA process schemes provide a significant supply of NH4+ and/or NO2− for the Anammox process. Additionally, the NO3− accumulation produced by NOB or by

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Anammox bacteria could be removed through DNRA-Anammox pathway, thus enhancing nitrogen removal [73]. Besides DNRA-Anammox applications, the use of DNRA process in combination with Comammox process (Sect. 3.6) appears as a cost-effective and efficient alternative for partial nitritation applications prior to Anammox process.

3.6 Complete Ammonium Oxidizers (Comammox) The recent advances in molecular techniques (Sect. 5) allow the discovery of new species and enhance biological diversity. Recently discovered complete ammonium oxidizers (Comammox) species “Candidatus Nitrospira inopinata,” which is known as a nitrite oxidizer, has completely changed the high oxygen requiring nitrification concept in wastewater treatment [74]. Comammox process (Fig. 11) is performed by only one type of bacterium in a single cell, which has its own Ammonia monooxygenase (AMO), hydroxylamine oxidoreductase (HAO), and nitrite oxidoreductase (NXR) enzymes, and hence, one organism performs the first process NO2−–N oxidation to NO3−–N by using NXR enzyme [75]. As being K-strategist (oligotroph), Comammox species have slow growth rate and high affinity for the substrate ammonium [76]. The existence of Comammox species in various environmental zones, e.g., in terrestrial habitats [77]; in rivers and lakes [78]; in WWTPs [74]; in drinking water systems [79], and in geothermal springs [74] was demonstrated by metagenomic analyses. Low dissolved oxygen (DO) conditions favor the growth of Comammox Nitrospira [80]. The number of studies evaluating the activity of Comammox species in domestic wastewater treatment is still quite limited. These studies were mainly focused on diversity of microbial population and abundance of Comammox species in this population without evaluating the nitrification process performance deeply. However, in the study of Senol et al. [81], enrichment of

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Comammox species from a mixture of initial seed materials from various environments, such as terrestrial environment (e.g., sediment, paddy field), activated sludge, and sand filter of drinking water plant was deeply investigated in a laboratory scale SBR reactor and a pilot scale integrated fixed film activated sludge (IFAS) type SBR reactor under limited DO conditions (< 0.5 mg O2/L) in terms of process efficiency, kinetics, and molecular diversity. The discovery of the Comammox bacteria appears as a promising alternative solution for both conventional denitrification and conventional Anammox process applications. For the latter case, Comammox should be applied together with another newly discovered technology DNRAS (Sect. 3.5).

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Wastewater-Based Resource Recovery

The paradigm shift in wastewater treatment is not only focusing on the use of energy-efficient new technologies, but also the recovery of water, energy, nutrients (N, P), and further recoverable materials from wastewater.

4.1 Reuse of Wastewater In view of the rapidly increasing shortage of water resources all around the world due to climate change, water reuse and reclamation have become one of the EU’s priorities fulfilling the ambitions of the European Green Deal to implement circular water management strategies [82]. In this context, there is a growing interest in the reuse of treated effluents from WWTPs. In 2021, global reuse of treated wastewater for agricultural purposes is expected to range between 1.5 and 6.6% [83]. The effluent of WWTPs can be reused for industrial applications, urban activities, groundwater re-charge, and most importantly for agriculture, which is the largest water consumer in the world, using some 70% of the World’s freshwater withdrawals [84]. The use of treated

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Fig. 11 Canonical two-step nitrification and Comammox [75]

wastewater for agricultural activities does not only provide source of water, but also of nutrients (e.g., nitrogen and phosphorous) required for plant growth. However, prior to use of wastewater for irrigation and/or agricultural activities, implementation of proper treatment processes is quite important to protect soil, crop, and human health. To meet the standards for restricted or unrestricted irrigation, secondary, tertiary, or quaternary treatment technologies can be used [85].

4.2 Reuse of Sludge in Land Applications Sewage sludge, also known as biosolids, contains valuable resources (N, P, K) with fertilizer value, and organic matter that improves the soil quality, acts as a soil conditioner [86]. Hence, the application of sewage sludge to agricultural land under strict regulations is gaining popularity. Direct land-spread of wastewater sludge is forbidden by law due to hazardous components of the untreated sludge (hazardous organic chemicals, heavy metals, pathogens, antibiotic resistant bacteria, etc.) [86, 87]. Many countries apply strict environmental standards (e.g., EU Sludge

Directive, 86/278/EC, USEPA Title 40 CFR 503) for sludge reuse in agricultural activities [86, 87]. In this context, sludge treatment facilities in WWTPs must be designed to produce “clean (non-toxic) sludge” that can be utilized in land applications and agriculture [88]. In WWTP, the most common way of treating sludge before using for land application is to apply anaerobic digestion, lime stabilization, thermal drying of raw sludge or thermal drying of digested sludge [88]. Based on USEPA 40 CFR Part 503, biosolids are divided into “Class A” and “Class B” designations, based on treatment methods [89]. The difference between Class A and Class B biosolids is solely based on pathogen concentrations. Class A sludge requires treatment processes like composting, heat drying, high-temperature aerobic digestion, or lime stabilization to reduce pathogens, including pathogenic bacteria, enteric viruses, and viable helminths ova below detectable levels, as set by USEPA 40 CFR Part 503 [90]. Class A biosolids can be used as fertilizer on farms, vegetable gardens, and sold to home gardeners as compost or fertilizer [89]. Unlike Class A, Class B sewage sludge treated with aerobic/anaerobic digestion and lime stabilization may still contain pathogens and hence cannot be

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used for land application at public contact sites (e.g., parks, golf courses, lawns) [91].

4.3 Energy Recovery from Sludge The increasing global energy demand (approx. 50% between 2010 and 2040) [92] increases attention on energy recovery from wastewater. In WWTP, the energy can be recovered through biogas production from sludge, heat recovery from thermal processes (e.g., sludge drying, incineration, pyrolysis, and gasification) and bioethanol production from sludge. In the context of Water–Energy–Food Nexus and Green Deal, second-generation (next generation) biofuels that are produced from non-food biomass like WWTP sludge have received great attention as an alternative to fossil fuels. Biogas that is generated from anaerobic digestion of primary and waste activated sludge of WWTP typically contains 55–65% methane and 35–44% carbon dioxide [93]. Methane, the primary component of biogas, can be used either to produce electricity and heat provided by Combined Heat and Power (i.e., CHP, co-generation) units [94, 95]. However, elevated concentrations of certain trace compounds present in biogas (e.g., sulfur compounds, siloxanes) can induce fatal damage to CHP system, especially corrosion damage [95, 96]. In WWTP, especially in developing countries, frequent failure of CHP systems may prevent efficient utilization of biogas for energy generation resulting in wasting of biogas by flaring [97]. Adsorption, absorption, and membrane-based gas separation processes have been widely used to remove sulfur and siloxane impurities from biogas [95]. In recent years, especially in megacities, management of sludge disposal generated from WWTP has become a major problem due to the limited availability of landfill sites. This increases the application of thermal sludge drying and sludge incineration technologies that offer over 90% reduction in sludge volume. Sludge drying processes are usually incorporated into cogeneration units to utilize the thermal energy

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generated by gas engines or gas turbines. The high volatile organic content and calorific content of dewatered/dried sewage sludge are at quite comparable levels of various biomass materials [98]. Hence, energy recovery from sludge incineration is an attractive route for eliminating its volatile organic matter while recovering energy. There are two different types of incinerators commercially available for biosolids incineration [99]: multiple hearth furnaces and fluidized bed furnaces. Fluidized bed furnaces are a newer technology that is more efficient, stable, and easier to operate than multiple hearth furnaces [99]. Today, fluidized bed incineration systems are designed in order to recover excess energy remaining after being used for internal combustion and steam production [100]. The produced electricity in these incineration plants via steam turbine is utilized for the internal power consumption of the incineration plant, and the remaining part is used for the WWTP [100]. Additionally, the end solid product of incineration, ash, can be reused in the manufacture of sintered materials (bricks, tile sand pavers, and lightweight aggregates), glass– ceramics, lightweight aerated cementitious materials, and regular cementitious materials [86]. Additionally, phosphate in ash can be recovered as a high value technical grade phosphoric acid using optimized sulfuric acid leaching process [86]. Bioethanol is a second source of renewable energy that can be produced from the sewage hydrolysate through fermentation [101]. As opposed to other alternative fuels (e.g., hydrogen), bioethanol integration to existing engine systems (5–85% mixture with gasoline) does not require any modification of the engines [102]. The separation of bioethanol from water is conducted by distillation, pervaporation separation, vacuum fermentation, adsorption, gas stripping, solvent extraction, and other alternative hybrid processes [103]. However, efficient separation and purification of bioethanol, which accounts for 40–80% of total process cost, is a major challenge in the practical application of the process [104].

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4.4 Phosphorus Recovery Phosphorus (P), which is a key component of DNA and ATP, is a vital element for all organisms [105]. However, P is a finite (limited) element, which occurs predominantly in phosphate rocks that are not evenly distributed across the earth and mineable with generation of huge quantities of wastes. However, human and animal excreta include plenty amount of phosphorus due to their dietary. Hence, in recent years, wastewater appears as an attractive source for the recovery of phosphorus. However, wastewater has a diluted phosphorus content generally at concentrations below 10 mg/L, which requires the recovery of wastewater from sludge including higher concentrated form of phosphorus especially in bacterial aggregates. Recovery of phosphorus from P-rich sludges can be accomplished in four different ways [106]: (i) direct use of stabilized P-rich sludge as a fertilizer, (ii) use of thermal treated P-rich sludge as a fertilizer, (iv) phosphorus recovery from sludge incineration ash, and (iv) struvite (magnesium ammonium phosphate, MAP) production from digested EBPR sludge. The first three options are discussed in detail under Sect. 4.2. Struvite (MgNH4PO4.6H2O) is an orthophosphate, containing magnesium, ammonium, and phosphate in equal molar concentrations [105]. Struvite, which is also known as MAP in the wastewater sector, has low water solubility and forms crystals under anaerobic conditions in digesters and downstream processes [107]. Struvite formation was firstly observed in 1937 in a sludge digestion system [107]. Since then, struvite scale deposits blocking valves, pipes, centrifuge bowls, and pumps have been known to be a nuisance in WWTP [108, 109]. As a solution to struvite problem, dilution of struvite crystals, chemical precipitation, and use of chemical inhibitors are used in WWTP [110–112]. Although struvite is a common operational problem in WWTP for a long time, controlled struvite crystallization in side stream generated from anaerobic digesters of WWTPs has gained interest as a route for phosphorus recovery in the last decade [108, 109]. In 2006, the first struvite

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recovery plant was integrated into a full-scale WWTP in the Netherlands [113]. Bench and pilot scale studies evaluating the optimization of potential methods (nucleation and crystal growth) for phosphorus recovery as struvite are still quite popular all around the world [105].

4.5 Nitrogen Recovery Nitrogen, which is essential for plants in its reactive forms as ammonium, nitrite, and nitrate, is a common fertilizer for agricultural activities. The invention of Haber–Bosch process in 1909 enabled the industrial production of ammonia from N2 gas in the atmosphere, provided increase in food production capacity. Humans and animals fed with N-based fertilized foods started to excrete higher urea and ammonium, which led to an increased nitrogen load in WWTP [113]. In conventional WWTP, the main approach for nitrogen management is to remove nitrogen from wastewater with biological nitrification and denitrification processes, which are energy intensive with a significant greenhouse gas footprint, while successfully complying with effluent discharge limits. Very recently, energyefficient novel nitrogen removal technologies, which do not produce greenhouse gases (Sect. 3), have been introduced to replace the conventional processes. However, none of these technologies, conventional or new ones, provides recovery of nitrogen, but loss of nitrogen by releasing to the atmosphere as N2. With WEF Nexus approach (Sect. 2), research activities have begun to shift toward recovery of nitrogen from wastewater streams to save energy and natural resources [114]. Nitrogen in wastewater can be mainly recovered in WWTP from mainstream or side stream (reject water). The recovery of nitrogen as urine from the source of generation is also another option. Mainstream recovery options must be evaluated in detail for energy requirements and recovery efficiencies since inlet wastewater streams are characterized by high volumes but low concentrations. Due to the high concentration of nitrogen in reject water/digestate, struvite, which is the common

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product mainly developed for phosphorus recovery (Sect. 4.4) from digestate, has been a subject of interest to optimize simultaneous phosphorus and nitrogen recovery [114]. Sourceseparated urine, which is separated at source (toilet) from other wastewater streams to prevent a dilution, is the most concentrated stream for nitrogen recovery [115]. Nitrogen recovery technologies can be classified as chemical, physical, and biological technologies [114]. Precipitation as struvite, ion exchange, and adsorption-based methods (carbon-based adsorbents, zeolite, concentrated ammonium/phosphate precipitation), bioelectrochemical systems (microbial electrolysis cell, microbial fuel cell, microbial desalination cell), stripping technologies and membrane processes are among those processes [114].

4.6 Valorization of Sludge to Biopolymers Biopolymers, which are classified as microbially derived intracellular (e.g., polyhydroxybutyrate (PHB in polyhydroxyalkanoates (PHAs) group) and extracellular polymers (EPS) can be recovered from sewage sludge [116]. Regarding to their potential contributions to circular economy, biopolymers have received great attention because of their possible use in various diversified applications. Although PHAs can have a potential to be used as filler for nonbiodegradable plastics, agriculture systems for prolonged release of fertilizers, agrochemicals and medicine, their high production cost limits their application [117]. Additionally, PHB marketability is limited due to its brittle nature, large spherulites, low mechanical strength, and its large extent of secondary crystallization [118]. Moreover, EPS recovered from waste sludge not only appear as an attractive alternative to synthetic polymers, but has also a potential use in innovative industrial/environmental applications (e.g., gelforming materials for paper industry, biosorbents, cement curing materials, and flame retardant materials). In the coming years, a

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significant contribution of EPS-based biomaterials to circular economy is expected due to their renewable origins [116]. PHBs are usually produced by microorganisms under stress conditions when carbon source is abundant [119]. PHA production process commonly includes major three steps with the use of different reactor configurations and microbes: (i) pre-fermentation, (ii) enrichment under feast and famine conditions, and (iii) accumulation [120]. Among these steps, prefermentation increases the cost of PHA recovery due to recalcitrant nature of waste activated sludge that hinders hydrolysis and acidogenesis and hence require application of pre-treatment technologies (e.g., thermal, chemical, or mechanical) prior to acidogenic fermentation of sludge to produce volatile fatty acids VFA) [119]. Selection of microorganisms with high PHA storage capacity and their enrichment is also another crucial step in PHA recovery [120]. For feasibility of PHA production in circular economy, a great research effort is dedicated to find potential new bacterial producers with optimization of fermentation process using low-cost carbon sources. EPS take a large fraction of sludge dry weight and can be secreted by the bacterial consortium during metabolism, and their accumulation helps to bridge bacterial cells and other particles into aggregates [116]. EPS forms are subdivided as bound EPS (B-EPS) and soluble EPS (S-EPS) with different sub-classes [121]. EPS recovery is heavily dependent on EPS production, efficient extraction, and purification. EPS production should be evaluated in terms of yield, composition and structure, and physical properties, which are heavily dependent on internal factors (e.g., type of microbial community, diverse metabolisms, etc.) and external environmental factors (e.g., operational parameters, substrates, exogenous substances) [116]. With the development of novel biological treatment technologies including granular sludge technologies, EPS extraction from granular processes (e.g., aerobic granules, anaerobic granules and Anammox granules) has appeared as a promising recovery option, since the EPS contents of granular sludges are

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significantly higher than of suspended-growth and biofilm type processes [122]. Up-to-date, several methods for the EPS extraction (e.g., centrifugation, sonication, blending and heating) have been developed for various kinds of sludges [123].

of antioxidants, photographic developers, insecticides, and blowing agents for plastics. Moreover, it is used for metal recovery by chemical reduction as useful insoluble elemental metal instead of metal hydroxide sludge [124]. There are also few studies [125, 126] evaluating the use of hydrazine for the synthesis of magnetic nanoparticles, which have widespread applications in biotechnology, biomedical, material science, engineering, and environmental areas. Very recently, Erdim et al. [56, 59] demonstrated that the use of zero valent iron nanoparticles enhanced the Anammox process in terms of growth, degradation, EPS secretion and granulation. There are limited number of studies demonstrating the hydrazine and hydroxylamine generation potential in Anammox systems. Very recently, Alpaslan Kocamemi and Celik [56] scanned the hydroxylamine and hydrazine levels in various laboratory scale and pilot scale Anammox reactors that were either newly started-up or operating for a long period of time. Based on the scanning results, the effluents were classified as the ones containing the highest hydrazine concentration, highest hydroxylamine concentration, and both hydrazine and hydroxylamine concentrations. The hydrazine concentrations varied between 2 and 104 ppb, while hydroxylamine ranged between 7 and 572 ppb [56]. Future research should be devoted to accumulating and recovery of these value-added two compounds from Anammox systems.

4.7 Potential Value-added Compounds from Novel Nitrogen Removal Technologies As being energy-efficient and greenhouse gas free processes, novel nitrogen removal technologies, especially Anammox process and hence deammonification applications (Sect. 3.3– 3.4) have become more popular in WWTP. Soon, mainstream Anammox (deammonification) systems will replace the conventional biological nitrogen removal facilities. Anammox is a unique process, in which hydrazine (N2H4) is synthesized biologically hydroxylamine (NH2OH) is another intermediate of the process (Fig. 12). Hydroxylamine is also generated as an intermediate compound in partial nitrification and Comammox processes. Both hydroxylamine and hydrazine are known as reducing agents, which are widely used in industry and pharmacy. The high chemical reactivity of hydrazine imparts wide spectrum use in industrial applications, e.g., in rocket fuels, in power plants, in the production

Fig. 12 Intermediate products of Anammox pathways

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The Use of Molecular Tools in the New Context of Wastewater Treatment

Wastewater treatment is a complex process that is a combination of physical, chemical, and biological processes. Microorganisms are integral parts of the treatment process responsible for the conversion and degradation of pollutants in wastewater. With the omics approach (Fig. 13), one can investigate the microorganisms in wastewater treatment systems and answer several questions such as what is the microbial community structure (16S rDNA)? What is their genomic capability (Metagenomics)? Which genes or pathways are actively used (Metatranscriptomics)? Next Generation Sequencing (NGS) is a highthroughput sequencing technology that provides accurate and quick DNA or RNA sequences from environmental samples [127]. NGS technologies, which were limited in use initially due to cost, expertise and equipment requirements, and the lack of standard techniques in data analysis and interpretation, have now become widely used in wastewater treatment research as these problems have been overcome. NGS applications can be divided into four groups: amplicon sequencing (targeting 16S rRNA gene or functional genes),

whole genome sequencing (WGS) of enriched or pure culture isolates from process, metagenomic sequencing, and meta-transcriptomic sequencing.

5.1 Targeted Sequencing of Amplified Gene Regions The most common target gene for amplicon sequencing is the 16S rRNA gene, which is universal to bacteria and archaea. With this approach, it determines the phylogeny or taxonomy of the members of the microbial community [128]. In this method, usually the V4 region of the 16S rRNA gene is targeted using primers 515F-806R. The reads obtained after sequencing can be analyzed using the Qiime2 program [129]. Briefly, cleaning of reads, obtaining OTU (Operational Taxonomic Unit) table, and determining representative reads are achieved by applying the DADA2 [130] or deblur [131] algorithm. At this stage, chimeric readings are also removed. The taxonomic annotation of OTUs is determined by comparing them with databases specific to the 16S rRNA gene, such as the Silva (https://www.arbsilva.de/), Greengenes (https://greengenes. secondgenome.com/), or RDP (http://rdp.cme. msu.edu/) databases. Qiime2 program also

Fig. 13 Multi-omics approach on microbial ecology studies

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provides the necessary tools for statistical analysis of data and calculation of alpha and beta diversity indexes. Amplicon sequencing of 16S rRNA genes is often used to understand microbial community structure. Wu et al. [132] performed 16S ribosomal RNA gene sequencing in around 1200 activated sludge samples taken from 6 continents, 23 countries, and 269 WWTP on a global scale. They detected 28 core OTUs, which are strongly correlated to activated sludge process. Furthermore, microbes of activated sludges show no latitudinal gradient. Studies conducted on a global scale will be important in elucidating the global distribution of microorganisms in WWTP and their relationship with the process. Although amplicon sequencing of 16S rRNA gene is powerful for identifying microbial taxonomy accurately at the phylum, class, order and family levels, reliable annotation at genus, species, or strain level is often imperfect. To increase accuracy of taxonomic annotation, the use of long read sequencing will be in our arsenal [133].

obtain more complete genomes from environmental samples. For example, Liu et al. [134] obtained eight high-quality circular bacterial genomes from the partial nitritation-anammox (PNA) reactor using the hybrid assembly method. Using a similar approach, Frank et al. [135] obtained the entire genome of the Kuenenia stuttgartiensis strain isolated from a membrane bioreactor. In their analysis, they detected the new protein type 3b (sulf) hydrogenase and an additional copy of the hydrazine synthase gene cluster in new strain. Therefore, obtaining a good quality finished genome clarify the biochemical and metabolic processes of these organisms.

5.2 Whole Genome Sequencing (WGS) WGS is a powerful method for understanding the genetic potential of isolated microorganisms. After DNA isolation from enriched or pure cultures obtained from wastewater treatment systems, the genome of the microorganism can be assembled with read using shotgun sequencing approaches. Daims et al. [74] used metagenomic binning technology to obtain the Comammox nitrospira genome. This work led to the discovery of a completely new process, complete ammonia oxidation, instead of the traditional separate nitrification–denitrification process. However, obtaining good quality finished genome is very problematic due to assembly of genome from short reads. Assembly and binning methods from short reads produce highly fragmented draft genomes. Therefore, the advantages of both long and short reads must be combined to

5.3 Metagenomic Sequencing Metagenomics sequencing can deliver information on the structure and function of an entire microbial community. With the help of metagenomics, metabolic processes can be constructed and predict the function of potential genes [136]. Metagenomics studies on wastewater treatment have mainly focused on nitrogen and phosphorus removal, pollutant degradation, and antibiotic resistance genes (ARGs) [137]. For example, shotgun metagenomics has been used to study microbial community relations in Anammox systems [138, 139] and increased the knowledge about denitrification microbial diversity by obtaining 23 metagenomes from metagenome of denitrifying microorganisms in the partial nitration-anammox (PNA) process. Chen et al. [137] concluded from metagenomic studies that large particles have high biodiversity and functional diversity, which may be the reason for the good denitrification performance of large granular-based PNA. Metagenomics approach can be very useful both for determining the diversity of nitrogen-cycling organisms and for finding new pathways in nitrogen metabolism. These studies could help to build more costeffective and sustainable biological nitrogen removal systems.

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5.4 Metatranscriptome Sequencing Metagenomics sequencing can deliver information on the structure and function of an entire microbial community. However, it is inadequate to detect microorganisms that play a role in these metabolic processes. Metatranscriptomics is directly sequence RNA transcripts belong to all microorganisms. Therefore, it produces superior understanding into microbial activity than metagenomics. Metatranscriptomics can provide a picture, of which microbes and pathways are functionally active. RNA sequencing is particularly crucial for identifying RNA viruses such as SARS-CoV-2. Nemudryi et al. [140] successfully sequenced the SARS-CoV-2 genome from wastewater using Oxford Nanopore sequencing technology and determined their phylogenetic origin. Multiple omics technologies can be used to resolve microbial community metabolic abilities at the genome, transcriptome levels, enabling more comprehensive interpretation of the collaboration between microbial populations.

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Future Development: Integration of Nanotechnology to Wastewater Treatment Applications

Despite of its long history and many outstanding benefits, biological wastewater treatment technology still faces tremendous challenges to meet today’s increasing demand for wastewater reuse and energy recovery [141]. One possible strategy would be to improve the metabolism of microorganisms by adding certain elements in the nanoform. Nanomaterials are typically defined as materials having the particle size of one billionth of a meter (nm = 10−9 m). Due to their extremely small size, they contain a significantly lower number of atoms, which in turn leads to very different properties as compared to their bulk materials. Nanomaterials are characterized by their unique physical, chemical, and biological properties. Because of their small size, high surface area, and ease of functionalization,

they offer unprecedented opportunities to improve the efficiency of wastewater treatment processes. Particularly, some processes such as the novel nitrogen removal Anammox process suffer from the long start-up periods. Although several types of nanomaterials have been suggested to be integrated with wastewater treatment processes, iron nanoparticles (nZVI) have received special attention as they are cheap, nontoxic, and can accelerate certain enzymatic reactions. Particularly, iron ions are very important for the enzymes needed by Anammox bacteria for energy conservation. Iron's presence in the anammoxome and its role in the electron transport chain have both been proven [142]. Most of the cellular iron is found in the anammoxome in the form of co-factors found in heme cytochromes and iron-sulfur (Fe-S) proteins [143], suggesting that this unusual organelle may not only be the site of energy production but also may function as an iron storage facility for enzymes that contain heme [144]. The long-term impact of nZVI addition on Anammox treatment performance, Anammox growth activity, and granulation and biofilm formation in a continuously operating SBR has been thoroughly examined by Erdim et al. [59] (Fig. 14). Their research showed that iron had various advantages for Anammox activity when applied at the nanoscale and in concentrations between 0.04 and 5000 ppb. Due to its enormous surface to volume ratio, nanoscale iron exhibited a considerably higher/faster reactivity and was easily able to diffuse into the compartments of the cell through pores. Additionally, nZVI has a significant reducing power that is advantageous for the growth of anaerobic bacteria, with a standard electrode potential of − 0.44 V and its small size [145]. Because nZVI had a valence of zero, it demonstrated a significant propensity to give electrons in water and produced hydrogen (H2), a useful electron donor. Additionally, nZVI use promoted the release of extracellular polymeric substances (EPS), which improved Anammox bacteria's granulation. The findings showed that intermittent dosage of low concentration nZVI was more

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Fig. 14 Effects of nZVI addition on Anammox bacteria [58]

effective at boosting Anammox bacteria's metabolic activity. Therefore, using nZVI is recommended as a feasible strategy, particularly to address the problems with traditional Anammox applications for wastewater treatment. However, to preserve the long-lasting benefits of nZVI for Anammox systems, dosing frequency and nanoparticle amount modification are required. Anaerobic digestion is another important biological process, in which application of nanoadditives are thoroughly investigated. At modest concentrations (about 10 mg/L), the addition of various nano-additives, like nZVI and iron oxides, produces notable improvements by promoting the activities of microorganisms and important enzymes [146–148], which increases gas production and improves effluent quality. The results of various papers concentrating on the effects of metallic nanomaterials on both AD processes of biological hydrogen-generation and methane synthesis have been discussed by Zhu et al. [149].

Among all metallic nano-additives, those based on iron showed the highest reported tolerable concentrations of over 150 mg/L. The presence of nano-copper entirely hindered the AD fermentation phase, but low concentrations of nano-ZnO and NiO significantly enhanced AD fermentation phases without having any negative effects.

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Microplastics Removal Performance Through Advanced Treatment Technologies: A Mini Review Hanife Sari Erkan and Guleda Onkal Engin

Abstract

Keywords

Microplastics, which are on the agenda of the scientific world due to their adverse effects on living organisms, are reported to be removed with high removal efficiencies in conventional wastewater treatment plants. However, high amounts of microplastics reach the natural environment as large volumes of treated wastewater are discharged from wastewater treatment plants. Especially in developed countries, stringent environmental regulations ensure upgrading and retrofitting the wastewater treatment plants with advanced treatment technologies. There are studies indicating that advanced treatment technologies, including membrane filtration and advanced oxidation could remove microplastics effectively. Considering the necessity to take measures under the Green Deal, the removal of microplastics becomes an important parameter for evaluating the performance of wastewater treatment technologies. This paper, therefore, provides an analysis of advanced treatment technologies for the removal of microplastics from wastewater.

Microplastics Wastewater Advanced treatment technologies Membrane filtration Advanced oxidation

H. Sari Erkan
G. Onkal Engin (&) Department of Environmental Engineering, Yildiz Technical University, Istanbul, Turkey e-mail: [email protected]




1










Introduction

Microplastics (MPs) are recognized as environmental pollutants that pose a serious threat to wildlife and public health [1–3]. MPs are defined as plastic particles having sizes of typically between 1 and 5000 µm [4]. MPs can either be prepared on purpose to be used as exfoliating agents in personal care products or formed from the breakdown of larger plastic debris [5]. A recent study showed that the MP abundance in seas and oceans is significant [6]. These MPs enter the food chain through organisms living in natural waters and, consequently, appear in humans [7– 9]. It is reported that microplastics can cause gastrointestinal problems when ingested by various organisms [10, 11]. In parallel, various chemicals such as plastic additives and heavy metals that are typically adsorbed on MPs pose potential toxicity to the fauna and flora [12]. Likewise, MPs have hydrophobic surfaces that allow for persistent organic pollutants (POPs) to adhere, particularly, when an MP has a large surface area and a small volume [13]. Similarly, heavy metals can hold on to surfaces of MPs [14, 15].

© The Author(s), under exclusive license to Springer Nature Switzerland AG 2023 E. Debik et al. (eds.), Wastewater Management and Technologies, Water and Wastewater Management, https://doi.org/10.1007/978-3-031-36298-9_16

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Microplastics originate from different sources including laundry, wear and tear of tires, industrial activities, landfill sites, marine activities, and so on [16]. Microplastics originating primarily from cosmetics, chemical, pharmaceutical, and textile industries reach aquatic ecosystems through storm water and wastewater [17, 18]. As a result of the discharge of wastewater treatment plant (WWTP) effluents into the aquatic media, microplastics accumulate in lakes and oceans [19]. Macroplastics, on the other hand, can reach the aquatic environment in various ways as a result of human activities and can be fractured by UV irradiation, waves, and wind to form secondary MPs [20]. Most of the microplastic research studies are concentrated on the effects of MPs on marine environment and marine ecosystems [21]. However, the number of articles is increasing year by year focusing on the MP pollution in other compartments of the environment and how to deal with it. A quick SCOPUS analysis for search words “microplastics” only and “microplastics”, “wastewater” and “treatment” demonstrated that there is an increasing trend in the investigation of MP removal and retention technologies at WWTPs; however, comparing the orange and blue lines in Fig. 1, there is still a big room for improvement. Wastewater treatment plants are shown as the major sources of MP pollution in marine systems [22–24]. Controlling the discharge of MPs from wastewater treatment plants is the most basic approach that can be taken to prevent MPs from entering water bodies. Although it is reported that the removal efficiencies may go up to 99% in some cases, the discharged MP number is still high considering the large volumes of daily discharged wastewater. However, it should be noted that wastewater treatment plants serve as final barriers to MP pollution [25], and therefore, it is important to select suitable technologies to prevent the release of MPs into the marine environment. Therefore, this chapter addresses the efficiency of treatment technologies used to remove MPs in wastewater treatment plants. It focuses on specific technologies used for MPs removal and their efficiencies in wastewater treatment plants.

H. Sari Erkan and G. Onkal Engin

For this purpose, different technologies including membrane technologies, filtration, and advanced oxidation are reviewed. This paper is, therefore, prepared to serve the scientific community to show the most efficient technologies that can be used in classical wastewater treatment plants for MP removal.

2

Urban Wastewater Treatment Plants

Wastewater treatment plants are regarded as major sinks and sources of land-based MP pollution [26, 27]. MPs abundance in effluents depends mainly on wastewater catchment properties, sewer system, population served, treatment processes applied, and operational factors [26, 28]. In a recent study conducted on three WWTPs in the USA, it was shown that a combined load of 500–1000 million MPs per day is discharged to water media [22]. In an earlier study [29] conducted at 17 different WWTPs in the USA, however, a lower estimation of MPs discharge in the effluent was found with a range of 50,000 and 15 million MPs per day in comparison with the study conducted by Conley et al. [22]. These results indicate that removal efficiencies may vary according to the received wastewater and treatment processes applied. It should be noted that the physicochemical properties, such as density, charge, and hydrophobicity of microplastics found in wastewater, also play an important role in the efficiency of the treatment technology [30]. It should also be noted that the identification methods used may create some differences in the results obtained, as there is no standardized measurement and identification method. To determine the efficiency of the treatment technology, MPs must be detected and defined properly. It is possible to reach a significant number of detection methods for the morphological characterization of MPs [31, 32]. MPs removal efficiencies of conventional urban wastewater treatment plants are reported to vary from 79.3% to 99% [26, 33, 34]. In primary treatment, where floating or settling is the specific removal technologies, MPs removal

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Fig. 1 SCOPUS analysis for search words “microplastics” only and “microplastics”, “wastewater” and “treatment” between 2005 and 2022 (May 2022)

efficiencies were reported to vary from 25.1% to 91.6% [35–39]. During the secondary treatment, MPs removal efficiencies increase from 64.4% to 99.1% [33, 37, 40, 41]. However, it was stated that changes in surface loading as a result of various problems encountered in treatment plants can cause MPs to resurface and thus to leave the system without being removed [42]. When MPs are removed from aqueous phase by settling, they are retained in the primary and secondary sludge. After several different treatment steps, waste sludge is either landfilled, incinerated, or land applied, thus forming pathways for MPs to enter different compartments of the environment. The studies investigating MPs abundance in waste sludge are relatively lower than studies conducted with wastewater samples. Several studies showed that the number of MPs detected per gram of waste sludge changed from 1 to 113 MPs [36, 43–45]. The obtained data reveal that the MPs detected in the sludge exceed those in the treated wastewater by one or two orders of magnitude. In addition, studies conducted in different parts of the world demonstrated varying results. The main reasons for this could be the

differences in wastewater characterization and the changes in the treatment plant configuration. The point to be considered here is the revision of waste sludge treatment and disposal technologies in a way so that further spread of MPs pollution could be prevented. Tertiary treatment technologies are used to meet increasingly stringent discharge requirements for inorganic and organic pollutants that cannot be removed during secondary treatment. It is reported that tertiary treatment generally removes 73% to 99.9% of microplastics [35, 46, 47]. In the following sections, several tertiary treatment technologies that have been investigated for their effectiveness in removing microplastics will be discussed.

3

Microplastic Removal Technologies

3.1 Sand Filtration Sand filtration is used in wastewater treatment plants as a final stage, especially in the removal

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of residual suspended solids and other particulate matter by adsorption or physical retention of solids. In the literature, sand filtration performance to remove MPs was evaluated by some researchers and found that sand filtration could be used to polish up the treated wastewater for further MP removal [48–50]. The removal of MPs was investigated using sand filtration in the final treatment step of two different municipal WWTPs in Germany [50]. In the same study, the effectiveness of the sand filter installed in the wastewater treatment plant of a PVC production facility was also investigated. The results revealed that sand filtration is quite successful in MPs removal and that both WWTPs and the PVC production plant have achieved a removal rate of over 99% [50]. Hidayaturrahman and Lee [49] examined two different WWTPs in South Korea for the effectiveness of sand filtration with membrane disk filter in MPs removal. While the removal efficiency in sand filtration, having a depth of 6.8 m and residence time of 1.08 h, was limited to 73.8%, it was reported that a slightly higher removal efficiency was obtained in membrane disk filtration (79.4%). The difference could be due to several factors such as wastewater characterization, public habits of water usage, and the existence of industrial plants in the wastewater catchment area. However, when both WWTPs are considered, it has been reported that the removal efficiency is around 99%. Similarly, Talvitie [48] assessed four different WWTPs and reported MPs removal efficiency of 97.1% by sand filtration.

3.2 Advanced Oxidation Microplastics forming polymers consist of long carbon chains that can be broken via advanced oxidation. However, plastic products may contain chemical additives that make them more flexible and resistant to various environmental factors such as weather conditions. The methods used for the removal of MPs generally provide the transition of MPs from the liquid phase to the solid phase (slurry or sludge). Therefore, it can

H. Sari Erkan and G. Onkal Engin

be said that they provide a partial solution to the removal of MPs. Advanced oxidation methods are used for the removal of persistent organic pollutants, by the production of hydroxyl (HO•), or sulfate (SO4−•) radicals having a high oxidation potential. In the literature, Fenton-like oxidation processes [51, 52], electro-oxidation [53], and photocatalysis [54] have been tested for MPs removal. In these studies, the targeted MPs were generally added to water samples to obtain test media. In most cases, only one type of polymer, such as polyethylene terephthalate (PET) or polystyrene (PS) was added into water samples. As an example, a study was carried out by Kiendrebeogo et al., in which polystyrene (PS) was degraded by anodic oxidation [53]. The authors reported that MPs can be degraded in two different ways by direct anodic oxidation and indirect electrochemical oxidation using mediators such as persulfate anions. Anodic oxidation takes place on the surface of the anode material, in contrast to electrochemical oxidation, where the oxidation occurs in aqueous solution. The authors conducted a parametric analysis experimenting different factors such as current intensity, anode material, and anode surface area, supporting electrolyte concentration and type. The overall results showed that 58 ± 21% of MPs could be degraded in 1 h operation time. When the operation time increased to 6 h, the efficiency increased to 89 ± 8%. It was stated that the increase in the removal efficiency slowed after 2 h operation time reaching steadiness. It was also found that complete mineralization was achieved. Photocatalytic degradation of polypropylene (PP) microplastic particles in water by visible light irradiation of zinc oxide nanorods immobilized on glass fiber substrates was investigated by Uheida et al. [55]. A 65% reduction in mean particle volume was reported at the end of the two-week irradiation period. In another study [56], the decomposition of polyethylene (PE) was investigated using ozone and a combination of ozone and ozone + H2O2. As a result of the ozonation process under certain

Microplastics Removal Performance Through Advanced …

conditions (pH: 12; ozone flow rates: 1, 3, and 5 L/min, operation time: 180 min), the density in the O–H and C = O bonds and the active groups on the PE microplastics were evaluated through Fourier transform infrared spectrophotometer (FTIR) by quantifying the O–H and C = O bonds on microplastic particles selected from the infrared spectra ranging from 4000 to 550 cm−1 [56]. The FTIR results showed that ozonation led to an increase in the intensity of the O–H and C = O bonds and thus a decrease in the transmittance percentage. In the same study, the combination of ozonation and H2O2 has also been shown to have a synergistic effect on the degradation of PE microplastics. The carbonyl index (CI) values were evaluated to determine the degree of oxidation obtained. The highest CI value was found to be 1.33 at pH 12 with a flow rate of 3 L/min using the combination of ozonation and H2O2 method, suggesting that this hybrid process can be used as a pretreatment option for MPs degradation [56]. Most of the studies that apply advanced oxidation process to MPs degradation appear to be conducted at laboratory scale. While some of these studies report that ozonation has no effect on MPs removal [57, 58], others report high removal efficiencies. For example, Hidayaturrahman and Lee [49] reported that MP removal efficiency was >99%, when a hybrid process of coagulation followed by ozonation was utilized. Singh and Sharma [59] emphasized that the physicochemical properties of polymers such as adhesion, surface tension, and solubility change during the ozonation, accelerating the degradation process. In another study, it has been reported that surface roughness and hydrophobicity of MPs decreased after UV-C irradiation and ozonation [60]. As a result, MPs precipitation increased, as revealed by the sinking velocity experiment conducted after the oscillation of the suspension having different water matrices. However, it should be noted that factors such as the need to produce ozone on-site and high electricity consumption may limit the use of ozonation in MPs removal.

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3.3 Membrane Bioreactors Membrane bioreactors (MBR) are used in combination and/or in substitution of conventional activated sludge process (CAS). MBRs typically constitute a membrane unit, usually microfiltration that is integrated into the activated sludge process. Conventional secondary treatment is not designed to break down MPs due to the relatively short hydraulic retention time (HRT) values, but MPs removal is reported to be via sludge [28, 61]. The use of MBRs in the treatment of domestic and industrial wastewater is increasing rapidly; therefore, its effectiveness in MPs removal seems to take an important place in research, including laboratory and real-scale studies. In a recent study carried out by Lares et al. [62], MPs removal efficiencies of CAS and MBR (a pilot plant operated at the WWTP) were compared by taking into consideration both wastewater and sludge samples. The highest concentrations of microplastics were found in influent samples (57.6 ± 12.4 MPs/L) and digested sludge samples (170.9 ± 28.7 MPs/g dry weight), indicating that MPs eventually accumulate in the sludge. They concluded that MPs were removed by the processes prior to the CAS system, and fibers were removed more efficiently (99.1%) compared to other MPs particles [62]. Different results regarding the effects polymer type and shape on removal efficiencies were reported by a number of researchers [49, 63]. The results indicated that cultural differences and, therefore, wastewater characteristics are the main factors affecting the polymer type and hence the MPs removal mechanisms. In another study conducted by Lv et al. [46], MPs removal performance of a full-scale MBR system was compared to that of an oxidation ditch. The influent MPs were removed more efficiently (82.1%) by the MBR system, when compared with 53.6% of the oxidation ditch in terms of MPs particle numbers. It has been stated that the MBR system allows the microplastics to remain in the reactor as the extracellular

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polymeric substances (EPS) produced in the mixed liquor that creates an adherent and viscoelastic medium. In addition, it has been reported that the passage of MPs to the effluent is largely prevented by microfiltration process (