Sustainable Agriculture 9048126657, 9789048126651

Citation preview

Environmental Chemistry for a Sustainable World 66

Nadia Morin-Crini Eric Lichtfouse Grégorio Crini  Editors

Emerging Contaminants Vol. 2 Remediation

Environmental Chemistry for a Sustainable World Volume 66

Series Editors Eric Lichtfouse , Aix Marseille University, CNRS, IRD, INRA, Coll France, CEREGE, Aix en Provence, France Jan Schwarzbauer, RWTH Aachen University, Aachen, Germany Didier Robert, CNRS, European Laboratory for Catalysis and Surface Sciences, Saint-Avold, France

Other Publications by the Editors Books Environmental Chemistry http://www.springer.com/978-3-540-22860-8 Organic Contaminants in Riverine and Groundwater Systems http://www.springer.com/978-3-540-31169-0 Sustainable Agriculture Volume 1: http://www.springer.com/978-90-481-2665-1 Volume 2: http://www.springer.com/978-94-007-0393-3 Book series Environmental Chemistry for a Sustainable World http://www.springer.com/series/11480 Sustainable Agriculture Reviews http://www.springer.com/series/8380 Journal Environmental Chemistry Letters http://www.springer.com/10311 More information about this series at http://www.springer.com/series/11480

Nadia Morin-Crini  •  Eric Lichtfouse Grégorio Crini Editors

Emerging Contaminants Vol. 2 Remediation

Editors Nadia Morin-Crini Chrono-environnement, UMR 6249 Université Bourgogne Franche-Comté Besançon, France

Eric Lichtfouse Aix Marseille University, CNRS, IRD, INRA, Coll France, CEREGE Aix en Provence, France

Grégorio Crini Chrono-Environnement, UMR 6249 Université Bourgogne Franche-Comté Besançon, France

ISSN 2213-7114     ISSN 2213-7122 (electronic) Environmental Chemistry for a Sustainable World ISBN 978-3-030-69089-2    ISBN 978-3-030-69090-8 (eBook) https://doi.org/10.1007/978-3-030-69090-8 © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature Switzerland AG 2021 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

Preface

Man’s use of pollutants began to affect the environment during the Industrial Revolution. Today, we are in the Pollutant Removal Age. Grégorio Crini, 2010

Figure. An aerated lagoon to treat wastewaters from a pulp and paper industry; source: G. Crini

This book, entitled Remediation of Emerging Contaminants, is the second volume on emerging contaminants published in the series Environmental Chemistry for a Sustainable World and written by 64 international contributors from 24 v

vi

Preface

countries. The first volume provides an overview of on emerging contaminants, their presence in the water compartment, and their potential health effects. This second volume focuses on methods for remediation of emerging contaminants. The first chapter by Nadia Morin-Crini et  al. details chemical and biological methods for the removal of emerging contaminants in water. The second chapter by Borislav Malinović et al. reviews the electrochemical treatments for removing pollutants. In Chap. 3, Eric Lichtfouse et al. present technologies for removing selenium from water and wastewater. The final chapter by Grégorio Crini et  al. summarizes advanced treatments for the removal of alkylphenols and alkylphenol polyethoxylates from wastewater. The editors extend their thanks to all the authors who contributed to this book for their efforts in producing timely and high-quality chapters. The creation of this book would not have been possible without the assistance of several colleagues and friends deserving acknowledgment. They have helped by choosing contributors and reviewing chapters and in many other ways. Finally, we would like to thank the staff at Springer Nature for their highly professional editing of the publication. Besançon, France

Nadia-Morin Crini

Aix en Provence, France

Eric Lichtfouse

Besançon, France

Grégorio Crini

Contents

1 Remediation of Emerging Contaminants����������������������������������������������    1 Nadia Morin-Crini, Eric Lichtfouse, Marc Fourmentin, Ana Rita Lado Ribeiro, Constantinos Noutsopoulos, Francesca Mapelli, Éva Fenyvesi, Melissa Gurgel Adeodato Vieira, Lorenzo A. Picos-­Corrales, Juan Carlos Moreno-Piraján, Liliana Giraldo, Tamás Sohajda, Mohammad Mahmudul Huq, Jafar Soltan, Giangiacomo Torri, Monica Magureanu, Corina Bradu, and Grégorio Crini 2 Electrochemical Treatments for the Removal of Emerging Contaminants ��������������������������������������������������������������������  107 Borislav N. Malinović, Jernej Markelj, Helena Prosen, Andreja Žgajnar Gotvajn, and Irena Kralj Cigić 3 Technologies to Remove Selenium from Water and Wastewater����������������������������������������������������������������������������������������  207 Eric Lichtfouse, Nadia Morin-Crini, Corina Bradu, Youssef-­Amine Boussouga, Mehran Aliaskari, Andrea Iris Schäfer, Soumya Das, Lee D. Wilson, Michihiko Ike, Daisuke Inoue, Masashi Kuroda, Sébastien Déon, Patrick Fievet, and Grégorio Crini 4 Advanced Treatments for the Removal of Alkylphenols and Alkylphenol Polyethoxylates from Wastewater�����������������������������  305 Grégorio Crini, Cesare Cosentino, Corina Bradu, Marc Fourmentin, Giangiacomo Torri, Olim Ruzimuradov, Idil Arslan-Alaton, Maria Concetta Tomei, Ján Derco, Mondher Barhoumi, Helena Prosen, Borislav N. Malinović, Martin Vrabeľ, Mohammad Mahmudul Huq, Jafar Soltan, Eric Lichtfouse, and Nadia Morin-Crini Index������������������������������������������������������������������������������������������������������������������  399 vii

Contributors

Idil Arslan Alaton  Istanbul Technical University, Istanbul, Turkey Mehran  Aliaskari  Institute for Advanced Membrane Technologies (IAMT), Karlsruhe Institute of Technology (KIT), Karlsruhe, Eggenstein-Leopoldshafen, Germany Mondher Barhoumi  Faculté des Sciences Bizerte, Zarzouna, Tunisia Youssef-Amine  Boussouga  Institute for Advanced Membrane Technologies (IAMT), Karlsruhe Institute of Technology (KIT), Karlsruhe, Eggenstein-­ Leopoldshafen, Germany Corina  Bradu  Department of Systems Ecology and Sustainability, PROTMED Research Centre, University of Bucharest, Bucharest, Romania Irena Kralj Cigić  Faculty of Chemistry and Chemical Technology, University of Ljubljana, Ljubljana, Slovenia Cesare Cosentino  Istituto di Chimica e Biochimica G. Ronzoni, Milan, Italy Grégorio Crini  Laboratoire Chrono-environnement, UMR 6249, UFR Sciences et Techniques, Université Bourgogne Franche-Comté, Besançon, France Soumya  Das  Department of Geological Sciences, University of Saskatchewan, Saskatoon, SK, Canada Sébastien  Déon  Institut UTINAM, UMR CNRS 6213, UFR Sciences et Techniques, Université Bourgogne Franche-Comté, Besançon, France Ján Derco  Institute of Chemical and Environmental Engineering, Department of Environmental Engineering, Faculty of Chemical and Food Technology, Slovak University of Technology, Bratislava 1, Slovak Republic Éva  Fenyvesi  CycloLab Budapest, Hungary

Cyclodextrin

Research

and

Development

Ltd,

ix

x

Contributors

Patrick Fievet  Institut UTINAM, UMR CNRS 6213, UFR Sciences et Techniques, Université Bourgogne Franche-Comté, Besançon, France Marc Fourmentin  Université du Littoral Côte d’Opale, Laboratoire de Physico-­ Chimie de l’Atmosphère (LPCA, EA 4493), ULCO, Dunkerque, France Liliana  Giraldo  Facultad de Ciencias, Departamento de Química, Universidad Nacional de Colombia, Bogotá, Colombia Andreja  Žgajnar  Gotvajn  Faculty of Chemistry and Chemical Technology, University of Ljubljana, Ljubljana, Slovenia Mohammad  Mahmudul  Huq  Department of Chemical and Biological Engineering, University of Saskatchewan, Saskatoon, Canada Michihiko  Ike  Division of Sustainable Energy and Environmental Engineering, Graduate School of Engineering, Osaka University, Suita, Osaka, Japan Daisuke  Inoue  Division of Sustainable Energy and Environmental Engineering, Graduate School of Engineering, Osaka University, Suita, Osaka, Japan Masashi  Kuroda  Graduate School of Environmental and Disaster Research, Tokoha University, Shizuoka, Shizuoka, Japan Eric  Lichtfouse  Aix Marseille University, CNRS, IRD, INRA, Coll France, CEREGE, Aix en Provence, France Monica Magureanu  Department for Plasma Physics and Nuclear Fusion, National Institute for Lasers, Plasma and Radiation Physics, Magurele, Romania Borislav  N.  Malinović  Faculty of Technology, University of Banja Luka, Banja Luka, Bosnia and Herzegovina Francesca  Mapelli  Biotechnology and Environmental Microbiology Lab-BEaM Lab, Department of Food, Environmental and Nutritional Sciences (DeFENS), University of Milan, Milan, Italy Jernej  Markelj  Faculty of Chemistry and Chemical Technology, University of Ljubljana, Ljubljana, Slovenia Juan Carlos Moreno-Piraján  Facultad de Ciencias, Departamento de Química, Grupo de Investigación en Sólidos Porosos y Calorimetría, Universidad de los Andes (Colombia), Bogotá, Colombia Nadia  Morin-Crini  Laboratoire Chrono-Environnement, UMR 6249, UFR Sciences et Techniques, Université Bourgogne Franche-Comté, Besançon, France Constantinos  Noutsopoulos  Sanitary Engineering Laboratory, Department of Water Resources and Environmental Engineering, School of Civil Engineering, National Technical University of Athens, Athens, Greece

Contributors

xi

Lorenzo A. Picos-Corrales  Facultad de Ciencias Químico Biológicas, Universidad Autónoma de Sinaloa, Ciudad Universitaria, Culiacán, Sinaloa, Mexico Helena  Prosen  Faculty of Chemistry and Chemical Technology, University of Ljubljana, Ljubljana, Slovenia Ana  Rita  Lado  Ribeiro  Laboratory of Separation and Reaction Engineering – Laboratory of Catalysis and Materials (LSRE-LCM), Faculdade de Engenharia, Universidade do Porto, Porto, Portugal Olim  Ruzimuradov  Turin Polytechnic University in Tashkent, Tashkent, Uzbekistan Andrea  Iris  Schäfer  Institute for Advanced Membrane Technologies (IAMT), Karlsruhe Institute of Technology (KIT), Karlsruhe, Eggenstein-Leopoldshafen, Germany Tamás  Sohajda  CycloLab Cyclodextrin Research and Development Ltd, Budapest, Hungary Jafar Soltan  Department of Chemical and Biological Engineering, University of Saskatchewan, Saskatoon, Canada Maria Concetta Tomei  Istituo di Ricerca sulle Acque, Consiglio Nazionale delle Ricerche, Monterotondo, Italy Giangiacomo Torri  Istituto di Chimica e Biochimica G. Ronzoni, Milan, Italy Melissa Gurgel Adeodato Vieira  Department of Processes and Products Design, School of Chemical Engineering, University of Campinas, Campinas, SP, Brazil Martin Vrabeľ  Institute of Chemical and Environmental Engineering, Department of Environmental Engineering, Faculty of Chemical and Food Technology, Slovak University of Technology, Bratislava 1, Slovak Republic Lee D. Wilson  Department of Chemistry, University of Saskatchewan, Saskatoon, SK, Canada

About the Editors

Nadia  Morin-Crini  is an analytical chemist at the University of Bourgogne Franche-Comté, Besançon, France. Her research interests include the extraction and analysis of chemicals from various solid and liquid environmental matrices. She is responsible for a technological platform that studies old and present environments. She has published more than 130 papers in peer-reviewed scientific journals.

Eric Lichtfouse  is a geochemist and professor of scientific writing at Aix-Marseille University, France. He has invented carbon-13 dating and published the book Scientific Writing for Impact Factor Journals. He is chief editor of the journal Environmental Chemistry Letters and the book series Sustainable Agriculture Reviews and Environmental Chemistry for a Sustainable World. Grégorio Crini  is an environmental polymer scientist at the University of Bourgogne Franche-Comté, Besançon, France. His research activities focus on the design of new polymer networks and the environmental aspects of polysaccharide chemistry. He has published more than 210 papers in international journals and is a highly cited researcher with an h-index of 40 and more than 13,000 citations.

xiii

Abbreviations

ACF Activated carbon fiber AOP Advanced oxidation process AOPs Advanced oxidation processes BDD Boron-doped diamond electrode BGA Boron-doped graphene aerogel BOD Biological oxygen demand BPA Bisphenol A CB Conductive band CCM Coal-based carbon membrane CE Counter electrode CF Carbon fiber COD Chemical oxygen demand Dow pH˗dependent n˗octanol-water distribution ratio DC Divided cell E1 Estrone E2 17˗beta˗estradiol EC European Commission EE2 17˗alpha˗ethinylestradiol FD Frequency of detection GC Glassy carbon GC-MS Gas chromatography coupled with mass spectrometry GM Galvanostatic mode GUS Groundwater ubiquity score HPLC High-performance liquid chromatography HPLC-UV High-performance liquid chromatography with UV detection HQ Hazard quotient HR Hazard ratio IQR Interquartile range ISPRA Italian National Institute for Environmental Protection and Research Partition coefficient between soil organic carbon and water Koc Kow n-octanol-water partition coefficient xv

xvi

LC-MS/MS LOD LOQ LS MEC n.d. OR pKa PM PM PM10 PM2.5 PM0.1 PNEC SCE Sw SHE SS TOC UC VB WE WWTP ZVI

Abbreviations

Liquid chromatography coupled with (tandem) mass spectrometry Limit of detection Limit of quantification Light source Measured environmental concentrations Not detected Odds ratio The negative base-10 logarithm of the acid dissociation constant (Ka) of a solution Potentiostatic mode Particulate matter Coarse particulate matter Fine particulate matter Ultrafine particulate matter Predicted no˗effect concentration Saturated calomel electrode Water solubility Standard hydrogen electrode Stainless steel Total organic carbon Undivided cell Valence band Working electrode Wastewater treatment plant Zero-valent ion

Chapter 1

Remediation of Emerging Contaminants Nadia Morin-Crini , Eric Lichtfouse , Marc Fourmentin, Ana Rita Lado Ribeiro , Constantinos Noutsopoulos, Francesca Mapelli Éva Fenyvesi , Melissa Gurgel Adeodato Vieira , Lorenzo A. Picos-­Corrales , Juan Carlos Moreno-Piraján , Liliana Giraldo , Tamás Sohajda, Mohammad Mahmudul Huq, Jafar Soltan, Giangiacomo Torri, Monica Magureanu, Corina Bradu, and Grégorio Crini

,

Abstract  Water pollution by emerging contaminants has become a major source of concern and a priority for society and public authorities. Emerging contaminants are a group of natural and synthetic chemicals and biological agents that are not routinely monitored or regulated in the environment and may have known or suspected N. Morin-Crini (*) Laboratoire Chrono-environnement, UMR 6249, UFR Sciences et Techniques, Université Bourgogne Franche-Comté, Besançon, France e-mail: [email protected] E. Lichtfouse Aix Marseille University, CNRS, IRD, INRA, Coll France, CEREGE, Aix-en-Provence, France e-mail: [email protected] M. Fourmentin Université du Littoral Côte d’Opale, Laboratoire de Physico-Chimie de l’Atmosphère (LPCA, EA 4493), Dunkerque, France e-mail: [email protected] A. R. L. Ribeiro Laboratory of Separation and Reaction Engineering – Laboratory of Catalysis and Materials (LSRE-LCM), Faculdade de Engenharia, Universidade do Porto, Porto, Portugal e-mail: [email protected] C. Noutsopoulos Sanitary Engineering Laboratory, Department of Water Resources and Environmental Engineering, School of Civil Engineering, National Technical University of Athens, Athens, Greece e-mail: [email protected] F. Mapelli Biotechnology and Environmental Microbiology Lab-BEaM Lab, Department of Food, Environmental and Nutritional Sciences (DeFENS), University of Milan, Milan, Italy e-mail: [email protected] © The Author(s), under exclusive license to Springer Nature Switzerland AG 2021 N. Morin-Crini et al. (eds.), Emerging Contaminants Vol. 2, Environmental Chemistry for a Sustainable World 66, https://doi.org/10.1007/978-3-030-69090-8_1

1

2

N. Morin-Crini et al.

adverse effects on the environment and human health. The list of these substances is particularly long and includes pharmaceuticals, personal care products and cosmetics, pesticides, surfactants, industrial products and additives, nanoparticles and nanomaterials, and pathogens. Many emerging contaminants are released continuously into the environment and can cause chronic toxicity even at low concentraÉ. Fenyvesi CycloLab Cyclodextrin Research and Development Ltd, Budapest, Hungary e-mail: [email protected] M. G. A. Vieira Department of Processes and Products Design, School of Chemical Engineering, University of Campinas, Campinas, SP, Brazil e-mail: [email protected] L. A. Picos-Corrales Facultad de Ciencias Químico Biológicas, Universidad Autónoma de Sinaloa, Ciudad Universitaria, Culiacán, Sinaloa, Mexico e-mail: [email protected] J. C. Moreno-Piraján Facultad de Ciencias, Departamento de Química, Grupo de Investigación en Sólidos Porosos y Calorimetría, Universidad de los Andes (Colombia), Bogotá, Colombia e-mail: [email protected] L. Giraldo Facultad de Ciencias, Departamento de Química, Universidad Nacional de Colombia, Bogotá, Colombia e-mail: [email protected] T. Sohajda CycloLab Cyclodextrin Research and Development Ltd, Budapest, Hungary e-mail: [email protected] M. M. Huq · J. Soltan Department of Chemical and Biological Engineering, University of Saskatchewan, Saskatoon, Canada e-mail: [email protected]; [email protected] G. Torri Istituto di Chimica e Biochimica G. Ronzoni, Milano, Italy e-mail: [email protected] M. Magureanu Department for Plasma Physics and Nuclear Fusion, National Institute for Lasers, Plasma and Radiation Physics, Magurele, Romania e-mail: [email protected] C. Bradu Department of Systems Ecology and Sustainability, PROTMED Research Centre, University of Bucharest, Bucharest, Romania e-mail: [email protected] G. Crini Laboratoire Chrono-environnement, UMR 6249, UFR Sciences et Techniques, Université Bourgogne Franche-Comté, Besançon, France e-mail: [email protected]

1  Remediation of Emerging Contaminants

3

tions, endocrine disruption in humans and aquatic life, and the development of antibiotic resistant bacteria. It is therefore necessary to mobilize efforts to protect human health and biodiversity. Two of the main sources of emerging contaminants are wastewater treatment discharges and agricultural practices. However, conventional wastewater treatment plants have not been designed to remove such contaminants. It is therefore important to develop effective treatment methods capable of eliminating the parent emerging molecules and their metabolites. This is a difficult and challenging task because most of emerging contaminants are recalcitrant substances. The methods must also target not only water chemicals but also water microbiols in order to reduce and eliminate the toxicity and impact of the treated wastewater. During the past two decades, several physical, chemical and biological technologies have been proposed for emerging contaminant removal. Each technology has its own advantages and constraints not only in terms of cost, but also in terms of efficiency, feasibility, and environmental impact. However, among the various treatment processes currently cited for wastewater treatment, only few are commonly employed by the industrial sector for technological and economic reasons. Extensive research on this topic highlights the growing interest of scientists in developing treatment systems that are increasingly effective in removing mixtures of trace pollutants, simple to implement from a technological point of view, economically viable and environmentally friendly, with little or no impact on the environment. The objective of this chapter is to present the recent state of knowledge on the advanced treatments proposed for the removal of emerging contaminants in wastewater when they are present in trace amounts in wastewater. After general ­considerations on wastewater treatment plant, the first part is focused on adsorptionoriented processes using conventional (activated carbon, clays) or non-conventional (cyclodextrin polymers, metal-organic frameworks, molecularly imprinted polymers, chitosan, nanocellulose) adsorbents. Biosorbents such as cyclodextrin bead polymers have great potential in environmental applications although they are still at the laboratory study stage. The second part presents examples of biological-based technologies for the degradation and elimination of emerging contaminants. Selected biological approaches include constructed wetlands, biomembrane reactors, strategies based on the use of algae, fungi and bacteria, and enzymatic degradation. The third part briefly presents the membrane filtration strategy that is already used as a tertiary treatment. The final part is focused on advanced oxidation processes that also represent one of the most promising strategies because of their simplicity and efficiency. Keywords  Emerging contaminants · Substances of global concern · Pharmaceuticals · Personal care products · Pesticides · Water pollution · Aquatic compartments · Wastewater treatment plants · Remediation · Bioremediation · Conventional removal technologies · Advanced treatments · Adsorption · Biosorption · Cyclodextrin polymers · Constructed wetlands · Membrane bioreactor technology · Membrane filtration · Disinfection · Ozonation · Advanced oxidation processes · Catalytic ozonation · Non-thermal plasma

4

N. Morin-Crini et al.

1.1  Introduction Emerging contaminants such as pharmaceuticals and personal care products represent a new global water quality challenge, with potentially serious threats to human health and the environment, including water resources. Indeed, the presence of emerging contaminants and biological agents in treated and untreated water has received increasing attention because the risk they pose to human health, ecosystems and water resources is not yet fully understood. For example, antibiotics and endocrine disruptors are among the best examples of “new” contaminants that are known to the public and are of concern. Traces of these substances are being found in water resources, including drinking water, throughout the world. Numerous studies on the presence and behavior of pharmaceuticals in the environment have demonstrated negative effects to aquatic ecosystems and wildlife in general. Another worrying phenomenon is the emergence of antibiotic resistance. Overuse of antibiotics has encouraged the development of antibiotic-resistant strains. Antibiotic resistant bacteria and antibiotic resistance genes pose a public health concern when they transfer antibiotic resistance to human pathogens. All these topics are of great interest to the scientific community and concern to our societies. The term emerging contaminants includes chemical or biological contaminants without a clearly defined regulatory status. They are often substances, not necessarily of new use, but newly identified, for which data concerning their presence, their fate in the environment and their potential impacts on health or the environment are patchy. Among the emerging pollutants are, in particular, pharmaceuticals subject or not to medical prescription (antibiotics, hormones, illicit drugs, etc.) for human or veterinary use, products for daily use (personal care products, cosmetics, detergents, disinfectants, antioxidants, etc.), products of industrial origin (surfactants, additives, solvents, flame retardants, nanoparticles, etc.), pesticides, and pathogens. The number of molecules concerned is constantly changing both in terms of parent substances and their degradation products (metabolites, for example resulting from biological treatment). In general, these synthetic or naturally occurring chemicals are not commonly monitored in the environment and they are consistently being found not only in wastewaters but also in water resources (groundwater, surface water), and even in drinking water. Conventional domestic wastewater treatment plants (often stabilization ponds, activated sludge systems) were not designed to treat recalcitrant organic pollutants including emerging substances (Bolong et al. 2009; Teijón et al. 2010; Jelic et al. 2011; Verlicchi et al. 2012; Petrie et al. 2015; Machado et al. 2016; Priac et al. 2017; Botero-Coy et al. 2018; Gogoi et al. 2018; Tolboom et al. 2019; Patel et al. 2019). It is also important to point out that the elimination of microorganisms is not an objective of wastewater treatment plants either (Morin-Crini and Crini 2017). Botero-­ Coy et  al. (2018) reported that the concentrations of some pharmaceuticals (e.g. erythromycin, metronidazole, sulfamethoxazole) in the effluent from a municipal wastewater treatment plant (Bogotá, Colombia) were even higher than in influent, suggesting negative removal efficiencies. Jelic et  al. (2011) and Verlicchi et  al. (2012) previously reported similar results within European effluents, suggesting

1  Remediation of Emerging Contaminants

5

that the wastewater treatment plants act as a concentrator of contaminants. This incomplete removal evidences the need of alternative or complementary treatment technologies (Jelic et al. 2011; Verlicchi et al. 2012; Botero-Coy et al. 2018; Serna-­ Galvis et al. 2019). Conventional water treatments often involve a combination of methods such as sand filtration coupled with a coagulation pre-treatment, + adsorption + disinfection step, which are usually not effective to remove emerging pollutants. It is therefore important to develop effective treatment methods capable of eliminating not only the parent emerging molecules but also their metabolite and transformation products. These methods must also meet not only water chemistry objectives (particularly in terms of chemical abatement of the residual pollutant load and quality standards) but also water microbiology objectives, in order to reduce and eliminate the toxicity and environmental impact of the treated wastewater. These are two research topics that have been mobilizing the scientific community. After general considerations on wastewater treatment plants, this chapter presents and discusses advanced treatment methods such as adsorption-oriented processes, biological-based technologies, membrane filtration and advanced oxidation processes. In each of the sections, a state of the art was first conducted and then the authors’ research is presented to illustrate each water treatment technology.

1.2  Municipal Wastewater Treatment Plants Human use of chemicals such as metals began to affect the environment during the “Industrial Revolution” of the second half of the eighteenth century. Since the end of the Second World War, the use of chemical, pharmaceuticals and agrochemical products has greatly increased over the years to meet the growing needs of consumers, who are also rising in number. However, economic development has been accompanied by environmental problems including water pollution. In response to this pollution problem, the first wastewater treatment plants were set up in the 1960s in France, for example. The first European regulations on aqueous discharges appeared a decade later. Today, we are in the “Emerging Pollutant Removal Age” (Crini and Lichtfouse 2018a, b) and, it is, therefore, not surprising that there have been considerable efforts to develop technologies to reduce contaminant emissions. Indeed, the presence of pesticides, pharmaceuticals and personal care products in water challenges the scientific community and water treatment stakeholders to find simple, practical, inexpensive, effective and environmentally friendly disposal solutions. An important question arises: how do rivers receive emerging contaminants? It is now recognized that the presence in aquatic environments is mainly correlated with the discharge of effluents from municipal wastewater treatment plants (Fig.  1.1), with a higher incidence in wastewater treatment plants in highly industrialized and urban areas (Sole et al. 2000; Goméz et al. 2006; Stackelberg et al. 2007; Kasprzyk-­ Hordern et al. 2008; Watkinson et al. 2009; Basile et al. 2011; Gonzalez et al. 2012; Meffe and Bustamante 2014; Machado et  al. 2016; Morin-Crini and Crini 2017;

6

N. Morin-Crini et al.

Fig. 1.1  Over-consumption of pharmaceuticals and personal care products and contamination of rivers by the same substances present in wastewater from sewage treatment plants. (Region: Sinaloa, Mexico; source: Lorenzo A. Picos-Corrales, Sinaloa, Mexico)

Priac et al. 2017; Gilabert-Alarcón et al. 2018; Mezzelani et al. 2018; Picos-Corrales et al. 2020). However, environmental contamination/pollution originates not only from domestic activity but also from the industrial and agricultural sectors, and other activities such as transport (Morin-Crini and Crini 2017). Indeed, even if all discharges comply with regulations, industrial and urban wastes can still contain a mixture of substances, known or unknown (metabolites, degradation products), and their concentration varies from a few ng/L to a few mg/L. All the substances we use in our daily lives, such as pharmaceuticals or personal care products, can enter the environment through excretion in human and animal urine and feces, through the flushing of unused medicines, or through our domestic uses (cooking, bathing), and results in nanogram per liter to microgram per liter, even after adequate treatment. For example, Watkinson et al. (2009) investigated the presence of 28 antibiotics in several Australian water matrices, including six rivers and drinking water. The results showed that four substances, i.e., monensin, erythromycin, sulfamethoxazole and norfloxacin, were more frequently detected in surface water, at

1  Remediation of Emerging Contaminants

7

concentrations ranging from low ng/L to 2.0 μg/L. The authors demonstrated that in one river that did not receive discharges from municipal wastewater treatment plants, the total concentration of the investigated antibiotics was significantly lower than those in the other five rivers, suggesting that municipal wastewater treatment plants were an important source of antibiotics to the streams. Meffe and Bustamante (2014), Machado et al. (2016) and Gilabert-Alarcón et al. (2018) reported similar conclusions in Italian, Brazilian and Mexican water resources, respectively. When wastewater is contaminated and remediation becomes necessary, the best purification approach should be chosen to reach the decontamination objectives as established by legislation (Crini and Badot 2007, 2010; Morin-Crini and Crini 2017). A purification process generally consists of five successive steps: (1) preliminary treatment or pre-treatment (physical & mechanical); (2) primary treatment (physicochemical & chemical); (3) secondary treatment or purification (chemical & biological); (4) tertiary or final treatment (physical & chemical); and (5) treatment of the sludge formed (gravity or mechanical thickening and dewatering, drying or incineration). In general, the first two steps are gathered under the notion of pre-­ treatment or preliminary step, depending on the situation. Pre-treatment removes debris and large solids, while primary treatment focuses on the removal of an appreciable load of organic matter and suspended solids from the influent. This pre-treatment stages, which can be carried out using mechanical or physical means, is indispensable before envisaging secondary treatment because particulate pollution (e.g. suspended solids, colloids, fats, etc.) hinders the posterior treatment, making it less efficient or damaging the decontamination equipment. Primary chemical treatment such as pH adjustment or pre-reduction of a high organic load may also be required. Before its discharge into the environment or its reuse, the pre-treated effluent must undergo secondary purification treatment using in most cases biological processes to produce a satisfactory final effluent. In certain cases, a final or tertiary treatment (including disinfection) can also be required to remove the remaining pollutants or the molecules produced during the secondary purification. Around the world, almost the same wastewater treatment scheme is used by following the same steps above listed. From a water engineering point of view, the only difference is the type of matrix to be treated (domestic, industrial or drinking). In general, for municipal/domestic wastewaters treatment, the removal of pollutants from the raw wastewater is primarily achieved by biological means (often activated sludge systems). Indeed, biodegradation is one of the main treatment applied to decontaminate wastewater. For industrial wastewaters,  e.g. from textile, pulp and paper or surface-treatment industry, pollutant removal is achieved by physical (sand filtration, adsorption onto carbons, membrane filtration, evaporation, extraction), chemical (precipitation, oxidation, electrocoagulation, electrochemical treatment, ion-exchange using organic resins) and biological (lagoon, membrane bioreactors) means, with research focusing on more efficient and less expensive systems combinations or new alternatives such as biosorption, biomass, advanced oxidation processes. In general, precipitation is – together with adsorption – one of the two main treatments applied at industrial scale (Crini and Badot 2007, 2010; Crini 2010). For the production of drinking water, filtration (sand, carbon) and adsorption (activated

8

N. Morin-Crini et al.

carbon) steps are used, coupled with coagulation/flocculation and disinfection (ozonation or chlorination) steps. The main objective of a conventional municipal wastewater treatment plant, receiving raw wastewater from domestic and industrial discharges, is to remove organic matter and nutrients, e.g. nitrogen and phosphorus. A technological process widely used in the world is the so-called activated sludge, due to its competitive cost and high efficiency. Currently, under optimized conditions, more than 95% of organic and mineral substances can be eliminated by biological-based methods used in municipal wastewater treatment plants. In general, these performance levels make possible to meet the regulatory requirements for the list of substances and water parameters analyzed. However, the efficiency of a conventional system varies depending on the wastewater characteristics and the treatment process used, e.g. activated sludge, membrane biological reactor, with a possible combination with a chemical treatment. As already mentioned, existing conventional wastewater treatment plants and drinking water treatment plants were not designed to remove pharmaceuticals or nanoparticles, which are difficult to biodegrade or filter, respectively. For example, substances such as acetylsalicylic acid, diclofenac, carbamazepine, atenolol or propranolol are difficult to degrade by conventional biological treatment: performance is estimated to be less than 10% (Patel et al. 2019). Several studies have nevertheless showed that certain substances such as endocrine disruptor compounds can be reduced (Priac et al. 2017; Gogoi et al. 2018; Patel et al. 2019). In a wastewater treatment plant, besides biodegradation, adsorption to sludge is the most effective process to eliminate substances. Numerous substances are poorly soluble hydrophobic compounds, poorly biodegradable and present a strong affinity for sludge. A typical example is the anti-inflammatory diclofenac that can be partially adsorbed on sludge (Barbosa et al. 2016). In Europe, treated sludge after stabilization is often disposed in the soil or reused for agricultural purposes and composting. A previous study by Johnson and Sumpter (2001) revealed that the essential sewage treatment can rapidly convert organics into biomass that is then separated from the effluent phase by settlement. However, not all compounds were completely broken down or converted into biomass in that study. For example, the steroid estrogens found in effluent corresponded to the products of incomplete breakdown of their respective parent molecules. Liu et al. (2009) also reported that endocrine disruptor compounds were not completely removed by municipal wastewater treatment plants and remained at fluctuating concentrations in the effluent. Thus, the discharge of such water may be the main reason for the wide distribution and occurrence of emerging contaminants in surface waters. The lists of substances monitored by the treatment plants are beginning to evolve, which will eventually force the plants to upgrade their operations to better treat wastewater. There are two main policies for water pollution control: the first is to focus on reducing pollutant emissions at source, e.g. France, and the second is to focus on downstream reduction at treatment plants, e.g. in Switzerland and Germany (Morin-Crini and Crini 2017). In the latter case, tertiary treatment is required. This final treatment is necessary to remove the remaining pollutants such as

1  Remediation of Emerging Contaminants

9

pharmaceuticals and their metabolites. However, the use of tertiary treatment worldwide, including in Europe, is still limited due to significant economic costs. However, it may be necessary in the future if further restrictions are applied (Crini and Lichtfouse 2018a, b).

1.3  Methods for the Removal of Emerging Contaminants Removing emerging contaminants from wastewater is a difficult and challenging task because most emerging contaminants are recalcitrant substances, often present at trace amounts in complex mixtures. Indeed, a major factor complicating the cleanup of many sites is the coexistence of organic molecules and mineral substances such as metals and metalloids, the so-called mixed wastes. Another complicating factor is the wide range of properties of contaminants, with different chemical structures and functionalities. This diversity of water pollutants requires a wide range of treatment methods that must not only be effective, but also technologically and economically feasible. To treat emerging contaminants, a multitude of techniques classified as physicochemical treatments, biological treatments, membrane filtration, advanced oxidation technologies, adsorption/biosorption have been proposed in the literature (Fig.  1.2) (Crini and Badot 2007, 2010; Acero et  al. 2013; Charles et  al. 2014; Secondes et al. 2014; Serna-Galvis et al. 2016, 2019; Bedoya-Ríos and Lara-Borrero 2018; Botero-Coy et  al. 2018; Crini and Lichtfouse 2018a, b; Kim et  al. 2018).

Fig. 1.2  Methods used to remove emerging compounds. (Source: Juan Carlos Moreno-Piraján, Bogotá, Colombia)

10

N. Morin-Crini et al.

Advanced treatments such as biosorption onto non-conventional materials (Wang and Wang 2016; Li et al. 2019a; Fenyvesi et al. 2020; Rojas and Horcajada 2020), filtration membrane coupled with a biological step, algal-based technologies, and oxidation processes combined with an adsorption pretreatment (Serna-Galvis et al. 2016, 2019; Rodriguez-Narvaez et al. 2017) have been proposed to treat domestic water. Recent studies showed that the latter techniques could be a promising way, especially advanced oxidation processes such as ozonation, electrochemistry, photocatalysis, solar-driven processes or solar photo-assisted H2O2 treatment (Polo-­ López et al. 2014; Secondes et al. 2014; Serna-Galvis et al. 2016, 2019). Indeed, advanced oxidation processes have been demonstrated to be very efficient in decreasing not only organic substances but also the pathogen load in contaminated water (Moreira et al. 2016). This is of special interest because of properly treated wastewater could be an alternative renewable water resource, for example for agriculture as an end-use point. All these techniques can also be considered for treating industrial water (Wang and Wang 2016; Crini et al. 2019).

1.4  A  dsorption-Oriented Processes for the Removal of Emerging Contaminants 1.4.1  A  dsorption of Emerging Contaminants on Activated Carbons Liquid-solid adsorption/sorption is a key method for removing contaminants, particularly organics, from wastewaters (industrial effluents) or drinking water sources. Indeed, this technique of contaminant removal can produce high quality water, while also being a process that is both technologically simple and economically feasible (Liu and Liptak 2000; Yang 2003; Crini 2010). Liquid-solid adsorption/sorption involves the accumulation of adsorbate substances from an effluent onto the external and internal surfaces of a material/adsorbent used as a mass separating agent (Crini and Badot 2007, 2010; Fomina and Gadd 2014; Crini and Lichtfouse 2018a, b; Crini et al. 2019; de Andrade et al. 2018; Jeirani et al. 2017). This surface phenomenon is a result of complicated interactions among the three components involved, i.e. the adsorbent, the adsorbate and the effluent (water or wastewater). The data from the literature and also from the industry show that the control of adsorption performances of a solid material in liquid-­ phase adsorption depends on the following factors: (i) the origin and nature of the solid such as its physical structure, chemical nature and functional groups; (ii) the activation conditions of the raw solid (physical treatment, chemical modifications); (iii) the influence of process variables such as contact time, initial pollutant concentration, solid dosage and dynamic conditions; (iv) the chemistry of the contaminant (as example for an organic molecule, its pKa, polarity, size and functional groups); and finally, (v) the effluent conditions, referring to its pH, ionic strength, temperature and the presence of multi-contaminants and impurities.

1  Remediation of Emerging Contaminants

11

Carbon-based materials are one of the oldest and most widely used adsorbents/ sorbents, employed since the beginning of the 1900s in the treatment of drinking water (Liu and Liptak 2000; Dąbrowski et al. 2005; Crini and Badot 2007, 2010; Khalaf 2016). These adsorbents have excellent textural properties (high porosity, high specific surface area, large range of particle sizes) and physicochemical properties (surface chemistry) which together account for their excellent adsorption capacity. The process that uses carbons at industrial scale is often carried out by feeding the effluent continuously through a packed bed of carbon (Fig. 1.3). These fixed-bed systems have an important advantage because adsorption depends on the concentration of the substance(s) in the solution being treated. The adsorbent is continuously in contact with fresh effluent; hence, the concentration in the water in contact with a given layer of adsorbent in a column is relatively constant. The process is efficient and economical from the point of view of implementation and operation, which is relatively simple (Geankoplis 1993) and does not generate water by-products as in the case of chemical oxidation or advanced oxidation processes (Rossner et al. 2009; Ahmed et al. 2015; de Andrade et al. 2020), which activities and toxicities still require further studies (Zhang et al. 2016a). Activated carbons are used as very good adsorbents of a wide range of contaminants and contamination, e.g. to adsorb organic matter in order to reduce the organic load in secondary and tertiary treatment in a municipal wastewater treatment plant, to treat organics such as pesticides, aromatic and phenolic derivatives, volatile aromatics, hydrocarbons and surfactants, to decrease chemical oxygen demand and biological oxygen demand, to discolor water (dye removal) or to treat substances raw water packed bed

Lab scale

Industrial scale

treated water Fig. 1.3  Fixed-bed type process used for adsorption of contaminants from water or wastewater, e.g. for the post-treatment of a pharmaceutical effluent treated by chemical oxidation. (Source: Grégorio Crini, Besançon, France)

12

N. Morin-Crini et al.

that cause specific taste or smell (Crini and Badot 2007, 2010; Couto et al. 2015; Ferreira et al. 2015; Crini et al. 2019; Khan et al. 2020). The adsorption efficiency depends on the type of carbon used, the properties of the adsorbate, as well as the composition of the effluent being treated (Crini and Badot 2010; Ahmed et al. 2015). Over the past three decades, activated carbons, in granular (GAC) or powder (PAC) form, have been widely used in conjunction with microorganisms for treating wastewater. Macroporous activated carbons can be used as supports for bacteria: the bacteria then degrade part of the adsorbed organic matter (biological elimination) and thus participate in the in situ regeneration of the adsorbent. Two combined processes are principally used. Powder activated carbon treatment is obtained by adding carbon to activated sludge. Biological activated carbon (BAC) is a filter made of biofilm-covered granular activated carbon. Commercial activated carbon materials have been employed to remove organic pollutants such as pesticides from wastewater and water. In the purification of drinking water, granular activated carbon filters have commonly been used as a polishing step to remove potentially harmful trace organics. This type of treatment is, in general, coupled to an ozonation stage, further improving the performance of the process. Carbon biological filters are, for instance, used for the detoxification of organic effluent loaded with inorganic ions (removal of iron, manganese, and nitrate). To obtain cost-effective technology (especially in the field of water recycling), carbon powder is used in conjunction with an ultrafiltration membrane or with other techniques, such as oxidation. In Europe, active carbon competes favorably with nanofiltration and begins to replace oxidation with ozone, in particular to remove pesticides (Crini and Badot 2010). The major advantage of carbon adsorption is that the solid adsorbent can be easily separated from the treated liquid. This allows easy and flexible process operation. However, as it is non-selective, common innocuous organics may interfere with the removal of hazardous compounds. Active carbon technology presents other disadvantages. Active carbons are quite expensive (the higher the quality, the greater the cost). Different qualities of carbon are strongly related to the raw material used, as well as it depends on the carbonization conditions and the way in which activation is performed (physical or chemical). And yet, even though the high absorbing power of active carbons is widely recognized, there are some drawbacks, namely the need of disposal of spent carbons, their rapid saturation, and their regeneration. This regeneration step of saturated carbon is also expensive, not straightforward, and results in loss of the sorbent. Finally, the carbon-adsorption process can be limited in cases of ultra-low concentrations of contaminants (Putra et al. 2009; Lian et al. 2014; Ahmed et al. 2015; Khan et al. 2020). Various removal efficiencies of emerging contaminants by commercial activated carbons have been published, due to the intrinsic physical and chemical properties of carbons, and the different experimental conditions (pH, contact time, type of water) used (Adams et al. 2002; Westerhoff et al. 2005; Snyder et al. 2007; Foo and Hameed 2010; Rivera-Utrilla et  al. 2013; Nam et  al. 2014). Interestingly, some hydrophilic pharmaceuticals and personal care products were successfully trapped via adsorption, despite of their low hydrophobicities. In fact, the mechanisms of adsorption are not yet fully elucidated (Rivera-Utrilla et al. 2013; Nam et al. 2014; Jeirani et al. 2017).

1  Remediation of Emerging Contaminants

13

For example, Adams et  al. (2002) showed that pharmaceuticals such as sulfamethazine, trimethoprim, and carbatox were effectively removed (81–98%) from water using powdered activated carbon. In contrast, Westerhoff et al. (2005) reported the wide ranges of removals (20–69%) for the pollutants such as acetaminophen, caffeine, diclofenac, naproxen, sulfamethoxazole, and atrazine by applying powdered activated carbon. Snyder et al. (2007) also demonstrated that 30–78% of pharmaceuticals were removed from river water samples using powdered activated carbon. Yu et al. (2009) reported a comprehensive study regarding the adsorption of two pharmaceuticals (naproxen and carbamazepine) and an endocrine disrupting (nonylphenol) compound from a full-scale drinking water treatment plant by activated carbon in granular or powder form. The results showed that activated carbon form affected pollutant adsorption to carbon. Powder carbon would be most appropriate for achieving 90% of pharmaceuticals removal at a concentration of 500  ng/L, whereas granular carbon would be more suitable for nonylphenol. The authors explained this result by the fact that nonylphenol was much less impacted by fouling than the other two molecules. Mailler et al. (2016) tested the use of micro-grain activated carbon (CarboPlus® technology) in a fluidized bed to remove pharmaceuticals, hormones and personal care products contained in treated wastewater from a municipal wastewater treatment plant (Seine Centre Paris, Colombes, France; 240,000 m3/day corresponding to 9000,000 population equivalent). Large-scale experiments using a 5 m high reactor with a surface of 4 m2 (10 g or 20 g of carbon per m3, flow rate 1400 m3/day, contact time between 10 and 20 min) have shown that a carbon-based configuration as tertiary treatment substantially improved the overall quality of the discharge waters. Indeed, micro-grain activated carbons can further reduce the residual pollution still present in biologically treated waters: reduction of biological oxygen demand (38–45%), chemical oxygen demand (21–48%), and dissolved organic carbon (13–44%). Besides, removal of ammonium, nitrite (total elimination) and total suspended solids has also been detected. On the other hand, carbon filters were most effective in removing pharmaceutical compounds (52 molecules) and hormones (10 molecules) and other emerging substances (57 molecules) such as alkylphenols, bisphenol A, parabens, pesticides (23 molecules) and sweeteners (50 to >90%). Thirteen campaigns of analytical monitoring have been realized. Pharmaceuticals had a good affinity for micro-grain activated carbon and high (>60%) or very high (>80%) removals were observed for most of the quantified compounds (22 substances of the 32 quantified): for example, atenolol (concentration 448 ± 400 ng/L; reduction of 92–97%), carbamazepine (74  ±  65  ng/L; 80–94%), ciprofloxacin (184  ±  95  ng/L; 75–95%), diclofenac (1120  ±  1400  ng/L; 71–97%), oxazepam (239 ± 110 ng/L; 74–91%) and sulfamethoxazole (1430 ± 1450 ng/L; 56–83%). The ten remaining substances that were less eliminated (25–75%) include for example acetaminophen, ibuprofen and iothalamic acid. Personal care products, alkylphenols, artificial sweeteners, benzotriazole, bisphenol A, pesticides and perfluorinated acids have also a good affinity for micro-grain activated carbon, while phthalates are not or poorly eliminated. Micro-grain activated carbon allowed obtaining performances comparable to conventional powdered activated carbon at a same fresh

14

N. Morin-Crini et al.

activated carbon dose, even if slightly higher removals are observed with powdered activated carbon for several compounds. The differences are explained by the nature of micro-grain activated carbon, which is reactivated, and more likely by the injection of FeCl3 with powdered activated carbon. However, micro-grain activated carbon removes nitrites and total suspended solids, contrarily to powdered activated carbon. In addition, the micro-grain activated carbon configuration leads to several significant operational advantages over powdered activated carbon for a similar cost (≈ 1000–1200 €/ton), such as the ease of operation, the non-necessity to use FeCl3 and polymer and the reactivability of the adsorbent (no waste to handle). Mailler et al. (2016) concluded that micro-grain activated carbon was more suited than powdered activated carbon for wastewater treatment at large-scale applications. Nam et  al. (2014), studying the adsorption of pharmaceuticals, pesticides and endocrine disruptors, also reported that in addition to carbon form, carbon dose and contact time affected the performance. Increased dosage of activated carbon and increased contact time allowed greater opportunity for the adsorbent to interact with target emerging contaminants, and enhanced adsorption removal was attained. Modifying the adsorbent dosage and contact time can maximize the removal of micropollutants. However, a cost-effectiveness analysis is necessary before implementing these solutions in field facilities. In addition, the adsorption of hydrophilic contaminants, such as caffeine, acetaminophen, sulfamethoxazole, and sulfamethazine, was significantly affected by the pH, whereas low temperature decreased the adsorption of hydrophobic pollutants such as naproxen, diclofenac, 2,4-­dichlorophenoxyacetic acid, triclocarban, and atrazine. For example, removal of acetaminophen and sulfamethoxazole changed significantly due to electrostatic interactions between the adsorbent surface and the ionic species of these substances. The charge state of ionizable pollutants can affect micropollutants at certain pH levels. Low temperature can decrease the diffusion of lipophilic pollutants during the hydrophobic interaction with the adsorbent. A decrease in adsorption removal in surface water samples was observed and this decrease was more significant for hydrophobic than hydrophilic compounds. The decline in the adsorption capacity in surface water samples was caused by the competitive inhibition of dissolved organic matter with pollutants onto activated carbon. Nam et al. (2014) pointed out the fact that it was important to consider the effects of dissolved organic matter when treating pollutants in a water treatment plant. Katsigiannis et al. (2015), studying the removal of emerging contaminants from water by granular activated carbon columns, demonstrated that the process removed endocrine disrupters (bisphenol A, triclosan) more efficiently than pharmaceuticals (ibuprofen, naproxen, ketoprofen) at concentrations up to 2 μg/L.  In all experiments, the performance was enhanced by increasing the depth of the carbon bed. Among the target compounds, triclosan had the highest removal efficiency while ibuprofen was the least effectively removed. At a surface velocity of 3.1  m/h, removal rates after 5  days of continuous operation were 74.7% for bisphenol A, 86.7% for triclosan, 57.4% for ibuprofen, 65.6% for naproxen and 61.4% for ketoprofen for a bed depth of 8 cm, while virtually complete removal was achieved for all target compounds at a bed depth of 16 cm. The authors explained their results by

1  Remediation of Emerging Contaminants

15

correlating with the pKa and the log Kow values of the target substances. During the experiments, microbiological activity was also observed, which stabilized the breakthrough curves that significantly extended the column’s service life. Katsigiannis et al. (2015) concluded that activated carbon in the form of granular activated carbon columns can provide for the effective removal of target compounds in a municipal wastewater treatment plant. Patiño et  al. (2015) proposed the adsorption of emerging pollutants (1,8-­dichlorooctane, nalidixic acid and 2-(4-methylphenoxy)ethanol) on functionalized multiwall carbon nanotubes. Recently, the use of carbon nanotubes has gained interest for the removal of different pollutants in the aqueous streams. These materials are a new member in the carbon family, including single walled and multi-walled nanotubes distinguished by their number of layers. They had exceptional mechanical properties, excellent electronic properties, large surface specific area, and their structure can be further tuned locally by oxidation treatment or surfactant treatment. Indeed, carbon nanotubes allow the surface modification in order to maximize the selectivity of adsorption through families of compounds, helping to improve the desorption behavior. When combined with membrane filtration, these innovative adsorbents demonstrated excellent removal of pharmaceuticals and personal care products with removals up to 95% (Kurwadkar et  al. 2019). Other examples of results obtained using carbon nanotubes are described in the review published by Khan et al. (2020). Carbon nanotubes are considered to be promising adsorbents for water and wastewater treatment, although their high price is an important parameter to consider and their adsorption capacity depends strongly on both the hydrophobicity of the adsorbates and the morphology of the materials used. The materials can also be used in photocatalytic degradation processes in combination with a filtration step for the removal of endocrine disruptors (Patiño et al. 2015; Kurwadkar et al. 2019; Khan et al. 2020). How to reduce the cost of adsorbents and improve their adsorption performance, while at the same time seeking environmentally friendly materials, has been a preoccupation of the research community since the 2000s. The development of cheaper, effective and novel materials for pollutant removal is currently an active field of research, as shown by the numerous studies published every year. Research is divided into two directions: the use of natural, cheap and abundant non-­conventional materials in raw or modified forms and the development of conventional/commercial and new (synthetic) materials. Several various studies have been carried out to replace active carbon by other adsorbent/biosorbent materials (Aksu 2005; Crini 2005, 2006, 2010, 2015; Ahmaruzzaman 2008; Vijayaraghavan and Yun 2008; Crini and Badot 2010; Ali 2012; Fomina and Gadd 2014; Kyzas and Kostoglou 2014; Couto et  al. 2015; Ferreira et al. 2015). Commercial materials adopted at industrial scale for potable water treatment, effluent treatment and decolorizing applications include zeolites, activated alumina, and silica gels (Crini and Lichtfouse 2018a, b; Crini et al. 2019). Other conventional materials are commercial ion-exchange resins and synthetic organic resins. Non-conventional adsorbents/sorbents include clays (e.g. bentonite), siliceous materials (e.g. perlite, alunite), waste materials from agriculture (e.g. rice

16

N. Morin-Crini et al.

husk, biochar) or forest industries (e.g. bark, sawdust), by-products from industry (e.g. red mud, sludge), coal derived from non-conventional resources, biosorbents (e.g. peat, chitosan, alginate, cyclodextrin polymers), aquatic macrophytes, and biomass (Adriano et al. 2005; Crini 2005, 2006; Bekçi et al. 2006, 2007; Kang et al. 2010; Braga et al. 2011; Liu et al. 2012a, b; Lian et al. 2014; Xie et al. 2014; Freitas et  al. 2017; Stofela et  al. 2017; Portinho et  al. 2017; Maia et  al. 2017, 2019; de Andrade et al. 2018; de Souza et al. 2018, 2019; Osei et al. 2019; Fenyvesi et al. 2020; Macedo et al. 2020). Which is the best adsorbent to replace activated carbon? There is no direct answer to this question because each adsorbent has advantages and drawbacks (Morin-Crini and Crini 2017; de Andrade et al. 2018). It is important to note that all non-conventional materials are basically at the laboratory stage in spite of unquestionable progress. Nevertheless, they will undoubtedly be at the centre of some extremely profitable commercial activities in the future (Morin-Crini and Crini 2017; de Andrade et al. 2018). Examples of such materials are described below.

1.4.2  R  emoval of Hazardous Contaminants Using Low-Cost Clay-Based Adsorbents Clays are natural well-known minerals having layered structures with large surface area, high porosity, and high chemical and mechanical stability (Beall 2003; Crini 2006; Bhattacharyya and Gupta 2008; Crini 2010; Lofrano 2012; Gupta and Bhattacharyya 2014; Zhu et al. 2016). In addition, these materials are low-cost and abundant on most continents and they are studied as alternatives to activated carbons (Mercurio et al. 2019). Low-cost is naturally an important advantage of clays as adsorbents, e.g. the cost of montmorillonite can be up to 20 times cheaper than activated carbons (de Andrade et  al. 2018). The performance of a clay material depends mainly on its chemical nature and pore structure. Clays are considered as strong candidates for ion-exchange and as host materials with a strong capacity to adsorb neutral and charged species (Bhattacharyya and Gupta 2008; Crini 2006). In general, pollutant removal efficiencies by clays are comparable or even higher than by commercial carbons. So, their use in adsorption-oriented processes for water purification and wastewater decontamination has become commonplace in industry today. The first recognition of their ability to sorb pesticides such as 2,4-­ dichlorophenoxyacetic acid (herbicide 2,4-D) from aqueous solutions was established in the 1970s (Beall 2003). Other practical applications of clays include purification of air, separation, sterilization and disinfection. Currently, there is great interest in the adsorptive processes using clays that are considered promising and effective in the environmental remediation of emerging contaminants. As effective adsorbents, clays interact with recalcitrant dyes used in cosmetics (Almeida et al. 2009; Rehman et al. 2013), pharmaceuticals (Maia et al. 2017) such as antibiotics, antihypertensives, anti-inflammatory drugs, beta-­blockers, and other psychoactive substances (caffeine), and pesticides (Grundgeiger et al. 2015).

1  Remediation of Emerging Contaminants

17

Trimethoprim is an antibacterial agent commonly used in human and veterinary medicine worldwide, often prescribed to be taken in combination with a sulphonamide. This substance was detected at levels between 0.6 and 7.6 μg/L in hospital sewage water in Sweden (Lindberg et al. 2004), at 0.16 μg/L in wastewater effluents from Holland (Batt and Aga 2005), between 40 and 705 ng/L in municipal wastewater treatment plants in US (Renew and Huang 2004), and ranging from 0.013 to 0.15 μg/L in US streams (Kolpin et al. 2002). Bekçi et al. (2006, 2007), studying trimethoprim adsorption onto montmorillonite, demonstrated that clay minerals were efficient to remove drugs. The performance was pH dependent and the maximum adsorption at the equilibrium was observed at pH  5.04. The modelling of adsorption results showed that the adsorption mechanism was a combination of two main interactions, i.e. physisorption and ion-exchange. Later, similar conclusions were reported by Molu and Yurdakoç (2010). Clay minerals has been proposed by more authors to sorb antibiotics (Parolo et al. 2008; Putra et al. 2009; Wang et al. 2011; Genç et al. 2013; Roca Jalil et al. 2015; Maia et  al. 2017). For example, Roca Jalil et  al. (2015) demonstrated that montmorillonite was a suitable adsorbent for ciprofloxacin, an antibiotic used in human and veterinary medicine. Its presence in hospital effluents and in domestic wastewater has been reported by Karthikeyan and Meyer (2006) and by Githinji et al. (2011), respectively. The highest performances were obtained at pH values lower than 10 (Roca Jalil et al. 2015). The adsorption capacities (up to 330 mg/g) were higher than that obtained onto commercial activated carbons (112  mg/g) (Wang et al. 2011; Genç et al. 2013). However, the ciprofloxacin aqueous solubility must be taken into account in adsorption tests. The ciprofloxacin adsorption mechanisms on montmorillonite depend mainly on pH (Roca Jalil et al. 2015). The adsorption studies of antibiotics on clays have showed that the molecular structure of the pharmaceutical had a direct influence on the adsorption capacity. This can be explained by the fact that antibiotics have molecules with protonable groups, dependent of the medium pH, affecting the interactions between the antibiotic species available at each pH and the adsorption sites on the surface clay mineral. Roca Jalil et  al. (2015) also showed that the adsorption mechanism was complex, due to a combination of several interactions including ion-exchange, electrostatic interactions, surface adsorption, hydrogen bonding and hydrophobic interactions, ionic interactions being the main interactions. Different clay-based materials were evaluated by Maia et al. (2017, 2019) for the removal of amoxicillin (antibiotic), propranolol (beta-blocker) and caffeine (stimulant). The authors demonstrated that the removal efficiency depended on the drug and material used and varied between 23 and 98% for amoxicillin, 21 and 89% for caffeine, and 29 and 100% for propranolol. High adsorption capacities were also described by Vieira and collaborators (de Andrade et al. 2018; Oliveira et al. 2018, 2019) for caffeine adsorption onto bentonite. Caffeine is a natural stimulant largely consumed, recalcitrant and persistent to conventional water treatments, and continuously released into the environment. Even when present in water at low concentrations, caffeine can negatively affect the metabolism of fish, amphibians, and reptiles (Santos-Silva et  al. 2018). Oliveira et  al. (2018, 2019) indicated that caffeine

18

N. Morin-Crini et al.

adsorption onto bentonite was due to a combination of physical and ion-exchange phenomena, with significant changes on the bentonite surface after adsorption. These changes were highlighted using several techniques (scanning electron microscopy, mercury porosimetry, helium pycnometry and nitrogen physisorption). Diclofenac is a nonsteroidal anti-inflammatory drug and due to its structure, the molecule presents low biodegradability and is recalcitrant to conventional sewage treatment plants. Diclofenac is widely detected in wastewater (Radjenovic et  al. 2007) and in European surface waters (Jux et al. 2002; Fernández et al. 2010). In the environment, this drug has toxic effects on bacteria, invertebrates and algae. Maia et al. (2019) studied the adsorption of diclofenac sodium onto a commercial organoclay (Spectrogel®) and showed that the interaction between diclofenac and the adsorbent was spontaneous and the process was endothermic. High performances were obtained in terms of abatement. A detailed characterization analysis demonstrated that adsorption takes place on the surface of the clay, without significant modifications. The authors concluded that adsorption onto Spectrogel® is a promising alternative for the removal of pharmaceuticals (Maia et al. 2019). Other examples of adsorption of pharmaceuticals from water and wastewater using non-conventional low-cost materials can be found in a comprehensive review published by de Andrade et al. (2018). Clay minerals were proposed to sorb pesticides (Polatti et al. 2006; Roca Jalil et al. 2013; Grundgeiger et al. 2015; de Souza et al. 2019). de Souza et al. (2019) studied adsorption of the herbicide 2,4-D from aqueous solution using an organo-­ modified bentonite clay. The maximum adsorption capacity for this herbicide obtained experimentally was 50.36 mg/g, found at a temperature of 298 K, being higher than other materials reported in the literature. Using a detailed thermodynamic study, the authors showed that the adsorption of 2,4-D in organophilic clays referred to a spontaneous, exothermal process of physical nature. The adsorbent can be easily regenerated when subjected to eluents such as mixtures containing fractions of ethanol/water (desorption = 95%). Another typical example is atrazine, an herbicide that has been widely used in agriculture around the World for the last 50 years. It is now banned in Europe but is still in use in other countries. The US Environmental Protection Agency has set the maximum contamination limit (MCL) for atrazine at 0.003μg/L. Grundgeiger et al. (2015) proposed organo-beidellites for atrazine adsorption, with the main mechanism being cation exchange. Increasing the adsorption capacity of clays is a critical objective in the environmental application of these materials for the removal of negatively charged organic contaminants and persistent non-polar and hydrophobic contaminants. Thus, thermal or chemical modifications (e.g. acid washing using HCl or H2SO4) have been proposed to alter the structure of clays and their physicochemical properties (Zhu et  al. 2016; de Andrade et  al. 2018). In general, modified materials present an enhanced porous structure, higher specific surface area, and higher surface acidity. These modifications may enhance the capacity of minerals in uptaking specific contaminants. Table  1.1 shows values for maximum drug adsorption capacity using clays treated by different methods.

1  Remediation of Emerging Contaminants

19

Table 1.1  Maximum drug adsorption capacity by the Langmuir isotherm, Qm (mg/g), obtained under certain experimental conditions of temperature, pH, agitation, equilibrium time and concentration of clay treated by different methods Adsorbent Commercial montmorillonite

Modification Acid treatment

Commercial montmorillonite K10 Bentonite – type Verde Lodo Commercial montmorillonite Mizulite Exfoliated vermiculite

Acid treatment Calcined

Operating Drug conditions Qm 129.51 Trimethoprim 30 °C; pH 3; 150 rpm; 7 h; 4 g/L Trimethoprim 25 °C; 150 rpm; 75.66 6 h; 4 g/L Caffeine 30 °C 40.53 Caffeine

25 °C

Thermal treatment

Gemfibrozil

Lightweight expanded clay aggregates

Thermal treatment

Gemfibrozil

Exfoliated vermiculite

Thermal treatment

Mefenamic acid

Lightweight expanded clay aggregates

Thermal treatment

Mefenamic acid

Exfoliated vermiculite

Thermal treatment

Naproxen

Lightweight expanded clay aggregates

Thermal treatment

Naproxen

Na-montmorillonite

In natura

Tetracycline

Na-montmorillonite

TMAa

Tetracycline

Na-montmorillonite

DDTMAb

Tetracycline

Na-montmorillonite

HDTMAc

Tetracycline

20 °C; pH 7; without agitation; 72 h; 130 g/L 20 °C; pH 7; without agitation; 144 h; 1000 g/L 20 °C; pH 7; without agitation; 72 h; 130 g/L 20 °C; pH 7; without agitation; 144 h; 1000 g/L 20 °C; pH 7; without agitation; 144 h; 1000 g/L 20 °C; pH 7; without agitation; 144 h; 1000 g/L 25 °C; pH 5.5; 24 h; 0.133 g/L 25 °C; pH 5.5; 24 h; 0.133 g/L 25 °C; pH 5.5; 24 h; 0.133 g/L 25 °C; pH 5.5; 24 h; 0.133 g/L

References Bekçi et al. (2006)

Bekçi et al. (2007) Oliveira et al. (2019) 67.19 Shiono et al. (2017) 1.05 Dordio et al. (2016)

0.02 Dordio et al. (2016)

2.02 Dordio et al. (2016)

0.02 Dordio et al. (2016)

0.07 Dordio et al. (2016)

0.02 Dordio et al. (2016)

341.77 Liu et al. (2012a) 555.54 Liu et al. (2012a) 888.87 Liu et al. (2012a) 740.43 Liu et al. (2012a) (continued)

20

N. Morin-Crini et al.

Table 1.1 (continued) Adsorbent Bentonite

Modification HDTMAc

Drug Amoxicillin

Montmorillonite

BDTMAd

Diclofenac

Operating conditions Qm 30 °C, 120 min; 27.85 pH 7; 250 rpm; 4 g/L 25 °C; pH 6.5 59.8

Montmorillonite

HDTMAc

Diclofenac

25 °C; pH 6.5

Commercial sentonite Spectrogel® – type C

DDMAe

Diclofenac

30 °C; 250 rpm 24 h; 10 g/L

References Zha et al. (2013)

de Oliveira et al. (2020) 54.4 de Oliveira et al. (2020) 36.58 Maia et al. (2019)

Organophilized with tetramethylammonium Organophilized with dodecyltrimethylammonium c Organophilized with hexadecyltrimethylammonium d Organophilized with benzyl decyltrimethylammonium e Organophilized with dialkyl dimethylammonium a

b

To conclude, from an industrial point of view, adsorption using clays is technologically simple and economically feasible, plus this represents a process that produces high quality water. This treatment has proved to be efficient in removing emerging contaminants that are non-biodegradable or chemically stable. Research, both fundamental and applied, is currently very active concerning the development of novel classes of materials, as well as the understanding of the mechanisms of activation, adsorption and regeneration. Nevertheless, studies on the adsorption of emerging contaminants, with a focus on continuous multicomponent systems at pilot scale, as well as in-depth technical-economic feasibility and life-cycle analysis studies are still scarce in the literature, making the applicability of these adsorbents still difficult.

1.4.3  R  emoval of Pharmaceuticals Using Cyclodextrin Bead Polymer Over the past two decades, much work has been done on the use of cyclodextrin polymers for the manufacture of complex emerging substances for environmental applications. These commercial polymers are becoming very popular for the tracking and removal of trace emerging substances. Conventional water treatment technologies are not very effective in reducing the concentration of these pollutants to a desirable level, prompting researchers to innovate and to propose complementary methods to conventional treatments. Inclusion of Emerging Contaminants by Cyclodextrin  Most of the pharmaceuticals interact with cyclodextrins. These special carbohydrates, produced from starch and characterized by cyclic structure with hydrophilic surface and moderately

1  Remediation of Emerging Contaminants

21

Fig. 1.4  Scheme of inclusion complex formation of cyclodextrin with a drug. (Source: Éva Fenyvesi, Budapest, Hungary; Marc Fourmentin, Dunkerque, France)

hydrophobic inner cavity, are able to include other compounds in their cavity. Since their discovery at the end of the nineteenth century in France, a great body of knowledge has been collected on the effect and mechanism of binding various organics also numerous pharmaceuticals among them (Crini 2014). Within the cavity of cyclodextrins, the water molecules are readily replaced by less hydrophilic guest compounds such as drugs of low water solubility (Fig. 1.4). In this way, inclusion complexes or clathrates are formed without creating or disrupting covalent bonds. The host cyclodextrin and the guest are reversibly connected by hydrophobic interactions, van der Walls forces, etc. therefore the complex can dissociate when the conditions change, e.g. upon dilution, heating or in the presence of other competing guests (Szejtli 1998; Crini et al. 2018). The stability of the inclusion complexes can be enhanced by hydrogen bonds between the hydroxyl groups on outer surface of cyclodextrin and the hydrophilic part of the guest protruding from the cavity. The most important consequences of the inclusion complex formation, such as ­enhancing the solubility of poorly soluble substances, protecting the included guest molecules against the environmental effects such as oxidation, hydrolysis, decomposition on heat and light, enzymatic degradation, and improving the taste, are utilized broadly by the pharmaceutical industry (Frömming and Szejtli 1994; Loftsson et al. 2005; Szejtli and Szente 2005; Uekama et  al. 2006). The number of active ingredients marketed in the form of cyclodextrin complexes is above 80 at the beginning of 2020 and is continuously increasing (CycloLab 2020). Among the three natural cyclodextrins, α-cyclodextrin, β-cyclodextrin and γ-cyclodextrin consisting of 6, 7 and 8 glucopyranose units, respectively, β-cyclodextrin possesses a cavity size with an internal diameter of 0.60–0.65 nm and a depth of 0.78 nm especially suitable for inclusion of a wide range of drug molecules (Szejtli 1998). While the unmodified β-cyclodextrin is used mostly as excipient in pharmaceutical formulations for oral administration, its highly soluble hydroxypropylated and sulfobutylated derivatives have also been approved for parenteral applications (EMA 2017). Cyclodextrin Bead Polymers  While the pharmaceutical industry aims to improve the solubility and consequently the bioavailability of drugs via complexation, the objective in wastewater purification is the opposite: to remove the dissolved non-­ biodegradable pharmaceuticals by sorption/biosorption (both terms are used in the

22

N. Morin-Crini et al.

Fig. 1.5  Structure of a cyclodextrin polymer obtained by a cross-linking reaction between cyclodextrin molecules and a cross-linking agent. (Source: Marc Fourmentin, Dunkerque, France; Éva Fenyvesi, Budapest, Hungary; Grégorio Crini, Besançon, France)

literature). As natural cyclodextrins are water-soluble, they must be transformed into water-insoluble sorbents/biosorbents, using three main synthesis routes: 1. by cross-linking with proper bi- or polyfunctional reagents, such as diepoxy compounds, diisocyanates, di- or polycarboxylic acids, fluoroterphthalonitriles, etc. (Wiedenhof et al. 1969; Crini et al. 1998; Yamasaki et al. 2008; Zhao et al. 2009; Alsbaiee et al. 2016; Xu et al. 2019; Yang et al. 2020); Fig. 1.5 shows a cylodextrin polymer obtained by a cross-linking reaction between cyclodextrin molecules and a cross-linking agent; 2. by copolymerization of polymerizable cyclodextrin derivatives with acrylic or vinyl monomers (Wimmer et al. 1992; Janus et al. 1999; He et al. 2012); 3. or by grafting on a macromolecular support such as chitosan, cellulose, alginate and silica (Crini et al. 1995; Sakairi et al. 1999; Aoki et al. 2007; Chung and Chen 2009; Kono et al. 2013; Omtvedt et al. 2019; Yamasaki et al. 2017). All these gels are preferably prepared in the form of tiny beads by emulsion/ suspension polymerization. The spherical shape of the particles has the advantage of high surface area and good accessibility of the cyclodextrin cavities in addition to the technological advantages in scaled up production. Owing to the several hydroxyl groups on cyclodextrins, these gels are highly hydrophilic and swell in water. The degree of swelling depends on the conditions of preparation, e.g. type, concentration and ratio of reactants, temperature, reaction time, etc. (Wiedenhof et al. 1969; Fenyvesi et al. 1979; Romo et al. 2006). The higher the degree of swelling the higher is the accessibility of the inner cyclodextrin cavities but the smaller the mechanical stability against compression. The binding capacity of epichlorohydrin-­cross-linked β-cyclodextrin polymer beads is higher in case of

1  Remediation of Emerging Contaminants

23

lower degree of cross-linking and reduced temperature (Zhu et al. 1997; Baille et al. 2000). In case of ionic or ionizable pharmaceuticals cyclodextrin polymers modified by ionic groups can be more efficient biosorbents due to the contribution of electrostatic forces to hydrophobic interactions. Depending on the conditions the binding mechanism can be complicated including inclusion and association complex formation, electrostatic interactions, ion exchange, chelation, etc. (Crini et al. 2016; Morin-Crini et al. 2018). Both the π–π interactions with the network and host-­ guest interactions with the cyclodextrin cavities may contribute to the uptake (Huang et al. 2020). Removal of Emerging contaminants by Cyclodextrin Bead Polymer Using Biosorption-Oriented Processes  Cyclodextrins and their polymers have been found useful for various environmental applications including wastewater treatment (Crini and Morcellet 2002; Gruiz et al. 2011; Landy et al. 2012; Morin-Crini and Crini  2013; Morin-Crini et  al. 2018;  Fenyvesi et  al. 2019). The early sorption/biosorption experiments with cyclodextrin polymer beads aimed at the removal of various organic contaminants such as phenols and dyes. Although these beads usually outperformed activated carbon, it became clear that cyclodextrin polymer beads are too expensive for the removal of compounds present at high concentration. Due to their specific affinity toward drugs (Fenyvesi et  al. 1996; Vyas et al. 2008; Mocanu et al. 2009), they found their application for the sorption of non-biodegradable emerging contaminants, such as pharmaceuticals, which survive the traditional wastewater treatment and are usually detected in purified wastewater at very low concentrations. The epichlorohydrin-cross-linked cyclodextrin polymers successfully removed hormones and other endocrine disrupting chemicals at nanomolar concentrations (Oishi and Moriuchi 2010). High removal rate (70% to >90%) was achieved for 17-β-estradiol, a contraceptive of emerging concern, in the range of 1  ×  10−11 to 1 × 10−8 mol/L concentration by using β-cyclodextrin polymer. Similar performance was observed for polymers prepared from γ-cyclodextrin and α-cyclodextrin. The removal ratio was hardly reduced when in addition to 17-β-estradiol the model solution was spiked also with cholesterol in 100-fold molar excess related to the hormone because cholesterol, another steroid typical in wastewaters but with no endocrine disrupting properties, has lower binding affinity to the cyclodextrin cavity. The small scale laboratory batch experiments with real municipal wastewater validated these results. It was recently published that carbonyldiimidazol-cross-­ linked β-cyclodextrin and γ-cyclodextrin polymer nanosponges efficiently removed the antipsychotic drug pimavanserin from single solutions (93% and 80% removal rates, respectively) and from postreaction raffinates in small scale batch experiments (0.1 g sorbent in 5 mL solution) (Hemine et al. 2020). The low mechanical strength of the gel beads does not allow the use of column technique in large scale. The higher swelling, e.g. lower degree of cross-linking, results in lower permeability, and thus lower elution rate of water through the gel bed is attained. Therefore fluidization technique was used in up-scaled demonstration of the technology. Another option is applying inorganic core such as silica or

24

N. Morin-Crini et al.

graphene oxide to get beads of high mechanical stability or introduction of rigid structures such as phenyl moieties into the polymer to form pores. For instance, silica coated with β-cyclodextrin using hexamethyl diisocyanate as cross-linking agent was used for the removal of emerging contaminants, also estrogen hormones among them from water (Bhattarai et al. 2012). More than 95% of 17-β-estradiol was removed from single component solution and more than 90% of most of the estrogens in multicomponent system even after 4 regeneration cycles. Also magnetic graphene oxide modified with β-cyclodextrin/poly(l-glutamic acid) showed excellent binding capacity for 17-β-estradiol (Jiang et al. 2016): the sorption capacity was 85  mg/g sorbent at 0.05  g/L sorbent to solution ratio and 25  °C using a model solution of ~7 × 10−7 mol 17-β-estradiol concentration. Single component solutions were eluted through a thin layer of β-cyclodextrin polymer with permanent porosity (cross-linked with tetrafluoroterephthalonitrile) to remove around 80% of ethinylestradiol and propranolol as model drugs at 1 × 10−4 mol/L concentration (Aisbaiee et al. 2016). In another study, a solution (8 mL) of 90 pollutants was eluted through the same porous β-cyclodextrin polymer (1 or 5 mg) to achieve nearly complete (over 95%) removal of several drugs including abacavir, amphetamine, cimetidine, codein, diclofenac, estrone, famotidine, fluoxetine, gemfibrozil, ibuprofen, testosterone, triclosan to mention only a few and several pesticides (Ling et  al. 2017). Figure  1.6 illustrates the diclofenac biosorption onto a cyclodextrin polymer. More recently, using the same cross-linking agent, the polymer (100 mg) obtained removed over 95% of both ethinylestradiol and estriol (~1 × 10−7 mol/L, 1 L) by filtration from a multicomponent model solution containing other emerging contaminants, too (Xu et al. 2019). A further possibility for preparing filters of good mechanical stability is sintering cyclodextrin polymer gel beads with a thermoplastic polymer such as polyethylene (Jurecska et  al. 2014). This process, however, results in reduced binding capacity. In laboratory experiments modeling post-purification of wastewater by fluidization technology, epichlorohydrin-cross-linked β-cyclodextrin polymer beads (20 g) were used to remove hormones, such as β-estradiol, ethinylestradiol and estriol

Fig. 1.6  Schematic illustration of the diclofenac sorption onto a cyclodextrin polymer. (Source: Marc Fourmentin, Dunkerque, France; Éva Fenyvesi, Budapest, Hungary; Grégorio Crini, Besançon, France)

1  Remediation of Emerging Contaminants

25

Fig. 1.7  Flow chart of wastewater purification technology applying cyclodextrin polymer beads for sorption of the emerging pollutants and micropollutants such as residual pharmaceuticals. (Source: Éva Fenyvesi, Budapest, Hungary)

(87–99%) from the spiked test solution (800 mL) containing some other emerging contaminants, too. The removal rate for the non-steroidal anti-inflammatory drugs was in the range of 15–70% (Nagy et al. 2014). Based on the positive results of this experiment, a scaled-up fluidization procedure was planned and performed in a pilot-scale wastewater purification plant (Fig. 1.7). β-Cyclodextrin polymer beads (1 kg) were applied to polish 300 L purified municipal wastewater spiked with 9 emerging pollutants at approximately 2 × 10−8 mol/L concentration (among others also drugs, e.g. diclofenac and 17-β-estradiol included in the watch list of substances for Union-wide monitoring in the field of water policy (European Water Framework Directive 2015/495/EU) (Fenyvesi et al. 2020). The hormones such as estradiol (>99.9%), ethinylestradiol (99.9%) and estriol (96.1%) were effectively removed. The removal rate for the non-steroidal anti-inflammatory drugs diclofenac, ibuprofen, ketoprofen and naproxen were 85.5, 86.9, 18.1 and 13.5%, respectively, following roughly the cyclodextrin-drug association constants showing that the main interaction was likely the inclusion complex formation. It should be noted that similarly low removal rate (18%) was obtained by Orprecio and Evans (2003) for naproxen applied at 5 × 10−6 mol/L concentration in single component solution (5 L) using similar epichlorohydrin-cross-linked β-cyclodextrin polymer (10 g) as column packing. The pilot-scale demonstration of the fluidization technology with the β-cyclodextrin polymer beads was successful, most of the pharmaceuticals as emerging contaminants, especially those on the European watch list, were efficiently removed in a short time. The beads are easily regenerated by extracting the sorbed components with methanol or ethanol (Orprecio and Evans 2003; Jurecska et  al. 2014; Fenyvesi et  al. 2020; Hemine et  al. 2020). The recovery of valuable drugs from industrial effluents is also conceivable after regeneration of the cyclodextrin polymer sorbent (Hemine et al. 2020).

26

N. Morin-Crini et al.

For the full scale application of the cyclodextrin-based biosorbents in wastewater treatment the first step is the scaled up production of these polymers. All the examples overviewed in this section used sorbents/biosorbents prepared in laboratory except the epichlorohydrin-cross-linked β-cyclodextrin polymer beads produced on pilot-plant scale. Several research groups are working on development of economic and environmental-friendly technologies for scaling up. The tetrafluoroterephthalonitrile-­cross-linked Dexsorb® adsorbents of Cyclopure Company, which can quickly and safely remove from water hundreds of pollutants including drugs, pesticides and short and long chain polyfluorinated alkyl substances, are probably close to the industrial production (Cyclopure 2020). All of these studies show that cyclodextrin polymers are a promising alternative in the treatment of wastewater containing a cocktail of emerging trace substances. Industrialists will now have to be convinced to use these materials in their municipal wastewater treatment plants.

1.4.4  M  etal-Organic Frameworks for the Removal of Emerging Contaminants In recent years, metal-organic frameworks have attracted attention as promising materials for the removal of emerging contaminants in effluents by adsorption-­ oriented and catalytic degradation processes (Dias and Petit 2015; de Andrade et al. 2018; Mon et al. 2018; Bedia et al. 2019; Li et al. 2019b; Dhaka et al. 2019; Rojas and Horcajada 2020). Metal-organic frameworks are a kind of crystalline materials with permanent porosity (more than 50% of their crystal volume) formed by the self-assembly of metal ions or their aggregates/clusters linked together by multifunctional organic ligands. They consist of small organic molecules, usually C6-rings, which are connected by metal clusters as a three-dimensional structure. Like cyclodextrin molecules, metal-organic frameworks have cavities where specific host-guest interactions can take place. These active sites can adsorb and degrade pollutants. In addition to superhigh pore volume, metal-organic frameworks also have values of surface area (1000–10,000  m2/g) superior to those of conventional materials such as carbons. They have also physicochemical properties easily tuned. Reviews by Dias and Petit (2015) and Bedia et al. (2019) describe the different methodologies that can be used for the synthesis of metal-organic frameworks, paying attention to the purification and activation steps. A typical example of a metal-organic framework is chromium terephthalate MIL-101 (Matériel Institut Lavoisier). It is a polymer built from trimeric chromium (III) triangular cluster complexes, bridged by linear terephthalate linkers. The material has a highly porous three-dimensional structure with large pores (>30 Å), high surface area (>3000 m2/g) and a huge cell volume (702,000 Å3). Because of their distinctive characteristics, metal-organic frameworks are very useful in chemistry and supramolecular chemistry (gas separation, adsorbent for

1  Remediation of Emerging Contaminants

27

carbon capture, hydrogen for energy storage), nanotechnology (electrocatalysis, photocatalysis, biocatalysis), materials science, drug delivery systems, biomedical imaging, sensing, fuel cells, and water and wastewater treatment (adsorption, semiconductor photocatalysis). They have many properties for the removal of pollutants that recommend them in water treatment. Indeed, metallo-organic frames have not only a high pore volume, large surface area and adjustable pore size, but also multiple topologies, hierarchical structure, adjustable surface chemistry (easily functionalized cavities), and high adsorption capacity (de Andrade et al. 2018; Rojas and Horcajada 2020). Their recyclability also gives metal-organic frameworks an advantage over conventional adsorbents. Metallo-organic frames can be synthesized on a large scale and are versatile materials as they can be shaped into monoliths, pellets, membranes or beads for columns, suitable for decontamination devices. An important challenge for using metal-organic frameworks as adsorbents in their stability in contact with water (de Andrade et al. 2018). Metal-organic frameworks can be used as adsorbent of pollutants or as platform for pollutant degradation through catalytic processes. This research is recent as the first work on metal-organic framework for dye removal and the first metal-organic framework used in the catalytic degradation of phenol were reported in 2010 by Haque et al. (2010) and in 2007 by Alvaro et al. (2007), respectively. In water and wastewater treatment, metal-organic frameworks have proven to be excellent adsorbents for the removal of harmful species such as pharmaceuticals and personal care products, artificial sweeteners and feed additives, agricultural products, organic dyes, BTEX compounds, pesticides, and industrial products such as alkylphenols, products of the photographic industry and plasticizers (Rojas and Horcajada 2020). Their performance is better than that of other adsorbents, especially activated carbons. However, their performance and selectivity with respect to the pollutants targeted for disposal must be regulated by judicious selection of the metal ion and organic linker. Numerous examples of applications in the field of water and wastewater treatment are presented and discussed in the following references: Dias and Petit (2015), de Andrade et al. (2018), Mon et al. (2018), Bedia et al. (2019), Li et al. (2019b), Dhaka et al. (2019), and Rojas and Horcajada (2020). The extensive data in the literature show that metal-organic frameworks, as a new class of porous materials, have a great potential for water purification by (selective) adsorption and catalytic degradation (photocatalysis) and could also find applications in other advanced oxidation processes (photo-Fenton, electro-Fenton), in soil remediation or in membrane filtration. This is a research topic in full development, particularly in the context of nanotechnologies, but is it is still in its early stages and requires a thorough assessment of parameters such as safety, lifetime, reusability and industrial conditions. In addition, the low stability of metal-organic frameworks in water is still a major challenge for their environmental application in industry (Rojas and Horcajada 2020). For a real application and compared to traditional semiconductors such as TiO2, two other aspects need to be improved. Firstly, the synthesis conditions must allow larger quantities of material to be obtained in continuous operation. Current processes, such as solvothermal methods, are limited by the slowness of the reactions, and have a relatively high cost (Li et al. 2019b).

28

N. Morin-Crini et al.

Second, research is also needed on their potential toxicity and possible health effects. Indeed, many metal-organic frameworks are constructed from toxic metals, e.g. Cd, Cr, Ag and Co, and some molecules such as the dyes used are also considered as emerging contaminants.

1.4.5  A  pplication of Molecularly Imprinted Polymers for Removal of Emerging Contaminants Over the past decade, several studies have reported innovative results on molecularly imprinted polymers and non-imprinted polymers for the removal of personal care products, pharmaceuticals, and endocrine disrupting compounds, in industrial effluents, surface waters, and drinking water (Murray and Örmeci 2012; Shen et al. 2013; Huang et al. 2015; Chen et al. 2016a). The molecular imprinting technique is an emerging technology in which the synthesis of a material is performed in the presence of a template molecule. Subsequent removal of the template provides a material with “memory” sites capable of selectively recognizing and re-binding to the original template of a mixture (Alexander et  al. 2006; Chen et  al. 2016a; Cantarella et al. 2017; BelBruno 2019). Molecularly imprinted polymers are prepared from cross-linked polymers containing cavities specific to a target analyte. These cavities are created by the copolymerization of cross-linking monomers and functional monomers with an imprinting molecule or template. After polymerization, the template is removed, leaving a cavity specific to the analyte. The molecularly imprinted polymer then selectively re-­ binds to the compound to be treated (Alexander et al. 2006; BelBruno 2019). The main advantages of the molecularly imprinted polymers are the ease of preparation and the ability to create “tailor-made” binding sites simply by adapting the synthesis procedure for the desired target molecule used as a template during polymerization (Huang et al. 2015; Cantarella et al. 2019). Molecularly imprinted polymers have traditionally been used as solid-phase extraction media for analytical chemistry, e.g. for pre-concentration and selective removal of substances (Chen et  al. 2016a; BelBruno 2019). Non-imprinted polymers are cross-linked polymeric materials that have macropores containing adsorption sites for organic molecules. They are synthesized using the same procedure as molecularly imprinted polymers, but in the absence of a template. They therefore have the same chemical properties as molecularly imprinted polymers but do not contain specific cavities (Murray and Örmeci 2012). Non-­ imprinted polymers exhibit strong hydrophobic interactions (non-specific binding) between organic pollutants and polymers. The main difference between molecularly imprinted polymers and non-imprinted polymers is their specificity. Molecularly imprinted polymers can selectively remove a target substance such as 17-β-estradiol from biological wastewater, whereas non-imprinted polymers can remove several substances simultaneously. Molecularly imprinted polymers are also generally

1  Remediation of Emerging Contaminants

29

more porous, having a greater surface area, which may also explain the higher binding efficiency of molecularly imprinted polymers. However, while the specificity of molecularly imprinted polymers is an advantage for solid-phase extraction application, the same may not be true for complex wastewater (BelBruno 2019). Molecularly imprinted polymers offer promising applications for water and wastewater treatment. They are advantageous for treatment of trace contaminants as they can be specifically designed to remove one or a group of target compounds (Alexander et al. 2006). This is an advantage over nonspecific technologies such as activated carbon (Pichon and Chapuis-Hugon 2008; Murray and Örmeci 2012). Molecularly imprinted polymers have been studied for their ability to remove, degrade and destroy, in combination with other technologies (advanced oxidation processes), emerging substances such as hormones (Le Noir et al. 2007; Zhongbo and Hu 2008; Chen et al. 2015), pharmaceuticals (Pichon and Chapuis-Hugon 2008; Madikizela et al. 2018; BelBruno 2019; Cantarella et al. 2019), endocrine disruptors (Fernández-Alvarez et al. 2009), personal care products (Alsudir et al. 2012) and pesticides (Murray and Örmeci 2012). Le Noir et  al. (2007), investigating the percolation of wastewater containing 17-β-estradiol through solid-phase extraction columns prepared with molecularly imprinted polymers (flow rate of 50  mL/min), demonstrated that these materials were capable to eliminate the equivalent of 22  ±  4  ng of 17-β-estradiol per L of effluent. The authors also showed that the use of molecularly imprinted polymers allows easy regeneration for subsequent reuse, yielding the same efficiency and reducing overall treatment costs. In another work, the same authors developed a method for regenerating molecularly imprinted polymer with a template of 17-β-estradiol using solvent extraction under ultraviolet light (Fernández-Alvarez et al. 2009). This method was able to regenerate and reuse both the polymeric material and the solvent, as well as the destruction of 17-β-estradiol. For that, acetone was used as a solvent under both ultraviolet-C and ultraviolet-visible light. After a 10-hour cycle, the molecularly imprinted polymer was completely regenerated and the acetone no longer contained residual 17-β-estradiol. Cantarella et  al. (2019) studied the adsorption of diclofenac to molecularly imprinted polymers synthetized by a simple and inexpensive bulk polymerization reaction. The authors also reported that materials can be regenerated and reused by simply washing with a 3 mL methanol/acetic acid solution in 2 min. The performance remained the same after at least four adsorption/regeneration cycles. Molecularly imprinted polymers showed exceptional adsorption capacity for the adsorption of diclofenac and the process was highly selective. In 10 min, 5 mg of molecularly imprinted polymer removed ∼90% of diclofenac from an aqueous solution at a concentration of 1 × 10−4 mol/L, while only 8% was removed by the corresponding non-imprinted polymer. The adsorption capacity was 110μmol/g and 13 μmol/g for the molecularly imprinted polymer and the non-imprinted polymer, respectively. Due to their easy and inexpensive synthesis, high efficiency and selectivity, easy regeneration and reusability, the authors concluded that the technology could be used on a large scale for water treatment.

30

N. Morin-Crini et al.

The use of molecularly imprinted polymers and non-imprinted polymers for the removal of emerging contaminants using adsorption-based or solid-phase extraction processes is a novel approach that has potential in the field of water engineering (BelBruno 2019). Polymers, particularly non-imprinted polymers, are inexpensive to manufacture and can be produced in large quantities (Alexander et  al. 2006). After use, the materials can be easily regenerated and reused several times. However, further research is needed to determine how best to integrate this technology into treatment plants. In addition, their potential toxicity and possible health effects are subject to discussion (Murray and Örmeci 2012).

1.4.6  A  dsorption of Emerging Contaminants onto Chitosan-Based Materials Like cyclodextrin polymers, chitosan-based materials have great potential in environmental applications, mainly for metal chelation and dye removal (Crini and Lichtfouse 2018a, b). Chitosan is a biopolymer derived from chitin. This aminopolysaccharide is capable of interacting and adsorbing a wide range of pollutants such as metals, dyes, organics and also emerging substances. In water and wastewater treatment, chitosan represents an alternative as an ecological complexing agent due to its low cost (chitin comes from shellfish wastes), its intrinsic characteristics (renewable, non-toxic and biodegradable resource, hydrophilicity) and its chemical properties (polyelectrolyte at acidic pH, high reactivity, coagulation, flocculation and biosorption properties) resulting from the presence of reactive hydroxyl and especially amine groups in the macromolecular chains. Its use in water treatment is justified by two other important advantages: firstly, its exceptional pollutant binding capacities and excellent selectivity, and secondly, its versatility (Crini and Lichtfouse 2018a, b; Morin-Crini et al. 2019). Chitosan-based materials can be used in solid form for the removal of pollutant from water and wastewater by filtration or adsorption processes or in liquid state, i.e. dissolved in acidic media, for applications in coagulation, flocculation, and membrane filtration technologies (polymer assisted ultrafiltration). Among the proposed materials, cross-linked chitosan hydrogels deserve particular attention (Morin-Crini et al. 2019). Chitosan-based materials have been investigated for the removal of emerging substances such as perfluorooctane sulfonate (Zhang et  al. 2011), bisphenol A (Dehghani et al. 2016; Zhou et al. 2019), amoxicillin (Adriano et al. 2005), sulfamethoxazole (Qin et al. 2015; Zhou et al. 2019), ciprofloxacin (Afzal et al. 2018), diclofenac (Rizzi et al. 2019), ketoprofen (Rizzi et al. 2019), and caffeine (Sanford et al. 2012). For example, Adriano et al. (2005) studied the adsorption of amoxicillin on chitosan beads. Amoxicillin ((2S, 5R, 6R)-6-[[(2R)-2-amino-2-(4-­ hydroxyphenyl)acetyl]amino]-3,3-dimethyl-7-oxo-4-thia-1-azabicyclo[3.2.0]heptane2-carboxylic acid) is an antibiotic that belongs to the beta-lactams groups. It is one of the most important commercial antibiotics due to its high bacterial resistance and

1  Remediation of Emerging Contaminants

31

large spectrum against a wide variety of microorganisms. More than 80% of orally administered amoxicillin in humans is excreted in the urine after 2 h of consumption. In conventional wastewater treatment systems, it is a difficult substance to eliminate due to its structure and amphoteric properties resulting from the presence of carboxylic (pKa = 2.68), amine (pKa = 7.49) and hydroxyl (pKa = 9.63) groups (Anastopoulos et al. 2020). Its charge changes gradually with pH, due to the different functional groups. The pH of wastewater affects not only the ionization of the molecules, but also the surface charge of the materials used as adsorbents for their removal. Adsorption results described by Adriano et al. (2005) showed that 0.5 g of cross-­linked chitosan were capable of removing amoxicillin from a 2 mL solution at concentrations ranging from 0.2 to 3 mg/L at pH = 6.5, with performances being concentration-dependent. The authors reported a maximum adsorption capacity of 8.71  ±  0.6  mg/g (equilibrium time 2  h). The adsorption mechanism was due to strong interactions between the carboxylic, amine and hydroxyl of the amoxicillin molecule and those of chitosan. From literature data, there is no doubt that chitosan-based hydrogels have high adsorption capacities. However, these materials are at the laboratory stage and pilot studies need to be conducted. Although various laboratories and a few companies can synthesize chitosan-materials to order, it is very difficult to find commercial sources of cross-linked hydrogels with guaranteed reproducible properties. However, the performance can vary depending on the conditions and the method the hydrogels are prepared. Despite an abundance of literature reports on industrial wastewaters, there is still little information available detailing comprehensive comparing various conventional commercial adsorbents under similar conditions. Like other non-conventional materials, further research is needed to determine how best to integrate this technology into treatment plants.

1.4.7  Nanocellulose as a Novel Adsorbent for Environmental Remediation Nanocelluloses are innovative materials with at least one dimension at the nanoscale that are attracting increasing interest in the field of nanotechnology and materials science. These materials are highly ordered β-(1→4) glucan chains that are produced naturally of by chemical processes. Nanocelluloses are mainly obtained from naturally occurring cellulose sources. Indeed, they can be obtained from biomass, plants or bacteria, using fairly simple, scalable and effective isolation techniques. Most nanocellulose is produced from lignocellulose. The term “nanocellulose” encompasses a number of cellulose-based materials, whose chemical and physical properties generally vary depending on their source and method of extraction: cellulose nanocrystals, nanofibrillated celluloses, and rigid bacterial nanocellulose (Moon et al. 2011; Lam et al. 2012).

32

N. Morin-Crini et al.

Pure nanocellulose is non-toxic, biodegradable and biocompatible. In addition, it has other advantages such as low cost, abundance, practically renewable, intrinsic properties (large surface area) and high reactivity (easy to modify), making it sustainable material for several applications including electronics, optoelectronics and engineering, paper industry, composites, antibacterial coatings, food packaging, cosmetics and personal hygiene products, medical applications (tissue scaffolds, drug delivery), bio-imaging, biosensors, enzyme immobilization, catalysis and energy storage and production (Putro et al. 2017; Thomas et al. 2018). Nanocellulose in its various forms (cellulose nanocrystals, cellulose nanofibrils, bacterial cellulose) is also promising material for environmental applications such as water treatment (Lam et al. 2012; Herrera-Morales et al. 2017, 2019; Mahfoudhi and Boufi 2017; Putro et al. 2017; Voisin et al. 2017; Abouzeid et al. 2018; Abujaber et  al. 2018; Mohammed et  al. 2018; Shak et  al. 2018; Wang 2019), air filtration (Nemoto et al. 2015), membrane filtration (Cruz-Tato et al. 2017; Abouzeid et al. 2018; Shak et al. 2018; Mautner 2020), flocculation (Shak et al. 2018) and catalytic degradation (Mahfoudhi and Boufi 2017; Thomas et  al. 2018; Shak et  al. 2018). Nanocelluloses have been used for the removal of metals, dyes, nitrates, organics and microbes from aqueous solution by filtration-, adsorption- or membrane-­ oriented processes. However, for industrial applications, native nanocelluloses have a low adsorption capacity due to their hydrophilic crystal structure. Therefore, it is necessary to incorporate functionalities by chemical modification of the cellulose primary hydroxyl groups to have suitable materials and composites suitable for hydrophobic contaminants. This functionalization is also necessary to increase the selectivity and affinity of the substances. Hokkanen et al. (2013) and Yu et al. (2013) reported that modification using succinic anhydrate and sodium bicarbonate can increase adsorption to nanocellulose up to tenfold. Shak et al. (2018) and reviewed the different strategies for the preparation of nanocellulose for applications in wastewater treatment. An important feature for an application in the water domain is the fact that nanocellulose materials are easily regenerated and reusable. Numerous studies have shown that nanocellulose-based materials are among the most promising green adsorbents, and comprehensive reviews have recently been published (Mahfoudhi and Boufi 2017; Putro et  al. 2017; Voisin et  al. 2017; Abouzeid et al. 2018; Mohammed et al. 2018; Shak et al. 2018; Wang 2019; Mautner 2020). Herrera-Morales et al. (2017, 2019) demonstrated that modified nanocelluloses were capable of effectively adsorbing pharmaceuticals such as sulfamethoxazole and acetaminophen (paracetamol). Abujaber et al. (2018) proposed magnetic cellulose nanoparticles electrostatically modified with ionic liquids to adsorb pharmaceuticals (paracetamol, ibuprofen, naproxen and diclofenac) in less than 30 min with extraction recoveries of 86%–16%. Although many studies have been published with promising results, nanocelluloses, and more generally nanomaterials, are still at the laboratory study stage (Shak et al. 2018).

1  Remediation of Emerging Contaminants

33

1.5  Biological-Based Technologies for the Degradation and Elimination of Emerging Contaminants The biological treatment approach, specifically the activated sludge process, is the most commonly used wastewater treatment technology. While it was originally designed to only remove organic carbon, it was subsequently extended to remove nitrogen and phosphorous. Conventional biological-based technologies are not capable of degrading the wide range of organic pollutants present in complex wastewater. These systems have been adapted and different strategies have been studied, including water pre-treatment, the implementation of complementary methods based on potabilization techniques such as ozonation and carbon adsorption, or pilot studies such as membrane filtration. These latter techniques, reserved for very specific water treatment, are already used in industry, but their cost is a brake on their development. Other techniques such as constructed wetlands, algal-based technologies and enzymatic degradation, and bioreactors are being developed.

1.5.1  Constructed Wetlands for the Removal of Emerging Contaminants Constructed wetlands are engineered systems based on the exploitation of plants for water remediation. These phytodepuration systems can be categorized according to the operational parameters, such as the type of flow (e.g. horizontal vs vertical, free water vs sub-surface), the hydraulic regime and the plant species. In every case, constructed wetlands take advantage of the numerous depuration processes that occur at the interface between the root system and the wastewater under treatment. These physical, chemical and biological processes are the same that may occur in natural wetlands and include adsorption on soil/sediment, volatilization, uptake and degradation (Gorito et  al. 2017). Constructed Wetlands are an environmental friendly, low-cost and easy to maintain technology suitable to remove total suspended solids, ammonia, phosphorous, and to reduce the chemical oxygen demand and biological oxygen demand. Thus, they are regarded as an interesting opportunity for wastewater treatment in developing countries (Mahmood et al. 2013) and for decentralized systems treating sewage produced by small communities (Zraunig et al. 2019). An increasing number of studies investigated the possible use of constructed wetlands for the removal of emerging organic contaminants including antibiotics and antibiotic resistance genes (Chen et al. 2016b; Huang et al. 2019), other pharmaceuticals and endocrine disruptors like bisphenol A (Syranidou et al. 2017; Meneghetti Campos et al. 2019) and synthetic dyes (Riva et al. 2019), highlighting the promising potential of phytodepuration also as tertiary treatment downstream conventional wastewater treatment plants that are not specifically designed to remove these types of contaminants, often present in wastewater treatment plant effluents (Castiglioni et al. 2006).

34

N. Morin-Crini et al.

The contaminant removal in constructed wetlands is due to the synergistic effect played by plant roots and their associated microbiome. The plant root system is an environmental niche that offer favorable conditions to microorganisms, which are specifically recruited through the release of root exudates and can help the plants to cope with adverse conditions (Vandenkoornhuyse et al. 2015), like the toxicity of emerging contaminants present in wastewaters (Riva et al. 2020). Moreover, microorganisms can play a direct role in contaminant degradation. Constructed wetlands can be designed using single or mixed plant species, selected among the most applied genera like Typha, Phragmites, Iris and Juncus. Interestingly, a constructed wetland co-cultivating the halophytes Tamarix parviflora, Juncus acutus, Sarcocornia perrenis, and Limoniastrum monopetalum, proved to efficiently decrease organic matter and pathogen concentration in the effluent (Fountoulakis et al. 2017). This result assumes a relevance for the bioremediation of wastewater produced by aquaculture, an anthropic activity of increasing environmental impact, also due to the intensive use of antibiotics. Constructed wetlands containing Iris pseudacorus and Phragmites australis, planted as single or mixed, were able to remove antibiotics and antibiotic resistance genes, with different removal efficacies according to the planting pattern and the target molecule. The tested plants showed the best removal performances in single species constructed wetlands. I. pseudacorus removed up to 77.64%, 68.70%, and 58.21% of enrofloxacin, sulfamethoxazole, and total antibiotic resistance genes, while P. australis showed removal efficiencies of 81.11%, 64.94%, and 56.26% for the same molecules and antibiotic resistance genes (Huang et al. 2019). Zhang et al. (2016b) setup a laboratory scale study under simplified conditions (i.e. hydroponic culture, artificially spiked wastewater) using different constructed wetland plant genera (i.e. Typha, Phragmites, Iris and Juncus) to investigate their remediation capacity towards two pharmaceuticals (ibuprofen and iohexol) characterized by different recalcitrance. Both ibuprofen and iohexol were uptaken by roots and partially translocated in the plant aerial portion, with diverse efficiencies depending upon the plant species and the considered molecule (Zhang et al. 2016b). Moreover, the experiment showed that ibuprofen was almost completely metabolized in the microcosms, and the degradation of iohexol could reach up to 80% of the initial amount using P. australis. The removal of ibuprofen and iohexol was due to both plant and microbe metabolisms, and the authors suggested a role of root exudates in the promotion of pharmaceuticals degradation, a well-known mechanism exploited for rhizoremediation of persistent contaminants in soil (Vergani et al. 2017). In a microcosm scale study using also P. australis in a vertical subsurface flow constructed wetland, high removal efficiencies were weekly shown for all micopollutants (except for 2-ethylhexyl-4-methoxycinnamate) in both spiked experiments with 36 multi-class pollutants and non-spiked freshwater aquaculture effluents containing atrazine, isoproturon, perfluorooctanesulfonic acid, clarithromycin, erythromycin, fluoxetine, norfluoxetine, and 2-ethylhexyl-4-­ methoxycinnamate (Gorito et al. 2018). The high removal capacity of P. australis was demonstrated also on the antiepileptic carbamazepine in a study that showed how several bacterial species isolated from P. australis endosphere (i.e. the internal root tissue) were able to remove and degrade the pharmaceuticals (Sauvêtre and

1  Remediation of Emerging Contaminants

35

Schröeder 2015). Likewise, bacterial endophytes were isolated from J. acutus plants grown in a pilot constructed wetland, treating water contaminated with bisphenol A, and many strains were able to tolerate and use bisphenol A, ciprofloxacin and sulfamethoxazole as sole carbon source (Syranidou et al. 2017). In the last decade, the role of bacteria dwelling the root system of constructed wetland plants has gained increasing attention, and the importance of microbial metabolism for the degradation of pharmaceuticals and other organic molecules suggested the possibility to exploit bacteria endowed with plant growth promoting capacities, tolerant and able to remove emerging contaminants in the so-called ‘microbial assisted phytodepuration’. For example, plant growth promoting bacteria improved the removal of the model azo-dye Reactive Black 5  in constructed wetland microcosms planted using J. acutus. Microcosms containing plants inoculated using two different single inocula and a mixed consortium of previously selected root-associated bacteria showed a higher decrease of Reactive Black 5 concentration in the effluent compared to the non-inoculated control ones, suggesting that this approach could be an interesting solution for the remediation of textile industry wastewater (Riva et al. 2019). On the whole, constructed wetlands represent a wastewater treatment technology of interest also for water reclamation, a current world-wide priority in the light of global warming and water crisis, although its benefits and risks should be carefully evaluated according to the final purposes of water reuse (Riva et al. 2020).

1.5.2  A  lgal-Based Removal Strategies for Emerging Contaminants One solution could be the use of biological materials such as algae and fungi (Parlade et al. 2018; Silva et al. 2019; Tolboom et al. 2019; Tomasini and León-­ Santiesteban 2019). Algae are known to be effective in the treatment of water contaminated with organic pollutants through biological processes such as bioremediation (phytoremediation) and biosorption. The use of algae for the removal of emerging contaminants has many advantages such as the use of low-cost materials, low capital investment, simple operation, reduced maintenance and the absence of formation of degradation by-products. Algae are also highly adaptive microorganisms and can grow autotrophically, heterotrophically or mixotrophically. They can grow in very harsh environmental conditions such as low nutrient levels, and extreme pH and temperature. Algae can acclimatize not only to change depending on the temperature and nutrient availability, but also salinity and light. In general, the characteristics of municipal wastewaters and industrial effluents (e.g. textile and pulp and paper industries) are suitable for algae cultivation, since these waters are a source of nutrients. Over the past two decades, extensive research has shown the potential application of advanced algal-based technologies for the removal of pollutants (Silva et al. 2019; Tolboom et al. 2019).

36

N. Morin-Crini et al.

Tolboom et al. (2019) recently reviewed algal-based removal strategies for the removal of emerging contaminants through efficient biological degradation. In laboratory-­scale studies, algae-based bioreactors such as open ponds and bubble column photobioreactors are capable of removing pharmaceuticals and endocrine disruptors. Open ponds are artificial ponds of limited depth (around 0.03–0.07 m) used for the cultivation of microalgae without agitation. Their advantages are low investment cost, ease of use and low operating cost. However, a lot of space is required for algae growth. Other problems include poor use of light by the cells, variations in pH and dissolved oxygen, water loss due to evaporation, diffusion of carbon dioxide into the environment, temperature fluctuations, and inefficient agitation. The operational factors that influence algae growth in these open systems are mixing, dilution rate and depth. Open ponds operate with a long hydraulic retention time to consume carbon dioxide during the day (photosynthesis) and provide oxygen for aerobic biodegradation. Closed systems have been designed to overcome most of the problems associated with open systems. Photobioreactors are closed tubular systems, mainly designed in vertical, horizontal and helical form (coil tube, flat plate, bubble column, air column or agitation tank). Each system has its own advantages and drawbacks (Silva et al. 2019). Closed bioreactors provide a tightly controlled environment for the isolation of the microalgal strain, ensuring increased productivity, biomass quality and the ability to explore a wider range of strains. The benefits are an easier pH and temperature control, higher volumetric efficiency, better use of the growing area, higher capture of radiant energy and less water loss. Closed systems also reduce the contamination risk and the loss of carbon dioxide to the atmosphere. This concept is promising, but the costs of such reactors are higher (expensive to install and to maintain) and are an obstacle to its development. Their design must be carefully optimized for each individual strain. A detailed discussion on these systems and their physicochemical characteristics can be found in the following references: de Godos et  al. (2012), Singh and Sharma (2012), Slade and Bauen (2013), Meneses-Jácome et  al. (2016), Matamoros et  al. (2015), Gouveia et al. (2016), Norvill et al. (2017), and Silva et al. (2019). Tolboom et al. (2019) reported high removal percentages (>90%) for metoprolol, triclosan, and salicylic acid, moderate (50–90%) for carbamazepine and tramadol and very low (90% was reported by Bai and Acharya (2017). Carbamazepine (Matamoros et al. 2015) and trimethoprim (de Wilt et al. 2016) showed less promising results with elimination rates of 62% and 60%, respectively. In fact, the removal efficiency depended mainly on the algae species used, such as Chlorella vulgaris, Chlorella sorokiniana and Nannochloris sp. For example, Escapa et al. (2017a, b), who evaluated the removal capacity of Chlorella vulgaris, Tetradesmus obliquus and Chlorella sorokiniana from wastewater containing

1  Remediation of Emerging Contaminants

37

paracetamol and salicylic acid using algae-based bioreactors, demonstrated that T. obliquus removed both pharmaceutical contaminants better than C. vulgaris in batch culture. The removal efficiency of salicylic acid by Tetradesmus obliquus was greater than 93% and that of Chlorella vulgaris was higher than 25%. Parameters that play an important role in pollutant degradation processes are not only the consortium of microorganisms present and its efficiency, but also the conditions used during degradation (temperature, seasons, retention time, environmental pH, dissolved oxygen, light periods). Performance also depends on chemical factors related to the pollutants, such as their structure and stereochemistry, concentration, toxicity and the presence of several pollutants. The mechanisms involved in algae photobioreactors are complex and are being elucidated (de Godos et al. 2012; Matamoros et al. 2015; Xiong et al. 2016; Silva et  al. 2019; Tolboom et  al. 2019). Nevertheless, the elimination processes often referred to are biodegradation, photodegradation and biosorption (cell sorption and bioaccumulation) to algae. Other mechanisms and interactions are also cited, such as volatilization (stripping), biotransformation, bioprecipitation (biomineralization), and oxidation/reduction reactions. Bioaccumulation of pollutants in algae cells can also induce the generation of reactive oxygen species, free radicals, and non-radical forms such as hydrogen peroxide and single oxygen. For example, de Godos et  al. (2012), studying the mechanism of elimination of tetracycline by Chlorella vulgaris, reported that photodegradation and biosorption were the main interactions to explain biodegradation. Matamoros et al. (2015) showed that biodegradation and photodegradation were the most relevant removal pathways for 26 contaminants in municipal wastewater in a study conducted in two pilot ponds of “high quality algae”. Volatilization was considered negligible for pharmaceuticals due to their low Henry’s constant values. However, for recalcitrant hydrophobic compounds, volatilization and sorption pathways were predominant. Xiong et al. (2016), studying the removal of carbamazepine by the microalgae Chlamydomonas mexicana and Scenedesmus obliquus, demonstrated that both species simultaneously promote the biodegradation, adsorption and bioaccumulation of carbamazepine. All the studies concluded that this technique can open up new opportunities for wastewater treatment and reduce environmental pollution that can have adverse effects on the ecosystem and human health. There are still areas to be explored or improved, such as the testing of other algal species (often the same species are used alone or in combination), the selection of different microorganisms with specific metabolism for different pollutants (currently there is a lot of research on the use of genetically modified microorganisms), the improvement of knowledge on the often slow kinetic and biodegradation mechanisms, and the optimization and better control of the operating parameters (temperature, pH, dissolved oxygen, etc.) as they are difficult to control on a large scale, which is a great challenge for this application. It is also necessary to acquire knowledge on other problems such as the incomplete transformation of certain recalcitrant pollutants.

38

N. Morin-Crini et al.

1.5.3  Use of Fungi for Removal of Emerging Contaminants Fungi have also been recognized for their ability to transform a wide range of recalcitrant compounds using non-specific intracellular and extracellular oxidizing enzymes (Spina et al. 2012). The use of fungi for the removal of emerging contaminants through biological processes such as mycoremediation and biosorption has the same advantages as those cited for the use of algae (Spina et al. 2012; Zhang et  al. 2013; Badia-Fabregat et  al. 2015; Asif et  al. 2017; Tomasini and León-­ Santiesteban 2019). Fungi are indeed microorganisms known for their effectiveness in treating water contaminated by pollutants such as pharmaceuticals. Fungal reactors (mycoreactors) can be suspended or immobilized growth systems. They can operate under aerobic or anaerobic conditions. This technology requires specific and controlled conditions in order to maintain a sustainable and efficient process. For example, the oxidative metabolism of fungi can be strongly affected by the presence of nutrients, pH, immobilization on different supports, and agitation or static growth conditions. Recent comprehensive reviews on fungal technologies for the degradation of pharmaceuticals and personal care products have been published (Asif et  al. 2017; Rodríguez-Rodríguez et  al. 2019; Silva et  al. 2019; Tomasini and León-­ Santiesteban 2019). Cruz-Morató et  al. (2013) studied the degradation of pharmaceuticals in non-­ sterile urban wastewater at the Universitat Autónoma de Barcelona (Spain) by Trametes versicolor in a fluidized bed reactor. Of the 80 pharmaceuticals analyzed, 13 were detected in the effluent of sterile urban wastewater, the most abundant belonging to the group of analgesic/anti-inflammatory compounds, especially naproxen (35.58 ± 4.8 μg/L) and ibuprofen (12.61 ± 1.79 μg/L). Complete elimination of both analgesics, ibuprofen and naproxen, occurred within 24 h after fungal treatment. A similar conclusion was reported by Marco-Urrea et al. (2009). Other analgesics as acetaminophen and codeine were initially detected at concentrations of 3.87 ± 0.41 μg/L and 0.02 ± 0.001 μg/L and were completely removed after 8 h and 2 days, respectively (Cruz-Morató et al. 2013). The analgesics ketoprofen and salicylic acid were also initially detected in wastewater at concentrations of 0.48 ± 0.07 μg/L and 0.85 ± 0.11 μg/L, respectively. Changes in concentrations during treatment showed unexpected behavior with increases and decreases, but after 8 days the concentration were 0.31 ± 0.04 μg/L (ketoprofen) and 1.24 ± 0.07 μg/L (salicylic acid), corresponding to a 35% removal of ketoprofen and a 46% increase in salicylic acid concentration. The possible release of these compounds may be explained by the deconjugation of glucuronides during biological treatment. Complete removal of 7 of the 10 initially detected pharmaceuticals was achieved in non-sterile conditions after 24  h of fungal treatment, while only 2 were partially removed and 1 of the pharmaceuticals tested increased its concentration. Antibiotics such as erythromycin (detected at 0.3 μg/L) and metronidazole (detected at 0.05  μg/L) were successfully removed: erythromycin was completely eliminated within 15 min while the elimination of metronidazole was obtained after 2 days.

1  Remediation of Emerging Contaminants

39

Carbamazepine, a well-known recalcitrant psychiatric drug in activated sludge treatment was also detected in the wastewater at 0.7μg/L. Cruz-Morató et al. (2013) showed that carbamazepine and its metabolites were also strongly eliminated by T. versicolor during batch treatment in a fluidized bed bioreactor under sterile conditions, in agreement with the results published by Zhang and Geiβen (2012). In addition, the Vibrio fischeri luminescence test (Microtox® test) showed a significant reduction in wastewater toxicity after treatment. Cruz-Morató et  al. (2013) concluded that it was possible to use a fluidized bed bioreactor to remove pharmaceuticals at environmentally relevant concentrations under non-sterile conditions by T. versicolor. In another work, the same authors treated hospital wastewaters (collected in the main sewer of the University Hospital of Girona, Dr. Josep Trueta, Girona, Spain) in a fungal bioreactor with Trametes versicolor, under sterile and non-sterile conditions (Cruz-Morató et  al. 2014). Preliminary analytical monitoring of hospital wastewater showed that the most frequently detected families of substances were analgesics, antibiotics, psychiatric drugs, endocrine disruptors and X-ray contrast media, at concentrations ranging from ng/L to mg/L. Results showed that 46 of the 51 detected pharmaceuticals were partially degraded and completely eliminated in non-sterile experiments after 8 days. The initial total amount of pharmaceuticals in the bioreactor was 8185μg in sterile treatment and 8426 μg in non-sterile treatment, and the overall load removal was 83.2% and 53.3% in their respective treatments. In particular, diclofenac (a recalcitrant compound in municipal wastewater treatment plant) and human metabolites of carbamazepine were efficiently removed. The lower removal detected in the non-sterile treatment is explained by the higher concentrations of caffeine (149 μg/L) and iopromide (419.7 μg/L), which are some of the most difficult compounds to be  degraded by the fungus (80–90%) for acetaminophen, ibuprofen, bezafibrate, estrogens (estrone, estradiol, estriol). Phan et al. (2015) reported that membrane bioreactor technology can generally achieve high pollutant removal efficiency with the effluent quality largely complying with the Australian guidelines for water recycling (except for caffeine, estrone and triclosan). Performance is a function of temperature (Suárez et  al. 2012; Gurung et  al. 2017) and pH-dependent (Sanguanpak et al. 2015). Suárez et al. (2012) evaluated the removal efficiency of sulfamethoxazole and erythromycin as a function of temperature and the results indicated that an increase in temperature resulted in 30% higher removal than in cold weathers. However, an increase in temperature to a high level, e.g. 45 °C, may inhibit metabolic activity (Besha et al. 2017). Similarly, low temperature impairs treatment efficiency. Gurung et al. (2017) also indicated that the removal efficiency of bisoprolol, diclofenac and bisphenol A was highly dependent on temperature,

1  Remediation of Emerging Contaminants

43

with maximum removal of 65%, 38% and >97%, respectively. Sanguanpak et al. (2015) reported that membrane bioreactors performed better for analgesics and anti-­ inflammatory drugs (e.g. ibuprofen, diclofenac and ketoprofen) under low pH conditions (the optimal pH was about 6), mainly due to the increased lipophilicity and the ionisability of pH-dependent molecules. Racar et  al. (2020) investigated the removal of several emerging substances, mainly pharmaceuticals, from wastewater from a treatment plant located in Čakovec (Croatia) by a membrane bioreactor. A 6-month analytical monitoring of wastewater showed 12 substances were systematically detected with a high variation in concentration. The highest values (up to 500 μg/L) were found in the winter period. Azithromycin (92.54  ±  113.90 μg/L), clarithromycin (50.49  ±  80.95 μg/L), and diclofenac (71.57 ± 57.41 μg/L) were the most prevalent, while the other measured pharmaceuticals showed wide variations in concentration, in particular acetamiprid, clothianidin, imidacloprid and thiamethoxam, a high concentration in November, while the other months had significantly lower concentrations. The authors demonstrated that membrane bioreactor technology achieves high removal rates (to levels below the limit of quantification) for all substances, regardless of the season. Vieira et al. (2020) recently compared the efficiency of adsorption, membrane separation and biodegradation to remove endocrine disruptors (parabens, bisphenols, phthalates, estrogens and nonylphenols) and pesticides and discussed the advantages and disadvantages of each process. Literature data shows that the three main removal mechanisms for substances that occur in membrane bioreactors are adsorption/sorption/biosorption (onto sludge flocs and bound microbial products), biological degradation (aerobic degradation, anaerobic degradation, metabolism and co-metabolism, ion trapping mechanisms) and membrane separation. For adsorption process, the efficiency largely depends on the physicochemical characteristics of emerging substances, e.g. hydrophobicity, hydrogen bond, electrostatic interactions, etc. For biological degradation, performance depends also on the biodegradability and bioavailability of the substances, and the condition used (oxidation-reduction potential plays an important role in the microbial diversity, enzymatic functions and activities). Suárez et  al. (2012) reported that musks (galaxolide, tonalide and celestolide) and estrogens (estrone and estradiol) can be well degraded under aerobic and anoxic conditions, whereas the transformation of ibuprofen, roxithromycin, erythromycin, citalopram and naproxen occurs only in the aerobic process. In contrast, the degradation of diclofenac, sulfamethoxazole, diazepam, trimethoprim and carbamazepine is much less effective in the presence of oxygen species. Membrane separation is another mechanism (size exclusion, charge repulsion) contributing to pollutant removal and it depends on the type of membrane used. Alvarino et al. (2017) demonstrated that the removal efficiency of diclofenac and roxithromycin in an ultrafiltration membrane bioreactor was higher than that in a microfiltration membrane bioreactor due to the retention by the cake layer. However, ultrafiltration does not eliminate all pollutants. The solution would be to generalize the use of nanofiltration and osmosis membranes, but then there is the problem of energy consumption and maintenance costs of the membranes, which can clog up more quickly.

44

N. Morin-Crini et al.

Most of the studies published in the literature on the behavior of membrane bioreactors in relation to emerging substances (mainly pharmaceuticals) have been conducted on a  laboratory scale. Pilot projects at industrial scale have yet to be conducted on a much wider range of substances including different families of pharmaceutical compounds, personal products and cosmetics, pesticides and industrial substances. Another important challenge is the presence of nanoparticles and nanoproducts in wastewater that can create clogging problems. The sustainable application of membrane bioreactor technology requires a better understanding of the fate of pollutants, research on biotransformation mechanisms and the combination of bioreactors with emerging technologies such as advanced oxidation processes (Borea et al. 2019; Monteoliva-García et al. 2019). The idea developed in this recent research theme is to use the redox reactions that occur on the surface of a conductive membrane (under anodic or cathodic polarization). These reactions can facilitate the transformation of refractory pollutants, which is a clear advantage over conventional processes, and reduce membrane fouling, while having negligible effects on microbial activities in the effluent. There are two other aspects that are becoming increasingly important in the field of water treatment. The first is the recycling of water, especially for irrigation of agricultural soils or golf courses. Recycling or reuse of wastewater could also be a way of supplementing available water supplies. Effluents from membrane bioreactor technology are capable of meeting or exceeding current drinking water regulations. However, there are several barriers to water reuse. For example, the public perception of recycled water for drinking water production is less than favorable, while the reuse of water for irrigation is generally accepted. The second is to consider wastewater as a resource and not as a waste. Conventional municipal wastewater treatment plants such as those applying activated sludge are energy-intensive, produce large quantities of residues (sludge) and fail to recover the potential resources available in wastewater. Furthermore, in this respect, wastewater is often considered as a waste. However, wastewater should be considered a resource because it contains organic matter, phosphorus, nitrogen, metals and energy. Membrane techniques could make it possible to recover compounds with high added value. Of course, there is the question of the presence of emerging substances. In this context, biological membrane technologies coupled with other methods could be useful.

1.6  R  emoval of Emerging Contaminants by Membrane Filtration Advanced treatment methods such as membrane filtration are potential technologies capable of removing a wide range of pollutants detected in water, even at trace levels, and are already used at industrial scale as tertiary treatment (Kunst and Košutić 2008; Ojajuni et al. 2015; Hu et al. 2017; Warsinger et al. 2018). In simple terms,

1  Remediation of Emerging Contaminants

45

the membrane filtration process is a physical separation method characterized by the ability to separate molecules of different sizes and characteristics. Its driving force is the pressure difference between the two sides of a special membrane. The technology involves passing a single feed stream through a membrane system that separates it into two individual streams, called permeate and retentate. The membrane that separates them is a physical barrier with highly specialized characteristics – a barrier that only selected components of the feed stream can pass through. There are four main types of membranes, defined according the size of the substance they must separate from the feed effluent: reverse osmosis, nanofiltration, ultrafiltration and microfiltration, in order of pore size. For wastewater treatment, the two preferred methods are reverse osmosis and nanofiltration, which are pressure-­driven membrane technologies that can remove contaminants to 0.1  nm and 1 nm, respectively. Both methods are diffusion and size exclusion controlled processes. They have the largest treatment capacity but require the greatest degree of pre-treatment. These technologies are known for their capacity to remove pharmaceuticals (ibuprofen, naproxen, trimethoprim, diazepam 17-β-estradiol, testosterone) and personal care products (triclosan, alkylphenols). Mehran Abtahi et  al. (2018) showed that the removal rates of diclofenac, naproxen, ibuprofen and 4-n-nonylphenol by nanofiltration can reach 77%, 56%, 44%, and 70%, respectively. Alonso et al. (2018), investigating the removal of ciprofloxacin in synthetic seawater by a commercial spiral-wound reverse osmosis membrane, showed that more than 90% of ciprofloxacin could be successfully removed by reverse osmosis membrane under constant pressure, and the maximum removal rate was 99.96%. In general, to treat pharmaceuticals and personal care products, nanofiltration has shown to be superior to other conventional filtration methods, in terms of effluent quality, easy operation and maintenance procedures, low cost, and small required operational space (Hu et al. 2017; Mehran Abtahi et al. 2018). However, reverse osmosis is the most effective in removing pesticides. In reality, there are no general rules because the performance depends not only on the membranes used and the pre-treatments undergone by the effluents, but also on several parameters such as the complexity of the effluents, the type(s) of substance(s), or their concentration. It is therefore necessary to carry out numerous tests at pilot scale before defining the choice of the membrane technique to be used. A recent review by Kim et al. (2018) provides a comprehensive overview of the most relevant works available in literature reporting the use of membrane filtration for the removal of emerging contaminants. The fundamental knowledge of forward osmosis (the force that drives the separation is concentration gradient), reverse osmosis (pressure-driven process), nanofiltration (pressure-driven process) and ultrafiltration (pressure-driven process) technologies has been described. This review also aimed to address several key parameters, including the physicochemical properties of emerging contaminants (solute molecular weight/size/geometry, charge, and hydrophobicity), water quality conditions (pH, solute concentration, temperature, background inorganics, and natural organic matter), and membrane properties and operating conditions (membrane fouling, membrane pore size, porosity, charge, and pressure) that influence the removal of emerging contaminants

46

N. Morin-Crini et al.

during membrane filtration, and separation mechanisms. Overall, the general emerging contaminant removal trend was as follows: (i) the removal efficiency for the membranes follows the declining order: reverse osmosis ≥ forward osmosis > nanofiltration > ultrafiltration; (ii) the retention of emerging contaminants by reverse osmosis and forward osmosis membranes is mainly governed by size/steric exclusion, while high retention can still be achieved due to hydrophobic (adsorption) and electrostatic (attraction) interactions for nanofiltration and ultrafiltration membranes; (iii) more polar, less volatile, and less hydrophobic organic emerging contaminants have less retention than less polar, more volatile, and more hydrophobic substances; (iv) while, in general, forward osmosis and reverse osmosis membranes show significant metal/toxic anion retention (>95%) regardless of water quality and operating conditions, metal/toxic anion retention by nanofiltration and ultrafiltration membranes is more efficient at neutral and alkaline conditions than at acidic values; and (v) while ultrafiltration alone may not effectively remove emerging contaminants, it can be employed as a pretreatment step prior to forward osmosis and reverse osmosis. The advantages of membrane filtration often cited are: a wide range of commercial membranes available from several manufacturers, modular design, small footprint, simple and efficient technologies, production of high quality treated wastewater, no chemicals required, low energy consumption, and well-known separation mechanisms. The main disadvantages are: high investment costs, high energy requirements, high maintenance and operating costs, rapid clogging of the membranes (pre-treatment is essential), formation of separate contaminant concentrates (retentate), which must be removed. A major disadvantage of membrane filtration compared to advanced oxidation processes is indeed that the substances are only transferred into concentrate streams and not destroyed, as in the case of advanced oxidation, and therefore the concentrate requires additional treatment and disposal. Despite this, in Europe, advanced tertiary adsorption processes on activated carbon, combined with a chemical oxidation step, are in competition with membrane technologies because the latter are particularly effective in removing complex mixtures of substances, even in trace amounts and in the presence of many interferents.

1.7  A  dvanced Oxidation Processes to Degrade Emerging Contaminants 1.7.1  R  emoval of Emerging Contaminants Through Wastewater Disinfection The main objective of disinfection is to either eliminate pathogens from water for the production of potable water or to reduce the pathogen content of treated wastewater in wastewater treatment plants. Disinfection is indeed the final treatment step for the production of potable water. It is an important step because the use of water disinfection as a public health measure reduces the spread of diseases.

1  Remediation of Emerging Contaminants

47

Pathogenic microorganisms are destroyed or inactivated using chemical or physical disinfectants such as chlorine, chlorine dioxide, hypochlorite, ozone, peracetic acid, bromide, iodine, non-ionizing radiation (ultraviolet radiation, ultrasonic radiation), and ionizing radiation (gamma ray). Among chemicals, chlorine and its compounds, chlorine dioxide and ozone, are the most common disinfectants used in the water industry. The main mechanisms of germicidal action of disinfectants are related to the direct oxidation of the cell of microorganisms by the disinfectant or to the alteration of the permeability of the cell wall, or even to the photochemical deterioration of their DNA or RNA (ultraviolet radiation) (Asano et  al. 2007). Disinfectants are also known to remove organic contaminants from water, which act as nutrients or shelter for microorganisms. They must also have a residual effect to prevent microorganisms from growing in the pipe after treatment, which would result in recontamination of the water (Collivignarelli et al. 2018). The main problem with disinfection, with the exception of ultraviolet radiation, is that the processes can lead to the formation of organic and inorganic disinfection by-products such as trihalomethanes, chlorite and chlorate, aldehydes, etc. (von Sonntag and von Gunten 2012). Advanced technologies include the combination of ozone and hydrogen peroxide, ozone and ultraviolet radiation, hydrogen peroxide and ultraviolet radiation, ultraviolet radiation with titanium dioxide, alone or combined with other processes, such as land filtration, membrane technologies, nanotechnology, photovoltaic method, solar photocatalytic, sonodisinfection etc. Disinfection processes using disinfectants, alone or in combination with additional physical/chemical agents, have also been proposed to eliminate emerging substances (Ikehata et al. 2006, 2008; Snyder et al. 2006; Gagnon et al. 2008; Kim and Tanaka 2009; Hey et al. 2012a, b; Noutsopoulos et al. 2013a; Yang et al. 2013). In municipal wastewater treatment, disinfection is usually the last step (tertiary or final treatment) before the treated wastewater is released to the aquatic environment. However, the treatment can also take place after primary and secondary wastewater treatment. Certain types of pollutants such as diclofenac (anti-inflammatory), boldenone (anabolic steroid), sulfamethoxazole (bacteriostatic antibiotic), and clofibric acid (lipid-regulating agent), can be effectively degraded at reasonably low doses for drinking water and wastewater treatment. Kim and Tanaka (2009) reported that photodegradation rates of up to 100% were obtained when diclofenac was treated using ultraviolet disinfection, but with ultraviolet dose which is 6 times higher than what typically used to disinfect drinking water. However, in general, direct ultraviolet irradiation at disinfection doses was no effective in removing trace pollutants such as hormones (Rosenfeldt and Linden 2004), bisphenol A (Chen et al. 2006), or ketoprofen (Real et al. 2009). A review of the abundant literature published over the past 20 years shows that the three main technologies studied for both disinfecting water from pathogens and removing substances are chlorination, ozonation and ultraviolet irradiation. These are the technologies used at industrial scale, either alone or in combination with other chemicals. For example, oxidative removal of trace organic contaminants in water (surface water, wastewater) can be achieved using ozone combined with

48

N. Morin-Crini et al.

hydrogen peroxide or ultraviolet combined with hydrogen peroxide (Kruithof et al. 2007; Gagnon et al. 2008; von Sonntag and von Gunten 2012; Yang et al. 2013). Chlorination  The main chlorination compounds used in the wastewater industry are chlorine (Cl2), sodium hypochlorite (NaOCl) and calcium hypochlorite (Ca(OCl)2). Of these, chlorine gas is generally used in large wastewater treatment plants, although a switch to sodium hypochlorite has been noted in many cases over the last decade due to safety concerns, while calcium hypochlorite is allowed in smaller plants. When chlorine gas is dissolved in water, it hydrolyzes according to Eq. (1.1). The hypochlorous acid produced during this reaction is a weak acid which dissociates easily in the aqueous solution according to Eq. (1.2). The sum of hypochlorous acid and the hypochlorite ion forms free chlorine. The relative distribution between the two chlorine species depend on both the temperature and pH of the wastewater (Fig. 1.8) and is very important because the acid form is a much more effective oxidant that the ionic form.

Cl 2 + H 2 O  HOCl + H + + Cl −

(1.1)



HOCl  OCl − + H +

(1.2)

The effect of chlorination on the removal of emerging contaminants has been widely documented (Hu et al. 2002; Petrovic et al. 2003; Westerhoff et al. 2005; Zhang and Grimm 2005; Greyshock and Vikesland 2006; Thurman 2006; Korshin 2006; Stackelberg et al. 2007; Simazaki et al. 2008; Benotti et al. 2009a, b; Acero et al. 2010; Quintana et al. 2010; Chen et al. 2013; Ga et al. 2014; de Jesus Gaffney et al. 2016). The degree of removal varies for different types of emerging contaminants. Nevertheless, all these studies on chlorination experiments are focused on pure water or drinking water rather than in the wastewater matrix. On the other

Fig. 1.8  Relative distribution among the dominant chlorine species versus pH at 25 °C for a chlorine concentration of 10 mg/L. (Source: Constantinos Noutsopoulos, Athens, Greece)

1  Remediation of Emerging Contaminants

49

hand, studies on the ability of chlorination to remove emerging contaminants from wastewater are rather rare (Renew and Huang 2004; Belgiorno et al. 2007; Nakamura et al. 2007; Ying et al. 2009; Li and Zhang 2011; Noutsopoulos et al. 2013a, b, c, 2015; Nika et al. 2016). Based on the extensive data in the literature, chlorination appears to be effective in the removal of several non-steroidal anti-inflammatory drugs, antibiotics, estrogens and antidepressants (diclofenac, naproxen, sulfamethoxazole, amitriptyline hydrochloride, methyl salicylate), endocrine disrupters (nonylphenol, bisphenol, triclosan and 17-β-estradiol), benzotriazoles and benzothiazoles, while other chemicals in these categories have significantly lower degradability, e.g. ibuprofen, ketoprofen, 17-β-estradiol and tolytriazole. The performance of the chlorination process in removing emerging contaminants is related both to the characteristics of the matrix (organic matter content of the wastewater, presence of total suspended solids, pH, etc.) and to the physicochemical characteristics of the chemicals. As suggested by Noutsopoulos et al. (2015), the performance of the chlorination process for the removal of targeted endocrine disruptors and non-steroidal anti-­inflammatory drugs was not affected by the pH for typical wastewater pH values around 7–8. On the other hand, some studies reported a significant effect of pH on the removal of new contaminants (Acero et  al. 2010; Gallard et  al. 2004; Pinkston and Sedlak 2004; Deborde and von Gunten 2008) for pH values significantly different from those prevailing in wastewater treated by secondary and tertiary treatment. In general, pH can effectively affect process performance at low values (lower than those prevailing in treated wastewater) that favor the prevalence of the strongest oxidizing species (i.e. hypochlorous acid). In addition, it is expected that not only the available free chlorine species (HOCl and OCl−), but also the chemical characteristics (pKa, chemical structure) of the compounds under different pH conditions may also affect chlorination performance. With respect to other wastewater characteristics, Noutsopoulos et al. (2015) reported that the effect of total suspended solids content of wastewater (and thus organic matter content) on the degradation of emerging contaminants during chlorination is more profound for chemicals with high Kow values (e.g. nonylphenol and its ethoxylates, triclosan), and thus a high affinity to be distributed to the particulate phase. In the same study, the effect of humic acids on the removal of emerging contaminants and non-steroidal anti-­inflammatory drugs by wastewater chlorination was rather minimal. In most studies, degradation of emerging contaminants appears to follow pseudo first-order kinetics and values of rate constant (k) and half-lives (t1/2) have been determined for many chemicals (Deborde and von Gunten 2008; Quintana et  al. 2012). According to Deborde and von Gunten (2008), the reactivity of hypochlorous acid with organic micropollutants refers to (i) oxidation reactions, (ii) addition reactions to unsaturated bonds, and (iii) electrophilic substitution reactions at nucleophilic sites. Thus, as explained by the authors, the effect of chlorine on aliphatic organic compounds (with the exception of compounds with sulfur or nitrogen-­containing fractions) is generally small, whereas for the most popular emerging contaminants (nonylphenol, bisphenol A and triclosan) and the steroid hormones estrogen, the reactivity of chlorine occurs mainly on the phenolic ring. A

50

N. Morin-Crini et al.

significant number of studies on endocrine disruptors and chlorination by-products of pharmaceuticals provide an analytical discussion of the relevant mechanisms (Hu et al. 2002; Hu et al. 2003; Petrovic et al. 2003; Gallard et al. 2004; Bedner and MacCrehan 2006; Korshin et al. 2006; Thurman 2006; Lei and Snyder 2007; Sharma 2008; Quintana et al. 2012; Bulloch et al. 2012; Soufan et al. 2012; Noutsopoulos et  al. 2015). In many cases, toxicity measurements show that some of the by-­ products of chlorination are more toxic than the original compounds. Chlorine Dioxide  This chemical agent has a stronger disinfectant activity than chlorine. Considering its higher cost compared to conventional chlorination, disinfection with chlorine dioxide is generally adopted in cases where minimizing the production of chlorine-based disinfection by-products is desirable. Because of its instability, chlorine dioxide (ClO2) is produced on-site by mixing a chlorine solution with a sodium chlorite solution. The literature on the effectiveness of chlorine dioxide in removing emerging contaminants during wastewater disinfection is rather limited. In their comprehensive review, Hey et  al. (2012a) investigate the effect of ClO2 on the removal of a wide range of 56 pharmaceuticals. According to this study, it was shown that, in addition to the effectiveness of chlorine dioxide in removing many emerging contaminants, approximately one-third of the target compounds studied were virtually unaffected by the oxidant, even at doses as high as 20 mg/L. In their follow-up study, Hey et al. (2012b) reported the effective elimination of three non-steroidal anti-inflammatory drugs (naproxen, diclofenac, mefenamic acid) and a lipid-regulating agent (gemfibrozil), while no elimination was reported for ibuprofen and clofibric acid. The inability of ClO2 to remove ibuprofen and carbamazepine has also been reported previously by Lee and van Gunten (2010), while satisfactory removal has been reported for sulfamethoxazole and 17α-ethinyl estradiol. In addition, Huber et al. (2005a, b) showed an effective elimination of estrogenic hormones at very low doses of ClO2 and a short contact time (5 min). The ability of chlorine dioxide to remove emerging contaminants has also been demonstrated for a range of antibiotics (Navalon et al. 2008; Wang et al. 2011). The variable efficacy of chlorine dioxide on the removal of pharmaceuticals was also recorded in the study by Sharma (2008). According to his conclusions, chlorine dioxide is an effective oxidant. UV Radiation  This process is a well-known disinfection step used to effectively remove bacteria, viruses and protozoa without producing toxic by-products. The main types of lamps are low-intensity low-pressure lamps, high-intensity low-­ pressure lamps and high-intensity medium-pressure lamps, the former being the most commonly used for disinfection purposes (Asano et al. 2007). The main germicidal mechanism of ultraviolet irradiation is associated with direct DNA damage, while the removal of organic pollutants is based on their direct photolysis during absorption of UV-C protons (at the wavelength of 254  nm). The effectiveness of ultraviolet irradiation on the removal of pathogens and organic micropollutants is highly dependent on the applied dose (in mJ/cm2 or mWs/cm2), calculated as the product of average ultraviolet intensity and contact time. The effectiveness of ultra-

1  Remediation of Emerging Contaminants

51

violet irradiation for the removal of emerging contaminants has been well documented, primarily for pharmaceuticals and endocrine disruptors (Andreozzi et al. 2003; Doll and Frimmel 2003; Lopez et  al. 2003; Rosenfeldt and Linden 2004; Vogna et al. 2004; Chen et al. 2006; Neamtu and Frimmel 2006; Pereira et al. 2007a, b; Canonica et al. 2008; Benotti et al. 2009a, b; Kim et al. 2009a; Kim and Tanaka 2009; Yuan et  al. 2009; Rosario-Ortiz et  al. 2010; Zhang et  al. 2010; Baeza and Knappe 2011; Salgado et al. 2011, 2012, 2013; Hansen and Andersen 2012; Pablos et al. 2013; Bennett et al. 2018; Mole et al. 2019). Kim et al. (2009a, b) reported that of the 42 pharmaceuticals studied under real wastewater conditions, only a few showed high removal (ketoprofen, diclofenac and antipyrine), while several others (particularly antibiotics, e.g. clarithromycin, erythromycin and azithromycin) showed very low degradation due to ultraviolet irradiation, even at doses above 2700 mJ/cm2. Consequently, Noutsopoulos et al. (2013c) concluded that with the application of the low-pressure ultraviolet doses generally adopted for pathogen removal (10–80  mJ/cm2), no significant removal should be expected for many emerging contaminants. Based on this study, the endocrine disruptors bisphenol A and nonylphenol and the non-steroidal anti-inflammatory drugs ibuprofen and naproxen showed a low degradability even at doses as high as 1000 mJ/cm2, confirming the results of previous studies (Rosenfeldt and Linden 2004; Vogna et al. 2004; Chen et al. 2006, Pereira et al. 2007a, b; Canonica et al. 2008; Yuan et al. 2009; Rosario-Ortiz et  al. 2010; Baeza and Knappe 2011; Salgado et  al. 2011; Pablos et al. 2012). The moderate effect of ultraviolet has also been confirmed by Bennett et al. (2018) for estrogens. The authors suggested that the complete degradation of estrogens could only be achieved at ultraviolet doses (500–100 mJ/cm2) much higher than those used for pathogen removal. As concluded by many researchers (Kim et  al. 2009a; Yuan et  al. 2009; Pablos et  al. 2013; Noutsopoulos et  al. 2013a), the structure of each chemical as well as its physical characteristics (i.e. decadal molar absorption coefficient at 254 nm wavelength) largely govern its sensitivity to degradation. Studies of ultraviolet radiation transformation by-products on emerging contaminants are rather limited (Salgado et  al. 2013; Bennett et  al. 2018) and the hypothesis that photodegradation intermediates may be more recalcitrant or toxic than parent compounds has therefore yet to be fully demonstrated. Several modifications have been suggested in order to improve the efficiency of ultraviolet irradiation such as the use of medium-pressure lamps (Kim et al. 2009a; Pereira et al. 2017) or the ultraviolet based advanced oxidation processes such as UV/H2O2, UV/Cl2 and VUV/O3 (Kruithof et al. 2007; Yuan et al. 2009; Xiang et al. 2016; Lian et al. 2017). For example, full-scale application of ultraviolet combined with hydrogen peroxide in an existing water treatment plant in North Holland for a running time of more than 2 years has proven that this practice was effective and reliable to control organic micropollutants (Kruithof et al. 2007). The UV/H2O2 process was installed between the sand filtration and granular activated carbon filtration processes. Substances such as mecoprop and diclofenac were removed by 98%, while the removal of the other compounds (pesticides) varied from 60% to 91%. Carbon filters could effectively remove residual hydrogen peroxide and at least conceptually, also any by-products formed in the oxidation process, as well as assimi-

52

N. Morin-Crini et al.

lable organic carbon that can feed microbes in biofilm formed within the water distribution lines. Ozonation  Ozone is a very active oxidant and therefore a very effective germicide. It has been widely demonstrated that ozone has a remarkable ability to eliminate not only bacteria and viruses, but also pathogenic protozoa, compared to chlorine and chlorine dioxide. Ozone is produced on site and an ozonation unit consists of the air compressor, including cooling, drying and filtration accessories, the ozone generator, the contact tank for ozonation and the waste gas destruction device. Due to the high cost of the method and the high doses required in the wastewater, compared to natural water treatment, ozonation was not considered before an attractive method for wastewater disinfection and its use was limited to water disinfection. However, after having corroborated its ability to remove emerging contaminants as well as its well-known disinfection performance, ozonation has received much attention over the last decade, especially in cases where specific provisions for the removal of emerging contaminants through wastewater treatment have been regulated (e.g. the Swiss Water Protection Act). Ozone reacts either through its molecular form (O3) or through the activity of hydroxyl radicals that are formed during its decomposition reactions in water. The ozone molecule reacts effectively with many compounds, especially those containing aromatic rings, unsaturations or heteroatoms, while hydroxyl radicals are powerful non-selective oxidants. Ozone tends to react preferentially with the hydrophobic fractions of organic compounds such as hydrophobic acid and neutral species. The effectiveness of ozonation is affected by several parameters such as temperature, pH, presence of organic compounds (e.g. chemical oxygen demand), natural organic matter, total suspended solids, nitrates, nitrites, etc. For example, it has been proposed that nitrites act as a radical scavenger, inhibiting the effectiveness of ozonation (Lee and von Gunten 2010). Several studies have been reported on the effectiveness of ozonation in removing emerging contaminants, although the majority of these studies use pure water rather than a wastewater matrix (Gehr et al. 2003; Xu et al. 2002; Huber et al. 2003, 2004, 2005a, b; Ternes et al. 2003; Kim et al. 2004; Larsen et al. 2004; Deborde et al. 2005; Irmak et al. 2005; Westerhoff et al. 2005; Buffle et al. 2006; Ikehata et al. 2006; Snyder et al. 2006; Wert et al. 2007; Benner et al. 2008; Dantas et al. 2008; Gagnon et al. 2008; Maniero et al. 2008; Coelho et al. 2009; Dodd et al. 2009; Lin et al. 2009; Leitner and Roshani 2010; Rosal et  al. 2010a, b; Schaar et  al. 2010; Stalter et  al. 2011; Tay et  al. 2010; Thompson et al. 2011; Zimmermann et al. 2011; Altmann et al. 2012, 2014; Lee and von Gunten 2012, 2016; Mawhinney et al. 2012; Reungoat et al. 2012; Kovalova et al. 2013; Lee et al. 2013, 2014; Snyder et al. 2006; Margot et al. 2013; Ahmed et al. 2017; Sun et al. 2017; Bourgin et al. 2018; Lacson et al. 2018; Paucar et al. 2018; Thanekar et al. 2018; Wang et al. 2018a). Gagnon et  al. (2008) studied the degradation of anti-inflammatory and anti-­ convulsant drugs in treated wastewater from a primary wastewater treatment plant in Montreal using ozonation. The analytical monitoring of treated wastewater before

1  Remediation of Emerging Contaminants

53

disinfection revealed concentrations ranging from 42 to 2556 ng/L for several pharmaceutical residues such as salicylic acid, clofibric acid, ibuprofen, naproxen, triclosan, carbamazepine, diclofenac, and 2-hydroxy-ibuprofen. After treatment, most of these substances were eliminated at a rate greater than 50% at an ozone dose of 10 mg/L. Higher elimination rates (up to 70%) were observed when 20 mg/L of ozone was used. The pilot projects carried out by the authors demonstrated that disinfection processes were a potential complementary method to degrade pharmaceutical residues still present after biological treatment. Snyder et al. (2006) also reported that in drinking water and wastewater experiments, the majority of estrogenic and androgenic steroids, pharmaceuticals, pesticides and industrial chemicals were eliminated by more than 90% in exposures to ozone commonly used for disinfection. Ozone has proven to be very effective in removing the majority of emerging trace substances from water. Pilot-scale experiments also demonstrated that for compounds that are highly resistant to ozone oxidation, such as atrazine, iopromide, meprobamate and tris-chloroethylphosphate, the removal performance was less than 50%. The addition of hydrogen peroxide for advanced oxidation was not very beneficial for contaminant removal compared to ozone alone. In addition, ozone and ozone combined with hydrogen peroxide have been shown to eliminate estrogenicity in vitro. According to extensive data in the literature, ozone has a high reactivity with pesticides, estrogens, endocrine disruptors, beta-blockers and many non-steroidal anti-inflammatory drugs and parabens, while a lower elimination capacity has been recorded for some antidepressants, antiepileptics, benzotriazoles, perfluorooctanesulfonic acids and perfluorooctanonic acids. Several researchers have postulated that decomposition of emerging contaminants is lower when using real water matrices and especially wastewater (Ternes et al. 2003; Lee and von Gunten 2010), rather than pure water, due to the presence of several parameters that could interfere with process performance. Recently, Kim et al. (2020) proposed new empirical models to predict ozonation kinetics and micropollutant reduction during ozonation. A disadvantage of the method is associated with the low mineralization obtained in some cases, which can lead to the production of reaction by-products that may exhibit stronger refractory behavior and possibly toxicity. Stalter et al. (2010a, b) provided some ecotoxicological evidence of increased toxicity following ozonation. Conversely, Nasuhoglu et  al. (2018) showed a decrease in estrogenic and androgenic activity of 98% and 68%, respectively, while anti-estrogenic activity remained unchanged. Several changes have been proposed to improve the efficiency of ozonation, such as the adoption of advanced ozone-based oxidation processes such as catalytic ozonation, O3/H2O2, O3/UV, O3/TiO2, ultrasonication aided ozonation, etc. (Comes et al. 2017). In view of the above, it is anticipated that through wastewater disinfection an appreciable removal of emerging contaminants can be achieved, therefore adding on their total abatement in wastewater treatment plants (when the removal through primary and secondary treatment is taken into consideration as well). Among different disinfection methods, chlorination and ozonation seem to provide better results, with regard to emerging contaminants removal, while ultraviolet as stand-alone

54

N. Morin-Crini et al.

disinfection method (without being upgraded to ultraviolet assisted advanced oxidation process) exhibit rather moderate performance. In order to guarantee satisfactory removal capacities, for as much as possible chemicals, high doses should be applied for all disinfection methods which exceed the ones that are typically used for the removal of pathogens. Further research is needed in order to conclude about the possible toxic characteristics of alternative disinfection methods by-products compared to those of their parent compounds. Conclusively, when disposal of treated wastewater in low dilution water streams (e.g. streams, rivers, lakes, shallow marine waters) is practiced or wastewater reuse is desirable (e.g. for agricultural use), the use of wastewater disinfection is mandatory to ensure a microbiologically acceptable water content. Optimizing disinfection methods to provide the removal of pathogens and the reduction of the emerging contaminant load that is released to the aquatic environment appears to be technologically feasible.

1.7.2  E  merging Contaminants in Industrial Wastewaters: Electrochemical, Photochemical and Ultrasonic Technologies for Their Removal For the scientific community, it is of great interest not only to study the presence of emerging substances in aqueous compartments (e.g. in terms of identification, quantification, behavior), but also to propose simple, adequate and effective methodologies for their elimination. Indeed, there is an urgent need for effective tertiary treatments capable of eliminating pharmaceuticals in biologically treated wastewater. This research theme is of particular interest to many research groups, as detailed below. Advanced oxidation processes represent one of the most promising strategies for the removal of emerging contaminants present in wastewater treatment effluents. Although advanced oxidation processes use different reagent systems, all techniques are based on the generation of reactive oxygen species such as the hydroxyl radical, which is a non-selective and very powerful oxidizing agent, used not only to degrade organic and inorganic substances but also to inactivate biological agents such as pathogenic microorganisms. Several types of advanced oxidation processes are presented in Fig.  1.9. The technology can improve biodegradability, enhance color removal, degrade and mineralize recalcitrant molecules, and reduce toxicity. Advanced oxidation processes are potential techniques for industrial wastewater treatment (e.g. removal of pollutants from pulp and paper and textile industry), and for drinking water production (to remove pathogens and organic compounds in combination with an adsorption step). Conventional municipal wastewater treatment plants have serious shortcomings that can also be addressed by advanced oxidation processes. However, these systems are not yet widely used in industry, mainly because of their high cost to treat large volume of effluents. The principles, performances, advantages, drawbacks and applications of advanced oxidation processes

55

1  Remediation of Emerging Contaminants

Oxidizing agent:

O3; H2O2; S2O82-

Catalyst:

homogeneous and heterogeneous

-

Energy:

UV radiation ionizing radiation electrical ultrasound, etc.

              

O3 in alkaline medium Catalytic ozonation H2O2/catalyst O3/H2O2/UV O3/H2O2/ultrasound S2O82-/catalyst S2O82-/UV •OH Photo-Fenton Heterogeneous photocatalysis Electrochemical oxidation Electro-Fenton Anodic oxidation O3/H2O2 Non-thermal plasma Etc.

Water emerging contaminants Partial or complete degradation

Organic compounds with lower weight and higher hydrophilicity;

CO2 + H2O + etc.

Fig. 1.9  Various types of advanced oxidation processes for wastewater treatment. (Source: Corina Bradu, Bucharest, Romania)

are detailed in numerous reviews (Ikehata et al. 2008; Yang et al. 2013; Oturan and Aaron 2014; Fernández-Castro et  al. 2015; Ribeiro et  al. 2015; de Araújo et  al. 2016; Mishra et al. 2017; Moreira et al. 2017; Miklos et al. 2018; Syam Babu et al. 2019; Wang and Zhuan 2020). Among the many advanced oxidation processes studied for the elimination of non-biodegradable compounds (the so-called refractory compounds), electrochemical, photochemical, photocatalytic and ultrasonic technologies are subject to particular attention. These technologies use single or combined methods (e.g. electro-Fenton, photo-electro-Fenton, sono-electrolysis etc.) in homogeneous (e.g. photo-Fenton: Fe2+/H2O2/UV) or heterogeneous systems (heterogeneous photo-­ catalysis: TiO2/H2O2/UV; anodic oxidation or photo-electrocatalysis). Different energy sources and catalysts are employed. The heterogeneous catalysis is often preferred to the homogeneous processes, due to the easier recovery of the catalyst. According to the energy source, there are three sub-types of processes, those using: (i) UV radiation (O3/UV, H2O2/UV, O3/H2O2/UV, or photo-Fenton Fe2+/H2O2/UV); (ii) ultrasound energy (O3/ultrasound, H2O2/ultrasound) and (iii) electrical energy (electrochemical oxidation, anodic oxidation and electro-Fenton). The general aspects of Fenton and photo-Fenton processes, electrochemical oxidation processes and sonochemistry were recently published by Ameta et al. (2018), by Radha and Sirisha (2018) and by Torres-Palma and Serna-Galvis (2018), respectively. Briefly, the Fenton process, based on the Fenton reagent, uses H2O2 and an iron soluble salt, generating hydroxyl radicals at atmospheric pressure and room temperature. High efficiency, relatively cheap reagents, no need of energy to activate H2O2 and the consequent easy implementation and operation are the advantages of such treatment. Some disadvantages are the generation of a secondary waste (sludge) and the narrow range of optimal pH (2.5–3.0). The photo-assisted Fenton process can be more efficient than Fenton alone, mainly due to the faster

56

N. Morin-Crini et al.

regeneration of Fe2+. Other related options are electro-Fenton, where Fe2+ is produced from sacrificial cast iron anodes, or even photo-electro-Fenton. Heterogeneous photocatalysis is another process that has been extensively investigated for water/ wastewater treatment and is based on the use of wide band-gap semiconductors which generate electrons and holes (and subsequent chain reactions including hydroxyl radicals) when irradiated with photons of energy higher than the semiconductor band-gap. The most widely used photocatalyst is TiO2, due to its outstanding activity, photochemical stability, good band gap energy, low cost and relatively low toxicity. Nevertheless, a drawback of the photo-assisted processes is the limited thickness of the water layer that is effectively penetrated by the UV radiation. Therefore, to increase the treatment efficiency shallow bed reactors should be used. These methods (Fenton process, photo-Fenton process, photocatalysis, ozonation-­ based processes) have been commonly studied, while others such as electrochemical technologies and sonolysis are less applied but deserve special attention in the literature. Advanced oxidation processes methods constitute a potential additional (secondary or tertiary) treatment for the elimination of pharmaceuticals, for example in wastewaters from the cities of Bogotá and Medellín (Colombia), due to their high elimination percentages. A chemical wastewater treatment using advanced oxidation processes may ideally produce the complete mineralization of organic pollutant, generating H2O, CO2 and other inorganic substances, or at least their transformation into more innocuous by-products. For example, the partial degradation of non-biodegradable organic substances can lead to biodegradable intermediates. For this reason, advanced oxidation processes can be used as pre-treatments before biological processes in a wastewater treatment plant. The electrochemical advanced oxidation processes have also been proposed for the degradation of pesticides and dyes, for the degradation of organic pollutants from wastewater and for water disinfection. Advantages often cited are: ease operation, high efficiency with possibility to mineralize compounds, no sludge production, possible coupling with other process, low temperature required for its operation, and possibility to use solar panel to decrease the energy consumption. Electrochemical technologies emerge also as a good alternative to carry out the on-site generation of disinfectant agents from the species naturally contained in wastewater. These technologies can be applied as a pre-treatment to transform recalcitrant compounds in biological wastewaters or in post-treatment before their discharge. However, until now most studies are conducted at laboratory scale and under controlled conditions. Further research and operational and investment costs assessment are necessary for scale-up new electrochemical technologies. Promising results using electrochemical, photochemical/photocatalytic and sonochemical processes have been published by Torres-Palma and collaborators (Giraldo et al. 2015; Serna-Galvis et al. 2016, 2019; Jojoa-Sierra et al. 2017; Valero et  al. 2017; Villegas-Guzman et  al. 2017; Torres-Palma and Serna-Galvis 2018). The authors studied the degradation of the antibiotic oxacillin in water by anodic oxidation with Ti/IrO2 anodes using an undivided stirred tank reactor (Giraldo et al. 2015). By using the best electrolyte and current density, complete oxacillin removal

1  Remediation of Emerging Contaminants

57

and total loss of antimicrobial activity was achieved after only 4 min of treatment and with low energy consumption. The degradation route and kinetics were indeed strongly affected by the supporting electrolyte (NaHCO3, Na2SO4, NaCl or NaNO3). Direct pathway was the most important route when NaHCO3 and Na2SO4 were used as supporting electrolytes. In presence of NaNO3, basic hydrolysis also contributed to the elimination of the pollutant. In the case of NaCl, the best supporting electrolyte, degradation via chlorinated oxidative species (mainly HOCl) was the most important route. The efficiency in the oxacillin degradation was not significantly affected by the initial pH (in the pH range of 3–9). Unfortunately, no mineralization was observed even after long exposure times (8 h). However, a decrease of 70% of the initial chemical oxygen demand was obtained and the level of biodegradability increased from 0.03 to 0.84, indicating that the system was able to transform the pollutant into highly oxidized and more biodegradable products with less antimicrobial activity. Additionally, different substances and radical scavengers present in wastewaters and natural water (glucose, isopropanol, and inorganic species) did not significantly affect the efficiency of the process. Finally, the more relevant initial aromatic by-products were identified and a degradation pathway of the electrochemical oxidation of the oxacillin antibiotic was proposed. Giraldo et al. (2015) concluded that electrochemical oxidation had a high potential to eliminate antibiotics. In another work, Serna-Galvis and co-workers demonstrated that high frequency ultrasound in the presence of additives was a selective and efficient advanced oxidation process to remove penicillinic antibiotics and to eliminate its antimicrobial activity from water (Serna-Galvis et al. 2016). The ultrasound is a sound wave with frequencies above 18 kHz and therefore cannot be heard by human ear. There are ultrasounds of high frequency, mainly used for medical applications, and ultrasounds of low frequency or power ultrasound, responsible for the phenomenon of cavitation. The sonochemical degradation of oxacillin was studied in simulated pharmaceutical wastewater. Oxacillin was transformed into by-products without antimicrobial activity and sonotreated water was completely mineralized using a subsequent biological process. Experiments showed that the antimicrobial activity was eliminated after 120 min of treatment and additives such as mannitol, calcium carbonate or their combination did not affect the sonochemical abatement of antimicrobial activity in terms of efficiency. A sonodegradation mechanism of oxacillin was proposed based on the evolution of four main by-products identified and their chemical structure. The identified by-products showed that the attack of the hydroxyl radical modified the penicillinic nucleus, which is the moiety responsible for the activity of the antibiotic. Although the ultrasound action also degraded part of by-­ products, the process was unable to achieve mineralization of the initial pollutant, even after a long period of ultrasonic irradiation (360 min). Nevertheless, the mineralization of the organic pollutants was completed through a subsequent biological treatment with a non-adapted microorganism from a municipal wastewater treatment plant. These results showed that the sonochemical treatment transformed the initial pollutant into substances that are biotreatable with a typical aerobic biological system. The authors concluded that the combination of a biological system with

58

N. Morin-Crini et al.

a sonochemical process was a promising alternative for the remediation of water containing oxacillin and other pharmaceutical additives. Torres-Palma’s group also studied the degradation of isoxazolyl penicillins by photo-Fenton, photocatalysis and ultrasound. The three processes achieved total removal of the antibiotic and antimicrobial activity, and increased the biodegradability level of the solutions (Villegas-Guzman et al. 2017). However, significant differences concerning the mineralization extent was observed depending on solution pH, chemical nature of additives (water matrix characteristics) and contaminant concentration. With TiO2 photocatalysis almost complete mineralization was reached, while ∼10% mineralization was obtained for photo-Fenton and practically zero for ultrasound activated process. Photo-Fenton and ultrasound processes were improved in acidic media (pH = 3), while natural pH favored TiO2 photocatalysis system. Bicarbonate and oxalic acid also improved photo-Fenton and ultrasound processes, respectively. According to both the nature of the added organic compound (e.g. glucose, 2-propanol and oxalic acid) and the pH of the process, inhibition, no effect or enhancement of the degradation rate was observed. The degradation in natural mineral water showed contrasting results according to the antibiotic concentration: ultrasound process was enhanced at low concentration of dicloxacillin followed by detrimental effects at high substrate concentrations. A contrary effect was observed during photo-Fenton, while TiO2 photocatalysis was inhibited in all of cases (Villegas-Guzman et al. 2017). In another work, Torres-Palma’s group investigated the elimination of the antibiotic norfloxacin in municipal wastewater, urine and seawater by electrochemical oxidation on Ti/IrO2 anodes (Jojoa-Sierra et al. 2017). This treatment was able to eliminate antimicrobial activity and the pollutant in all matrices. However, the results showed that matrices such as seawater or municipal wastewater containing chloride or bicarbonate ions exhibited a better degradation rate, while nitrate ions or urea, found in urine, reduced the efficiency of the process. Concerning the pH, the efficiency of the process in the presence of chloride ions followed the order: 9.0 > 7.5 > 6.5 > 3.0, showing a strong dependence on the antibiotic speciation. Anionic antibiotic form was more easily degraded than the zwitterionic and cationic forms. The antibiotic degradation occurred through both direct elimination at the electrode surface and mediated oxidation, via the electrogeneration of oxidative agents, such as active chlorine species and percarbonate ions, which came from chloride and bicarbonate oxidation, respectively. The mechanism of by-product formation was also studied to understand the efficiency of the technology to eliminate norfloxacin and its associated antimicrobial activity in complex matrices. The identification of three primary norfloxacin by-products demonstrated that the initial attack of the active chlorine species, mainly HOCl, occurred at the secondary amine of the piperazine ring followed by chlorination of the benzene ring. Torres-Palma’s group also demonstrated that electrochemical advanced oxidation was a pertinent approach for Staphylococcus aureus disinfection in municipal wastewater treatment plants (Valero et al. 2017). Staphylococcus aureus is a facultative anaerobic, Gram-positive, coccoid bacterium, which has the capacity to colonize almost every tissue of the human body, causing different diseases. Pathogenic

1  Remediation of Emerging Contaminants

59

organisms can be found in municipal effluents at relatively high concentration since the elimination of microbes is not an objective of conventional wastewater treatment plants. Thus, antibiotic-resistant pathogens and their genes can be easily spread through wastewaters. The results published by Valero et al. (2017) showed that the photo-electro-Fenton process led to a bacterial reduction of −0.9 log units and a dissolved organic carbon reduction of 14%, while −5.2 log units of bacteria and 26% of dissolved organic carbon were removed by using the photo-electro-­ Fenton process. The increase in current intensity in the photoelectro-Fenton system augmented the production of H2O2, resulting in increased bacterial inactivation. However, mineralization extent slightly increased or remained practically the same. When comparing the influence of Fe2+ and Fe3+ on photo-electro-Fenton, similar Staphylococcus aureus inactivation was observed, while dissolved organic carbon removal was higher with Fe2+ (31%) than with Fe3+ (19%). This water disinfection technology is of interest because in the reclamation of municipal wastewater, a disinfection stage is always required to obtain a high quality effluent, as established by the Colombian legislation. The degradation of seventeen contaminants of emerging concern in real effluents from the municipal wastewater treatment plant of Bogotá by sonochemical advanced oxidation processes, was studied by Serna-Galvis et al. (2019). The municipal water treatment plant operates at 4 m3/s, about 350,000 m3/day, with a removal efficiency of 40% biological oxygen demand and 60% of suspended solids. However, it is not efficient to completely remove emerging contaminants, as confirmed by a previous work of the same research group (Botero-Coy et al. 2018). Analytical monitoring of the effluents showed the presence of several pharmaceuticals with variable concentrations, in the range of 0.01–2 μg/L (Serna-Galvis et al. 2019). Among the substances, the antihypertensives (valsartan, losartan) were those at the highest concentrations followed by a cocaine metabolite (benzoylecgonine), an anti-­ inflammatory (diclofenac) and the antibiotics ciprofloxacin and norfloxacin. The obtained results confirmed that ultrasound water treatment is an efficient alternative method for the treatment of these emerging contaminants present in low concentrations in the effluent of the municipal wastewater of Bogotá treatment plant. For example, after 30 min of treatment the concentration of sulfamethoxazole, an antibiotic used to treat a variety of bacterial infections, notably diminished from 0.37 to 0.14 μg/L. The ultrasonic system complemented with iron (II), UVA light and oxalic acid presented a strong potential for the contaminant elimination in real and complex effluents. The main disadvantage of the method is however the high consumption of electrical energy. Martínez-Pachón et al. (2019) studied the advanced oxidation of antihypertensives losartan and valsartan by photo-electro-Fenton at near-neutral pH using natural organic acids and a dimensional stable anode-gas diffusion electrode system under light emission diode lighting. Valsartan and losartan are used in the treatment of hypertension diseases, and are considered among the emerging contaminants that are difficult to treat. In their experiments, organic acids as citric, tartaric and oxalic acids were used as complexing agents of iron ions in order to maintain the performance of the Fenton reaction at near-neutral pH value. The authors showed that

60

N. Morin-Crini et al.

after 90 min of electro-Fenton treatment using the optimized conditions, a degradation of 70% of valsartan and 100% of losartan were achieved. The total degradation of the two antihypertensives was achieved with a photo-electro-Fenton for the same time period. The degradation performance was attributed to the increase of the initial dissolved iron in the system in the presence of the organic ligands, facilitating the Fe3+/Fe2+ turnover in the catalytic photo-Fenton reaction and consequently, hydroxyl radical production. The increased photo-activity of the complexes was also associated with their high capability to complex Fe3+ and to promote ligand-to-­ metal charge transfer, which was of key importance to feed Fe2+ to the Fenton process. The results showed that the system evaluated was more efficient to eliminate sartan family compounds using light emission diode lighting in comparison with traditional UV-A lamps used in this type of works. Moreover, three transformation products of valsartan degradation and two transformation products of losartan degradation were identified by high-resolution mass spectrometry using hybrid quadrupole-­time-of-flight mass spectrometry. The several organic compounds still present at the end of the photo-electro-Fenton treatment were effectively treated in a subsequent aerobic biological system (Martínez-Pachón et al. 2019). The review of the literature related to wastewater treatment issue revealed that the electrochemical, photochemical and ultrasonic based technologies have a tremendous power to degrade emerging pollutants such as antibiotics, although some processes are not capable to completely mineralize the organic pollutants. The two main advantages of these advanced oxidation processes are: (i) the oxidizing species are generated in situ (no need of chemicals storage and handling) and (ii) most of processes (except the Fenton type) do not require a rigorous control of solution pH. Other advantages are: operation control simplicity, reactor design compactness, adaptability of the technology to various organic loads of wastewater and effectiveness in disinfection. For the electrochemical treatment, one of the main challenges to its successful implementation for industrial application is to reduce energy consumption and cost (including the cost of electrodes). Another important challenge is to improve the understanding of the reaction mechanisms and in particular to identify potential toxic intermediates. It is also important to improve the long-term stability and electrolytic efficiency of the materials. In this respect, progress is expected from the application of nanotechnologies.

1.7.3  C  arbon Nanomaterials in Catalytic Ozonation of Emerging Contaminants An effective technology to destroy organic emerging contaminants is catalytic ozonation. This treatment uses a catalyst to decompose ozone into a number of strong oxidant radicals such as .OH (Wang et al. 2016c). These radicals quickly oxidize organic compounds (Fig. 1.10).

1  Remediation of Emerging Contaminants

61

CO2 H2O

O3

. OH

CATALYST

.O 2

Pollutants

Compounds with lower weight

Fig. 1.10  Radical pathway of catalytic ozonation process. (Source: Mohammad Mahmudul Huq, Saskatoon, Canada)

The principal component of a catalytic ozonation process is the catalyst. These catalysts could primarily be categorized into two: homogenous or heterogeneous. Homogeneous catalysts are water soluble metal salts which are not recoverable after an ozonation process, thus their application involves chemical costs as well as further pollution. On the contrary, heterogeneous catalysts are solid particles that can be recovered, regenerated and reused. Numerous works have reported different kinds of materials as catalysts in water treatment processes, namely metal oxides, activated carbon and carbon nanomaterials (Restivo et al. 2012, 2013, 2016; Rocha et al. 2015; Wang et al. 2016b, c). Many reports have shown superior performance of metal oxide-based catalysts. However, these catalysts can be subject to metal leaching into effluent water (Rocha et al. 2015). Consequently, increasingly more research effort has been made for the development of metal-free catalyst. Among metal-free catalysts, carbon materials such as activated carbon, multi-walled carbon nanotubes, graphene oxide and reduced-graphene oxide are the most studied ones. These materials have been studied both as sole catalysts and as support materials. This section focuses fundamentally on the studies that concern application of carbon-­based nanomaterials (i.e. multi-walled carbon nanotubes, graphene oxide and reduced-graphene oxides) both as catalysts and catalyst support materials in catalytic ozonation processes. Among carbon nanomaterials, multi-walled carbon nanotubes are the most widely studied catalyst for ozonation processes due to their strong catalytic activity and re-usability. Multi-walled carbon nanotubes were first reported as catalyst in ozonation process by Liu et al. (2009). These materials achieved 80% conversion of oxalic acid in 40 min as compared to around 2% conversion by non-catalytic ozonation. The authors also studied the catalytic performance of oxidized multi-walled carbon nanotubes. For this purpose, nanotubes were oxidized by pre-ozonation treatment at various degrees. This pre-ozonation treatment implanted different acidic groups (such as -COOH, -OH, etc.) on catalyst surface leading to inferior performance of multi-walled carbon nanotubes. The catalytic activity decrease can be attributed to the negative charge generated by the acidic groups which causes a

62

N. Morin-Crini et al.

drop of pHpzc value nearly to the pH of oxalic acid aqueous solution. This leads to less adsorption of anion species of oxalic acid, eventually to less degradation of oxalic acid during catalytic ozonation. In a similar study, Gonçalves et al. (2010) investigated the role of surface chemistry of multi-walled carbon nanotubes on catalytic ozonation of oxalic acid. This study gave a very similar picture as the previous one in which the surface acidic sites hinder multi-walled carbon nanotubes’ catalytic ability. The authors also compared the performances of multi-walled carbon nanotubes with those of commercial activated carbon, concluding that multi-walled carbon nanotubes are more effective catalysts since they impose less internal mass transfer limitations on the reactants during the catalytic ozonation reaction. A catalytic reaction is mass transfer limited when the participating reactants experience mass transfer resistance inside micro-­ pores before reaching the active sites. Activated carbons possess a large quantity of micro-pores, which is likely to impose significant mass transfer resistance. On the other hand, multi-walled carbon nanotubes barely possess any micro-pores. The same research group compared the catalytic performance of commercial multi-­ walled carbon nanotubes with that of activated carbon in catalytic oxidative degradation of sulfamethoxazole, an antibiotic (Gonçalves et  al. 2013). In terms of sulfamethoxazole degradation, catalytic ozonation with multi-walled carbon nanotubes was comparable to non-catalytic ozonation. Nevertheless, catalytic ozonation led to higher degree of mineralization expressed as total organic carbon removal. Surprisingly, activated carbon showed superior catalytic activity compared to multi-­ walled carbon nanotubes in terms of total organic carbon removal. This result was ascribed to the fact that sulfamethoxazole is more readily adsorbed in micro-pores and not to its oxidative degradation. The authors also studied the toxicity effects of sulfamethoxazole oxidation by-products by Microtox® bioassays. It was found that the lowest toxicity is achieved when the ozonation was carried on in the presence of multi-walled carbon nanotubes. Restivo et al. (2012) studied the activity of multi-walled carbon nanotubes in the ozonation of metolachlor, an herbicide. The use of the carbon based material did not show much improvement of the single ozonation process in the removal of target pollutant. Both catalytic and non-catalytic oxidation products of metolachlor showed higher toxicity when compared to the toxicity of non-treated metolachlor, as shown by Microtox® bioassay test. However, the end-products were found less toxic with catalytic ozonation when compared to the non-catalytic ozonation. The degradation of atrazine (an herbicide) by multi-walled carbon nanotubes catalyzed ozonation was found to be somewhat slower (lower apparent reaction rate constant) than single ozonation (Fan et al. 2014). Nevertheless, in terms of total organic carbon and toxicity reduction, catalytic ozonation was found superior. One of the main research directions in this field is to improve the catalytic activity of carbon nanotubes by modifying their surface chemistry and specific surface area by different methods (e.g. heteroatom-doping, oxidation, grinding, etc.). Qu et al. (2015) studied carboxylated carbon nanotubes in catalytic ozonation of indigo. The modified nanotubes showed higher indigo removal in terms of indigo concentration, total organic carbon and toxicity removal. Yet, this study is inconclusive as

1  Remediation of Emerging Contaminants

63

it did not compare the results with those obtained with pristine carbon nanotubes. Soares et  al. (2015) improved multi-walled carbon nanotubes’ catalytic performance by increasing their specific surface area by shortening their tube size. This size reduction was achieved by ball-milling. The ball-milled multi-walled carbon nanotubes showed fairly improved performance in catalytic ozonation of oxalic acid. Rocha et al. (2015) showed that reduced graphene oxide works as a catalyst in ozonation of oxalic acid. Catalytic ozonation with reduced-graphene oxide turned out to be 50% more efficient in the pollutant removal than non-catalytic ozonation (for 120 min reaction time). Wang et al. (2016b) used reduced-graphene oxide as catalyst for ozonation of p-hydroxybenzoic acid when nearly full mineralization of p-hydroxybenzoic acid was achieved in 60  min. This study identified carbonyl groups as the active sites and suggested that superoxide radicals (.O2−) and singlet oxygen (1O2) are the dominant species responsible for p-hydroxybenzoic acid degradation. Graphene oxide, the parent material of reduced-graphene oxide, can also be catalytically active in many reactions as its surface is rich in oxygen functional groups. For instance, the ozonation of N,N-diethyl-m-toluamide, a widely used pesticide, in the presence of graphene oxide was investigated by Liu et al. (2016). This process showed increased removal of the target pollutant as compared to non-­ catalytic ozonation. But lack of mineralization data raises doubts on its applicability. Ahn et  al. (2017) showed that over-oxidized graphene oxide produces significantly higher amounts of .OH than graphene oxide or reduced-graphene oxide, which is likely to lead to higher removal of recalcitrant organics. Song et al. (2019b) studied both graphene oxide and reduced-graphene oxide as catalysts for catalytic ozonation of p-chlorobenzoic acid and benzotriazoles. They found graphene oxide more efficient in degrading p-chlorobenzoic acid while reduced-graphene oxide showed higher activity in the case of benzotriazole. However, the authors noted that graphene oxide suffers from gradual degradation during the catalytic ozonation process. This deterioration of graphene oxide material is most likely caused by corrosive attack of ozone or .OH (Radich et al. 2014). In an extensive study, Wang et al. (2018c) synthesized reduced-graphene oxide starting from graphite from used lithium ion battery. The obtained material showed higher removal of oxalic acid than commercially reduced-graphene oxide. A strong correlation between the amount of defective sites on reduced-graphene oxide and its catalytic activity was found. This correlation was again justified by density functional theory calculations. Heteroatom doping of graphene oxide and of reduced-graphene oxide can significantly improve their catalytic activity. N, B, P and S have been studied as dopants for reduced-graphene oxides. Rocha et  al. (2015) doped reduced-graphene oxide with N from different nitrogen precursors such as melamine and urea. This doping process implanted three N containing functionalities into the reduced-­ graphene oxide namely, pyridinic, pyrrolic and quaternary-N. The improved performance of the N-doped reduced-graphene oxide in mineralizing oxalic acid and phenol was attributed to these functional groups which are said to work as active centers. Moreover, these N-functional groups increase the pHPZC, making the material more positively charged at the pH of oxalic acid solution (3.0). This enables

64

N. Morin-Crini et al.

increased adsorption of oxalate anions (which is the dissociated form of oxalic acid at pH of 3.0), favoring its catalytic surface reaction. This could also explain the lack of improvement in the phenol degradation, which is found in molecular form at the aqueous solution pH (pKa around 10). Bao et al. (2016) also confirmed improved catalytic ozonation performance of N-doped reduced-graphene oxide. Yin et  al. (2017) showed catalytic ozonation using N and P doped reduced-graphene oxides is far more efficient than non-­ catalytic ozonation in the degradation of sulphamethoxazole. Wang et al. (2019a, b) used a novel microwave method to dope N into reduced-graphene oxide. This method generates higher amounts of N-doping that eventually leads to higher removal of 4-nitrophenol and oxalic acid. The improved performance of this microwave method was attributed to the higher degree of graphene oxide reduction and generation of more defects and carbon dangling bonds. In a very detailed study, Song et al. (2019a) synthesized and tested N, P, B and S-doped reduced-graphene oxide or catalytic ozonation degradation of p-­chlorobenzoic acid and benzotriazole. They also studied bromate (BrO3−) elimination capacities of these processes. Bromates are carcinogenic by-products of many ozonation processes. Although P-doped reduced-graphene oxides showed the fastest elimination of both p-chlorobenzoic acid and benzotriazole, the authors concluded that in terms of normalized pseudo-first order reaction constant (kobs), N-doped reduced-graphene oxide shows the fastest removal. The normalization was done by dividing the kobs by atomic percentage of the corresponding heteroatoms in each of the as prepared reduced-graphene oxide. On the other hand, S-doped reduced-graphene oxide is found to be unstable as it increases the total organic carbon content of the aqueous solution during the catalytic ozonation process. Wang et al. (2018b) doped multi-walled carbon nanotubes with F using HF as F precursor. The materials showed significantly increased catalytic activity in ozonation of oxalic acid as compared to N-doped multi-walled carbon nanotubes. Highly electronegative active sites like N and O are inferred to decompose ozone by nucleophilic attack. Interestingly, this study showed that an excessive doping of electronegative atoms is counter productive. Nevertheless, given the difficulty in handling highly toxic HF, F doping may not be feasible at commercial level. Usually, catalytically active metals and metal oxides are deposited on porous materials such as alumina, zeolite, activated carbon, silica and various metal oxides (Ghuge and Saroha 2018). Carbon nanomaterials such as multi-walled carbon nanotubes and graphene oxide/reduced-graphene oxides have also been used as support materials because of their excellent compatibility with metals and metal oxides, resistance to adverse environment, mechanical strength and excellent electron transfer ability (Lin et  al. 2011; Khan et  al. 2015). Studies have also shown synergy between active catalyst materials and these carbon materials (Sampaio et al. 2015). This synergy has been exploited in a few catalytic ozonation studies as well. The greatest number of works deals with the synthesis of supported manganese and iron oxides and their use in the oxidative degradation of emerging pollutants such as pesticides and pharmaceuticals (Sui et al. 2012; Li et al. 2015; Bai et al. 2017; Wang et al. 2016a). Sui et al. (2012) synthesized a MnOx/multi-walled carbon

1  Remediation of Emerging Contaminants

65

nanotube composite for catalytic ozonation of ciprofloxacin, a persistent antibacterial agent. MnOx/multi-walled carbon nanotube showed 87.5% ciprofloxacin removal in 15 min as opposed to 40.2% removal with unsupported MnOx and 26.7% removal with non-catalytic ozonation. Li et al. (2015) reported the synthesis of a sea urchin-like α-MnO2/reduced-graphene oxide composite for catalytic ozonation of bisphenol A. The process showed significantly higher bisphenol A removal efficiency compared to that with pristine α-MnO2, pristine reduced-graphene oxide and non-catalytic ozonation. The authors concluded that reduced-graphene oxide is inactive in catalytic ozonation of bisphenol A. Wang et  al. (2016a) synthesized γ-MnO2/reduced-graphene oxide composite for catalytic ozonation of 4-­nitrophenol, showing improved degradation and mineralization compared to non-catalytic ozonation. Prepared γ-MnO2/reduced-graphene oxide showed improved performance over commercial MnO2. Wang et  al. (2019a) prepared a CeO2/oxidized-carbon nanotube composite for catalytic ozonation of phenol. The composite catalyst achieved nearly 100% total organic carbon removal in 60  min with virtually no activity loss up to 5 cycles. Depositing reduced-graphene oxide on metal oxide can also be fruitful as shown by Ren et al. (2018). In this report, MnFe2O4 nano-fiber catalyst was improved by reduced-graphene oxide doping. Overall, catalytic ozonation processes with carbon nanomaterial-based catalysts offer the following advantages: (i) higher degree of organic pollutant degradation as compared to activated carbon catalysts (ii) minimal metal leaching issue and (iii) compatibility under a wide range of conditions. However, the separation method of catalyst from treated water is an important issue to be addressed before technology scale-up for industrial application.

1.7.4  N  on-thermal Plasma: A New Candidate Water Treatment for the Removal of Emerging Contaminants Growing interest in finding effective solutions for the removal of emerging contaminants from water led to the investigation of unconventional water treatment methods. Among them, non-thermal plasma is a promising approach and is now considered the youngest member of the so-called advanced oxidation processes family. Recently, significant research efforts were devoted to enhancing the efficiency of plasma treatment of water contaminated with harmful organic pollutants such as pharmaceuticals and pesticides (Magureanu et  al. 2018). The efforts are directed to elaborate technical and economical feasible solution, but also to bring new insight from the point of view of reaction mechanism and final characteristics and quality of the treated water. Non-thermal plasma can be generated directly in the liquid or in the gas phase. Plasmas in contact with water are very complex systems, which produce a diversity of molecular, ionic and radical reactive species responsible for the degradation of organic pollutants. In these systems, the generation of reactive species is initiated by the collision of highly energetic electrons (formed by the electrical discharge) with

66

N. Morin-Crini et al.

the gas constituents and water molecules (in the gas phase, or at the gas-liquid interface). The nature of active species and their availability in the liquid phases depend on the plasma reactor configuration (e.g. corona, dielectric barrier discharges, gliding arc, plasma jet etc.), on the discharge characteristics, solution properties and gas composition (Park et al. 2013). A number of reactive oxygen species and reactive nitrogen species have been detected in gas and liquid phase of the cold plasma discharge systems, such as: ∙OH, HO2∙, H2O2, O∙, ∙O2−, O3, ∙NO, ∙NO2, NO2−, NO3−, ONO2− etc. (Lukes et al. 2012; Locke et al. 2012; Bruggeman et al. 2016). Ozone, hydrogen peroxide and hydroxyl radical are the most investigated reactive oxygen species in the plasmas in contact with water systems. Numerous studies deal with the identification and quantification of these three species by different spectrophotometric and chromatographic methods (Ono and Oda 2002; Park et al. 2006; Xiong et  al. 2015; Guo et  al. 2019a; Kanazawa et  al. 2013; Marotta et  al. 2011; Lukes et al. 2004; Bilea et al. 2019). The hydroxyl radical is a key reactive oxygen species, considered to be the main actor in the oxidative degradation of organic pollutants. It is actually the species that makes the plasma process to be framed as an advanced oxidation process. Due to its short lifetime, ∙OH interacts with organic pollutants very near the gas-water interface (Kanazawa et al. 2011; Marotta et al. 2011; Ajo et al. 2015). In the bulk liquid phase, hydrogen peroxide, a stable molecular reactive oxygen species, is formed mainly by recombination of hydroxyl radicals (Locke and Shih 2011). The accumulation of H2O2 in water during the plasma treatment was highlighted by numerous works (Locke and Shih 2011; Magureanu et al. 2016; Bradu et al. 2017; Bilea et al. 2019). The production and loss mechanisms of ∙OH in discharges formed over water are discussed in detail in (Bruggeman and Schram 2010). The radical recombination reaction leading to stable molecular species may limit the process effectiveness (Locke et al. 2012). However, the hydrogen peroxide is a potential source of hydroxyl radicals, which could be used judiciously. For instance, H2O2 decomposition with the regeneration of ∙OH radicals could be promoted by an adequate catalyst. Indeed, numerous research groups report advantages of the combined plasma-catalytic oxidation process in the degradation of recalcitrant pollutants in water (Parvulescu et al. 2012 Jović et al. 2014; He et al. 2014; Hama Aziz et al. 2018; Guo et al. 2019a). The proposed catalysts are soluble transitional metal salts (e.g. Fe2+, Fe3+, Mn2+ and Co2+ salts) (Dojčinović et al. 2011; Jović et al. 2014) or heterogeneous catalysts (e.g. TiO2, activated carbon and graphene based materials) (He et  al. 2015; Hama Aziz et  al. 2018, Vanraes et  al. 2015, 2017; Guo et  al. 2019a, b, c). Another long-lived species generated in non-thermal plasma in oxygen-­ containing gaseous atmosphere is ozone. Appreciable O3 concentrations were detected in the gas phase of electrical discharges in contact with water (Lukes et al. 2005; Marotta et al. 2011; Magureanu et al. 2016). However, ozone diffusion across the gas-liquid interface is limited and its concentration in water is often under the detection limit of employed analytical methods (Marotta et al. 2011; Dobrin et al. 2013; Magureanu et al. 2016). Nevertheless, the transfer of plasma generated O3 from the gas to the liquid could be improved to facilitate its reaction with the target

1  Remediation of Emerging Contaminants

67

organic pollutants or their degradation by-products (both by direct: involving molecular ozone attack; and indirect path-way: involving ∙OH). In this respect, plasma-ozonation systems were proposed (Magureanu et  al. 2016; Bradu et  al. 2017). In this case, the water to be treated is continuously circulated between the two reactors: a plasma reactor (in which the electrical discharge is produced over water surface) and second reactor in which the effluent gas from the plasma reactor was bubbled into the aqueous solution (Fig.  1.11). With this combined plasma-­ ozonation system faster removal of the target compound (e.g. methylparaben and herbicide 2,4-D) and higher degree of mineralisation was obtained compared to the single plasma or ozonation processes. The improvement was attributed to the

Fig. 1.11  Experimental set-up for plasma-ozonation system. (Monica Magureanu, Magurele, Romania)

68

N. Morin-Crini et al.

enhanced transfer of O3 in the water and its further interaction with the H2O2 accumulated in solution, leading to increased generation of ∙OH (peroxone reaction). When non-thermal plasma is generated in air, beside reactive oxygen species, reactive nitrogen species can be formed in significant amounts. Nitric oxide (∙NO) is an important secondary species, produced through the interaction of the parent (N2) and/or primary species (N∙, O∙, ∙OH) in gas phase. In further reactions, ∙NO can effectively fix oxygen atoms through the interaction with O∙ or with O-donor species produced in the discharge (e.g. O3 and HO2∙) to form nitrogen dioxide (∙NO2) (Brisset and Pawlat 2016; Aritoshi et al. 2002). In humid air and in the aqueous phase, the formation of nitrous and nitric acids (HNO2 and HNO3) takes place. The presence of peroxynitrous acid (ONOOH) in water was also reported. It was suggested that this peroxy-acid is produced in the reaction between nitrous acid and hydrogen peroxide (reaction favoured in acidic pH), or via the reaction between dissolved nitric and nitrous oxides and different radical reactive oxygen species (∙OH, HO2∙ and ∙O2−) (Goldstein et al. 2005; Moussa et al. 2007; Lukes et al. 2012; Tian and Kushner 2014). The generation of peroxynitric acid (O2NOOH) through the reaction between peroxynitrous acid and hydrogen peroxide has also been proposed (Boehm et al. 2018; Ikawa et al. 2016; Nakashima et al. 2016).Thus, a variety of nitrogen-containing species is present in the aqueous phase as well. The reactive nitrogen species accumulation in water depends on their solubility and their lifetime. The dominant species for electrical discharges in contact with water are considered to be the nitrogen oxy- and peroxy-acids and their conjugate ions: HNOx/ NOx− (Tian and Kushner 2014). Detailed analysis of the generation, transport and interactions of reactive nitrogen species in plasma in contact with water can be found in (Locke et al. 2012; Bruggeman et al. 2016; Bradu et al. 2020). Even if the reactive nitrogen species involvement in the organic compounds degradation was less studied, there are solid evidences that these species participate in the degradative oxidation pathway of water pollutants. As an example, nitro-­ substituted by-products and N,N-dimethyl-nitroaniline have been detected in the degradation of methyl orange in corona discharge in contact with water by Cadorin et al. (2015). It was assumed that the organic molecules interact with peroxynitrous acid either directly or indirectly via dissociation into ∙NO2 and ∙OH (Moussa et al. 2007; Cadorin et al. 2015). A general scheme for the reactive oxygen species and reactive nitrogen species dynamic in a non-thermal plasma system for water treatment is presented in Fig. 1.12. A large variety of discharge configurations has been used to produce plasma in liquid or gas-liquid environments, and can be classified into two categories: discharges generated directly in liquid (i.e. with both electrodes submerged in liquid) and discharges generated in the gas phase, in contact with liquid (Bruggeman and Leys 2009; Jiang et al. 2014; Magureanu and Parvulescu 2016; Locke et al. 2012). The early studies on plasma removal of aqueous pollutants addressed mainly organic dyes, due to facile observation of the solution decolorization (Malik et  al. 2002; Sugiarto et al. 2003; Burlica et al. 2004; Grabowski et al. 2007; Magureanu et al. 2007, 2008; Stará et  al. 2009). An analysis of reported results revealed that the

1  Remediation of Emerging Contaminants

69

Fig. 1.12  Schematic representation of reactive oxygen species and reactive nitrogen species dynamic for non-thermal plasma in contact with water. (Monica Magureanu, Magurele, Romania; Corina Bradu, Bucharest, Romania)

energy efficiency of the process is significantly higher for the plasma in gas phase as compared to discharges directly in liquid, especially for large surface to volume ratio of the solution to be treated, i.e. in case of thin liquid films or liquid spray (Malik 2010). Most of the recent studies on plasma degradation of water contaminants have been carried out using electrical discharges generated in gas phase in contact with liquid. Very simple geometries, such as corona, dielectric barrier discharges or gliding arc above liquid, have often been reported for the removal of antibiotics (El Shaer et al. 2020; Smith et al. 2018; Sarangapani et al. 2019; Xu et al. 2020; Zhang et al. 2018; Acayanka et al. 2019) and pesticides (Hijosa-Valsero et al. 2013; Hu et al. 2013; Li et al. 2013; Singh et al. 2016, 2017). Since the penetration depth of the plasma-generated reactive species into the solution is very small, in such configurations the volume of plasma-treated solution is generally very small, in the milliliter range (Smith et al. 2018; Zhang et al. 2018; Xu et al. 2020). Smith et al. (2018) reported the complete removal of ampicillin in 1 mL solution of high concentration (20 mM) after only 3 min of treatment with a dielectric barrier discharge above liquid. No information on the discharge power is provided, and thus an evaluation of the degradation efficiency is not possible. If treatment of larger solution volumes is attempted the time required for pollutant removal becomes much longer. Using a pin-to-water corona discharge, El Shaer et al. (2020) obtained almost complete removal of doxycycline with concentration of 50 mg/L in 50 mL water after 90  min treatment, while the degradation of oxytetracycline was even slower. Acayanka et al. (2019) needed 120 min to remove approximately 80% of the initial amoxicillin in 500 mL 0.1 mM aqueous solution using a gliding arc above liquid. Obviously, the treatment time is not an accurate measure of the efficiency of the plasma process, since the degradation depends on a number of factors, such as the molecular structure of the target compound, its concentration, the solution volume

70

N. Morin-Crini et al.

and properties, the gaseous atmosphere, the input power and the discharge characteristics, to name only a few. An example to illustrate the large extent of this influence is the comparison between the above-mentioned results of El Shaer et  al. (2020) and the data reported by Singh et al. (2016, 2017), who also used a pin-to-­ water corona to degrade the pesticides carbofuran and 2,4-D and achieved complete removal within less than 10 min treatment. Information on the energy efficiency of the removal process was provided by Xu et al. (2020), research group that investigated the degradation of norfloxacin (initial concentration 10 mg/L in 10 mL water) by a dielectric barrier discharge. Although fast removal of the antibiotic has been achieved (4  min), the reported efficiency (defined as the amount of pollutant removed per unit of energy consumed in the process) was rather low, i.e. in the range of tens of mg/kWh. The addition of H2O2 and Fe2+ significantly improved the results, optimum catalyst concentration leading to the reduction of treatment time to 0.5 min. The positive effect of iron catalyst was attributed to the Fenton reaction in the presence of plasma-generated H2O2 (Li et al. 2013; Jović et al. 2014; Hama Aziz et al. 2018; Xu et al. 2020). A comparison between the corona discharge above liquid and a corona generated in gas bubbles inside the solution demonstrated much faster degradation of the target antibiotics in the second case (El Shaer et al. 2020). This discharge geometry, with one or several hollow needles submerged in liquid as high voltage electrode and plasma produced in gas bubbles at the tip of the needles, has been employed by several research groups for the removal of antibiotics, mostly in combination with catalysts (Wang et al. 2018d; Guo et al. 2019a, b, c). A slightly different configuration, with the gas blown through a tube containing the high voltage needle electrode has also been used (He et al. 2014, 2015; Hu et al. 2019). Several authors produce plasma in the gas and bubble the effluent gas through the solution to be treated (Kim et al. 2013; Kim et al. 2015; Lee et al. 2018; Tang et al. 2018a, b, 2019; Wang et al. 2019c; Li et al. 2020). It is unlikely that highly reactive species with short lifetime would reach the liquid, so in this case plasma is simply used as a source of ozone. Besides the generation of large amounts of reactive species in the plasma, their efficient transfer to the treated liquid is essential for the enhancement of pollutants removal efficiency. This has been demonstrated for instance in the experiments of Acayanka et al. (2019), where the authors compared the degradation of amoxicillin by a gliding arc plasma above the target solution with the results obtained when the solution is sprayed through the discharge region. Faster degradation of the antibiotic has been reported for the spray configuration than in the batch mode (i.e. three times larger rate constant) and 2.5 times higher energy yield. Hijosa-Valsero et al. (2013) have also confirmed the importance of large surface-to-volume area by comparing the removal of several pesticides in water using either a batch dielectric barrier discharge geometry or a dielectric barrier discharge with falling liquid film. The energy efficiency for the removal of atrazine is 10 times higher in the configuration with liquid circulation, while for the insecticides lindane and chlorfenvinfos, it exceeds one order of magnitude. Such more elaborated reactor design has been extensively investigated for the degradation of various pesticides (Hijosa-Valsero et al. 2013; Jović et al. 2013, 2014; Vanraes et al. 2015, 2017; Bradu et al. 2017; Yu et al. 2017;

1  Remediation of Emerging Contaminants

71

Hama Aziz et al. 2018) and antibiotics (Magureanu et al. 2011; Rong and Sun 2014; Rong et al. 2014; Xin et al. 2016; Iervolino et al. 2019). Most authors used coaxial geometry, with the liquid pumped upwards through a cylindrical inner electrode and then flowing as a thin film on the outer surface of this tube, thus being in direct contact with plasma. However, planar geometries have also been employed, with the liquid film flowing either horizontally (Xin et al. 2016) or vertically (Hama Aziz et al. 2018) between the electrodes. One of the challenges in the dielectric barrier discharge with falling liquid film is to produce a stable thin film of liquid. Hama Aziz et al. (2018) obtained a homogeneous and stable solution layer of thickness estimated to 150 mm, flowing along large area (68 × 29 cm) glass sheets and used this planar dielectric barrier discharge configuration to degrade the herbicide 2,4-­ D. A comparison between plasma treatment and other advanced oxidation processes from the point of view of the energy yield revealed that the dielectric barrier discharge in combination with Fenton oxidation is the most efficient treatment process, followed in this order by ozonation, plasma alone, photocatalytic ozonation and, at last photocatalysis. Although the good removal efficiency of ozonation is confirmed, the authors mention its major drawback related to low mineralization. For improved degradation of the target compound and its intermediate oxidation products, either photocatalytic ozonation, or plasma combined with Fenton oxidation are recommended (Hama Aziz et al. 2018). It is now generally accepted that the addition of Fe2+ to the plasma treatment significantly improves the removal of aqueous contaminants, another example being the herbicides mesotrione and sulcotrione (Jović et al. 2013; Jović et al. 2014). In this case it has been found that the effect of Fe2+ exceeds that of Mn2+ and Co2+. The combination of plasma with TiO2 catalysts also appears successful for pollutants removal. He et al. (2014) reports a considerable rise in the removal rate of the antibiotic tetracycline, from 61.9% to 85.1%, accompanied by the enhancement of total organic carbon removal, from 25.3% with the plasma alone to 53.4% in the presence of TiO2. Another approach to increase the efficiency of pollutants removal by plasma adopted by Vanraes et al. (2015, 2017) is to locally enhance the pollutant concentration in the plasma region. They used a coaxial dielectric barrier discharge with falling film and added an activated carbon textile mesh with extremely large surface area over the inner electrode. This highly adsorptive material was found to significantly contribute to the removal of target compounds in plasma. A method to improve the mass transfer of the plasma-generated ozone into the treated liquid is to bubble the effluent gas from the plasma through the solution (Gerrity et al. 2010; Magureanu et al. 2011; Bradu et al. 2017). Therefore, as presented earlier, dual plasma-ozonation systems have been developed. Very high energy efficiency has been reported in such systems, either employing dielectric barrier discharge with falling liquid film for the removal of β-lactam antibiotics (reaching 105 g/kWh for amoxicillin) (Magureanu et al. 2011), or with a corona discharge above liquid to degrade the herbicide 2,4-D (5 g/kWh) (Bradu et al. 2017). One of the most efficient plasma systems reported up to now is based on a pulsed corona reactor similar to an electrostatic precipitator, with the liquid introduced as

72

N. Morin-Crini et al.

droplets or jets through the plasma zone and short high voltage pulses (Panorel et al. 2013; Preis et al. 2013). Energy yields of tens of g/kWh have been achieved in this system for the removal of various pharmaceuticals and these high values have been attributed to the large contact area between the plasma and the liquid (Ajo et  al. 2015). This configuration has recently been adapted for the treatment of hospital wastewater at pilot scale (50 L) (Ajo et al. 2018) and tests have been run with promising results for both untreated sewage of a public hospital and for biologically treated wastewater effluent of a healthcare institute. Another report of a pilot-scale plasma system has been done by Gerrity et al. (2010) using the reactor developed by Aquapure Technologies Ltd. for the degradation of several pharmaceuticals in trace concentrations. The pilot unit contains a plasma reactor, based on a pulsed corona above water, and an ozone contactor which uses the ozone-rich gas from the plasma reactor. The tests have been done on tertiary-­treated wastewater and spiked surface water with contaminants concentrations of tens to hundreds of ng/L. Rapid degradation of the target compounds has been demonstrated and the authors concluded that plasma treatment may be a possible alternative to more common advanced oxidation processes, since the energy requirements for pollutants degradation are comparable and no additional feed chemicals are needed. To increase the performance of non-thermal plasma processes for the degradation of harmful water pollutants, significant research efforts have been focused on the optimisation of the electrical discharge, including reactor design, aimed at improving the energy yield. Numerous studies have been dedicated to the understanding of the process chemistry through investigation of the plasma-generated reactive species, as well as the identification and quantification of the intermediate degradation products. An issue that requires more attention is the characterisation of the plasma treated water from both chemical and (eco)toxicological points of view. There are only few studies dealing with this complex characterisation and more efforts are needed in order to correctly evaluate the plasma treatment performances. Nevertheless, it is worth mentioning that progress has been made in developing new combined processes such as plasma-ozonation or plasma-catalysis and performing tests on real wastewater effluents (like those from hospitals) at pilot-scale. In this context, non-­ thermal plasma appears to be a new potential candidate for water treatment used for emerging contaminants removal.

1.8  Conclusion The remediation of contaminated water in general, and that of emerging substances in particular, is not only a source of concern for our societies, but also a major subject of debate, at both industrial and political levels for all water stakeholders, and of research for scientists. Indeed, emerging substances study and their treatment have become a relevant research topic for scientists focused on water engineering

1  Remediation of Emerging Contaminants

73

issues. However, the challenge is not simple, as it is difficult to remove trace contaminants from complex mixtures of substances in a way that is chemically effective, technologically simple, economically viable, and environmentally friendly. In this chapter, we have described advanced treatment methods that have been proposed for this purpose, including wastewater disinfection, adsorption onto conventional or non-conventional materials, constructed wetlands, membrane bioreactors and other biological-based strategies, membrane filtration, and advanced oxidation processes such as electrochemical technologies and catalytic ozonation. Of the various disinfection methods, chlorination and ozonation appear to give better results in terms of removing both pathogens and emerging contaminants. However, further research is needed to determine the possible toxic characteristics of disinfection by-products compared to their parent compounds. The approach consisting in the use of ozonation and adsorption on activated carbon has been used for about 10 years in countries such as Switzerland and Germany, due to its efficiency, simplicity and technical feasibility at industrial scale, and also for economic reasons. Another feature of these technologies is the fact that they can be easily integrated into existing treatment facilities. One disadvantage is the rapid saturation of the carbon filters, which must then be regenerated. This step is difficult and expensive. Extensive research on non-conventional adsorbents highlights the growing interest of scientists in developing systems that are increasingly effective in removing mixtures of trace pollutants, simple to implement from a technological point of view, economically viable and environmentally friendly, with little or no impact on the environment. Materials such as cyclodextrin polymers, metal-organic frameworks, molecularly imprinted polymers, chitosan-based materials and nanocelluloses have great potential in environmental applications. However, they are still at the laboratory study stage. Further research is needed to determine the means of integration of these adsorbents into treatment plants. Biological technologies used for the degradation of new contaminants and the reduction of their negative impact on the environment are also a field of research in full development and significant advances are expected in the future. Biological approaches include constructed wetlands, biomembrane reactors, strategies based on the use of algae, fungi and bacteria, and enzymatic degradation. Membrane filtration is already used as a tertiary treatment capable of removing a wide range of pollutants and pathogens detected in water, but this treatment has not become widespread due in part to its high cost but also to problems of membrane clogging. Finally, advanced oxidation processes also represent one of the most promising strategies because of their efficiency and simplicity (they can also be integrated into plants as primary, secondary and tertiary methods). Significant advances are expected in the next few years, although here again, investment, operation and maintenance costs must be taken into account. Industrialists will now have to be convinced to use these technologies in their municipal wastewater treatment plants. What are the prospects for the management of emerging substances? Firstly, increased political measures to tackle pollution at source would help mitigate the impact of pollution. Pollution reduction requires the application of good practices

74

N. Morin-Crini et al.

for water users and industrialists: reduction of industrial discharges, avoid mixing domestic (containing antibiotics) and industrial water (containing metals; recent studies have shown a relationship between the presence of metals and the phenomenon of bacterial resistance), reduction of pesticides and other plant protection products, regulation of soil fertilization, more control or prohibition of hazardous substances, etc. All of this obviously has a cost. Secondly, the implementation of complementary water treatment systems (known as tertiary decontamination) in wastewater treatment plants would make it possible to move towards zero pollution discharge and thus increase the availability of water resources. In addition, the microbiological quality of treated wastewater is a parameter that is increasingly being taken into account. These effluents represent, in some cases, a significant source of contamination by pathogens and/or antimicrobial resistant bacteria for the receiving environment. However, in many countries, most wastewater treatment plants do not apply any disinfection step before the discharge of their water. Finally, wastewater is still an undervalued resource of water, energy, nutrients and other valuable by-products (metals). Recycling, reuse and recovery of what is normally considered waste could alleviate water stress and bring many social, economic and environmental benefits. In this context, advanced oxidation processes combined with biodegradation, adsorption and membrane filtration are promising, effective and environmentally friendly treatments to remove new pollutants and pathogens from wastewater and can be part of wastewater reuse policies, even if their high cost still prevents their widespread use. Acknowledgements  Nadia Morin-Crini and Grégorio Crini (Besançon, France) thanks the FEDER (Fonds Européen de Développment Régional) for its financial support (NIRHOFEX Program: “Innovative materials for wastewater treatment”) and the Université de Franche-Comté (France) for the research grant awarded to Guest Professor C.  Bradu. Ana Rita Lado Ribeiro (Porto, Portugal) acknowledges the support from Projects: PTDC/QUI-QAN/30521/2017  – POCI-01-0145-FEDER-030521 – funded by FEDER funds through COMPETE2020 – Programa Operacional Competitividade e Internacionalização (POCI) and by national funds (PIDDAC) through FCT/MCTES and Base Funding – UIDB/50020/2020 of the Associate Laboratory LSRE-­ LCM – funded by national funds through FCT/MCTES (PIDDAC). Monica Magureanu (Magurele, Romania) acknowledges financial support from UEFISCDI  – project 18 BM/2019. Francesca Mapelli (Milan, Italy) acknowledge the Cariplo Foundation for its financial support (GA n° 2018-0995-WARFARE “Novel wastewater disinfection treatments to mitigate the spread of antibiotic resistance in agriculture”).

References Abargues MR, Ferrer J, Bouzas A, Seco A (2018) Fate of endocrine disruptor compounds in an anaerobic membrane bioreactor (AnMBR) coupled to an activated sludge reactor. Environ Sci Water Res Technol 4:226–233. https://doi.org/10.1039/c7ew00382j Abouzeid RE, Khiari R, El-Wakil N, Dufresne A (2018) Current state and new trends in the use of cellulose nanomaterials for wastewater treatment. Biomacromolecules 20:573–597. https://doi. org/10.1021/acs.biomac.8b00839

1  Remediation of Emerging Contaminants

75

Abujaber F, Zougagh M, Jodeh S, Ríos A, Javier F, Bernardo G, Martín-Doimeadios RCR (2018) Magnetic cellulose nanoparticles coated with ionic liquid as a new material for the simple and fast monitoring of emerging pollutants in waters by magnetic solid phase extraction. Microchem J 137:490–495. https://doi.org/10.1016/j.microc.2017.12.007 Acayanka E, Tarkwa JB, Laminsi S (2019) Evaluation of energy balance in a batch and circulating non-thermal plasma reactors during organic pollutant oxidation in aqueous solution. Plasma Chem Plasma P 39:75–87. https://doi.org/10.1007/s11090-­018-­9946-­7 Acero JL, Benitez FJ, Real J, Roldan G (2010) Kinetics of aqueous chlorination of pharmaceuticals and their elimination from water matrices. Water Res 44:4158–4170. https://doi.org/10.1016/j. watres.2010.05.012 Acero JL, Benitez FJ, Real FJ, Roldan G, Rodriguez E (2013) Chlorination and bromination kinetics of emerging contaminants in aqueous systems. Chem Eng J 219:43–50. https://doi. org/10.1016/j.cej.2012.12.067 Adams C, Wang Y, Loftin K, Meyer M (2002) Removal of antibiotics from surface and distilled water in conventional water treatment processes. J Environ Eng 128:253–260 Adriano WS, Veredas V, Santana CC, Gonçalves LRB (2005) Adsorption of amoxicillin on chitosan beads: kinetics, equilibrium and validation of finite bath models. Biochem Eng J 27:132–137. https://doi.org/10.1016/j.bej.2005.08.010 Afzal MZ, Sun X, Liu J, Song C, Wang S, Javed A (2018) Enhancement of ciprofloxacin sorption on chitosan/biochar hydrogel beads. Sci Total Environ 639:560–569 Ahmaruzzaman MD (2008) Adsorption of phenolic compounds on low-cost adsorbents: a review. Adv Colloid Int Sci 143:48–67. https://doi.org/10.1016/j.cis.2008.07.002 Ahmed MB, Zhou JL, Ngo HH, Guo W (2015) Adsorptive removal of antibiotics from water and wastewater: Progress and challenges. Sci Total Environ 532:112–126. https://doi.org/10.1016/j. scitotenv.2015.05.130 Ahmed I, Iqbal DK (2017) Enzyme-based biodegradation of hazardous pollutants – an overview. J Exp Biol Agricultural Sci 5:402–411. https://doi.org/10.18006/2017.5(4).402.411 Ahmed M, Zhou J, Ngo H, Guo W, Thomaidis N, Xu J (2017) Progress in the biological and chemical treatment technologies for emerging contaminants removal from wastewater: a critical review. J Hazard Mater 323:274–298. https://doi.org/10.1016/j.jhazmat.2016.04.045 Ahn Y, Oh H, Yoon Y, Park WK, Yang WS, Kang JW (2017) Effect of graphene oxidation degree on the catalytic activity of graphene for ozone catalysis. J Environ Chem Eng 5:3882–3894. https://doi.org/10.1016/j.jece.2017.07.038 Ajo P, Kornev I, Preis S (2015) Pulsed corona discharge in water treatment: the effect of hydrodynamic conditions on oxidation energy efficiency. Ind Eng Chem Res 54:7452–7458. https:// doi.org/10.1021/acs.iecr.5b01915 Ajo P, Preis S, Vornamo T, Mänttäri M, Kallioinen M, Louhi-Kultanen M (2018) Hospital wastewater treatment with pilot-scale pulsed corona discharge for removal of pharmaceutical residues. J Environ Chem Eng 6:1569–1577. https://doi.org/10.1016/j.jece.2018.02.007 Aksu Z (2005) Application of biosorption for the removal of organic pollutants: a review. Process Biochem 40:997–1026. https://doi.org/10.1016/j.procbio.2004.04.008 Alexander C, Andersson HS, Andersson LI, Ansell RJ, Kirsch N, Nicholls IA, O’Mahony J, Whitcombe MJ (2006) Molecular imprinting science and technology: a survey of the literature for the years up to and including 2003. J Mol Recognit 19:106–180 Ali I (2012) New generation adsorbents for water treatment. Chem Rev 112:5073–5091. https:// doi.org/10.1021/cr300133d Ali MEM, Abd El-Aty AM, Badawy MI, Ali RK (2018) Removal of pharmaceutical pollutants from synthetic wastewater using chemically modified biomass of green alga Scenedesmus obliquus. Ecotoxicol Environ Saf 151:144–152. https://doi.org/10.1016/j.ecoenv.2018.01.012 Almeida CAP, Debacher NA, Downs AJ, Cottet L, Mello CAD (2009) Removal of methylene blue from colored effluents by adsorption on montmorillonite clay. J Colloid Interf Sc 332:46–53. https://doi.org/10.1016/j.jcis.2008.12.012

76

N. Morin-Crini et al.

Alonso JSJ, Kori NE, Melian-Martel N, Del Río-Gamero B (2018) Removal of ciprofloxacin from seawater by reverse osmosis. J Environ Manag 217:337–345 Alsbaiee A, Smith BJ, Xiao L, Ling Y, Helbling DE, Dichtel WR (2016) Rapid removal of organic micropollutants from water by a porous β-cyclodextrin polymer. Nature 529:190–194. https:// doi.org/10.1038/nature16185 Alsudir S, Iqbal Z, Lai EPC (2012) Competitive CE-UV binding tests for selective recognition of bisphenol A by molecularly imprinted polymer particles. Electrophoresis 33:1255–1262 Alvarino T, Torregrosa N, Omil F, Lema J, Suarez S (2017) Assessing the feasibility of two hybrid MBR systems using PAC for removing macro and micropollutants. J Environ Manag 203:831–837 Alvaro M, Carbonell E, Ferrer B, Llabrés IXFX, Garcia H (2007) Semiconductor behavior of a metal-organic framework (MOF). Chem Eur J 13:5106–5112 Altmann D, Schaar H, Bartel C, Schorkopf DLP, Miller I, Kreuzinger N, Möstl E, Grillitsch B (2012) Impact of ozonation on ecotoxicity and endocrine activity of tertiary treated wastewater effluent. Water Res 46:3693–3702. https://doi.org/10.1016/j.watres.2012.04.017 Altmann J, Ruhl AS, Zietzschmann F, Jekel M (2014) Direct comparison of ozonation and adsorption onto powdered activated carbon for micropollutant removal in advanced wastewater treatment. Water Res 55:185–193. https://doi.org/10.1016/j.watres.2014.02.025 Ameta R, Chohadia AK, Jain A, Punjabi PB (2018) Chapter 3: Fenton and photo-Fenton processes. In: Advanced oxidation processes for waste water treatment, pp 49–87. https://doi.org/10.1016/ B978-­0-­12-­810499-­6.00003-­6 Anastopoulos I, Pashalidis I, Orfanos AG, Manariotis ID, Tatarchuk T, Sellaoui L, Bonilla-­ Petriciolet A, Mittal A, Núñez-Delgado A (2020) Removal of caffeine, nicotine and amoxicillin from (waste)waters by various adsorbents. A Rev J Environ Manag 261:110236. https://doi. org/10.1016/j.jenvman.2020.110236 Andreozzi R, Raffaele M, Nicklas R (2003) Pharmaceuticals in STP effluents and their solar photodegradation in aquatic environment. Chemosphere 50:1319–1330. https://doi.org/10.1016/ S0045-­6535(02)00769-5 Aoki N, Kinoshita K, Mikuni K, Nakanishi K, Hattori K (2007) Adsorption of 4-nonylphenol ethoxylates onto insoluble chitosan beads bearing cyclodextrin moieties. J Inclusion Phenomena Macrocyclic Chem 57:237–241. https://doi.org/10.1007/s10847-­006-­9190-­2 Aritoshi K, Fujiwara M, Ishida M (2002) Production and removal mechanisms of discharge NOx treatment in N2/O2 gas mixture. Jpn J Appl Phys 41:3936–3942. https://doi.org/10.1143/ JJAP.41.3936 Asano T, Burton FL, Leverenz HL, Tsuchihashi R, Tchobanoglous G (2007) Water reuse: issues, technologies, and applications. McGraw Hill, Metcalf Eddy Inc. AECO, New  York. ISBN: 978-0-007-145927-3 Asif MB, Hai FI, Singh L, Price WE, Nghiem LD (2017) Degradation of pharmaceuticals and personal care products by white-rot-fungi – a critical review. Cur Pollut Rep 3:88–103. https:// doi.org/10.1007/s40726-­017-­0049-­5 Badia-Fabregat M, Lucas D, Gros M, Rodríguez-Mozaz S, Barceló D, Caminal G, Vicent T (2015) Identification of some factors affecting pharmaceutical active compounds (PhACs) removal in real wastewater. Case study of fungal treatment of reverse osmosis concentrate. J Hazard Mater 283:663–671 Baeza C, Knappe DR (2011) Transformation kinetics of biochemically active compounds in low-­ pressure UV photolysis and UV/H2O2 advanced oxidation processes. Water Res 45:4531–4543. https://doi.org/10.1016/j.watres.2011.05.039 Bai X, Acharya K (2017) Algae-mediated removal of selected pharmaceutical and personal care products (PPCPs) from Lake Mead water. Sci Total Environ 581–582:734–740. https://doi. org/10.1016/j.scitotenv.2016.12.192 Bai Z, Yang Q, Wang J (2017) Catalytic ozonation of dimethyl phthalate using Fe3O4/multi-­ wall carbon nanotubes. Environ Technol 38:2048–2057. https://doi.org/10.1080/0959333 0.2016.1245360

1  Remediation of Emerging Contaminants

77

Baille WE, Huang WQ, Nichifor M, Zhu XX (2000) Functionalized β-cyclodextrin polymers for the sorption of bile salts. J Macromol Sci A 37:677–690. https://doi.org/10.1081/ma-­100101117 Bao Q, Hui KS, Duh JG (2016) Promoting catalytic ozonation of phenol over graphene through nitrogenation and Co3O4 compositing. J Environ Sci 50:38–48. https://doi.org/10.1016/j. jes.2016.03.029 Barbosa MO, Moreira NFF, Ribeiro AR, Pereira MFR, Silva AMT (2016) Occurrence and removal of organic micropollutants: an overview of the watch list of EU Decision 2015/495. Water Res 94:257–279. https://doi.org/10.1016/j.watres.2016.02.047 Basile T, Petrella A, Petrella M, Boghetich G, Petruzzelli V, Colasuonno S, Petruzzelli D (2011) Review of endocrine-disrupting-compound removal technologies in water and wastewater treatment plants: an EU perspective. Ind Eng Chem Res 50:8389–8401. https://doi.org/10.1021/ ie101919v Batt AL, Aga DS (2005) Simultaneous analysis of multiple classes of antibiotics by ion trap LC/ MS/MS for assessing surface water and ground water contamination. Anal Chem 77:2940–2947 Beall GW (2003) The use of organo-clays in water treatment. Appl Clay Sci 24:11–20. https://doi. org/10.1016/j.clay.2003.07.006 Bedia J, Muelas-Ramos V, Peñas-Garzón M, Gómez-Avilés A, Rodríguez JJ, Belver C (2019) A review on the synthesis and characterization of metal organic frameworks for photocatalytic water purification. Catalysts 9:52. https://doi.org/10.3390/catal9010052 Bedner M, MacCrehan WA (2006) Transformation of acetaminophen by chlorination produces the toxicants 1,4-benzoquinone and n-acetyl-pbenzoquinone imine. Environ Sci Technol 40:516–522. https://doi.org/10.1021/es0509073 Bedoya-Ríos DF, Lara-Borrero JA (2018) Occurrence of endocrine disruptor chemicals in the urban cycle of Colombia. In: Ahmed RG (ed) Endocrine disruptors. IntechOpen. https://doi. org/10.5772/intechopen.78325 Bekçi Z, Seki Y, Yurdakoç MK (2006) Equilibrium studies for trimethoprim adsorption on montmorillonite KSF. J Hazard Mater 133:233–242. https://doi.org/10.1016/j.jhazmat.2005.10.029 Bekçi Z, Seki Y, Yurdakoç MK (2007) A study of equilibrium and FTIR, SEM/EDS analysis of trimethoprim adsorption onto K10. J Mol Struct 827:67–74. https://doi.org/10.1016/j. molstruc.2006.04.054 BelBruno JJ (2019) Molecularly imprinted polymers. Cehm Rev 119:94–119. https://doi. org/10.1021/acs.chemrev.8b00171 Belgiorno V, Rizzo L, Fatta D, Della Rocca C, Lofrano G, Nikolaou A, Naddeo V, Meric S (2007) Review on endocrine disrupting-emerging compounds in urban wastewater: occurrence and removal by photocatalysis and ultrasonic irradiation for wastewater reuse. Des 215:166–176. https://doi.org/10.1016/j.desal.2006.10.035 Benner J, Salhi E, Ternes T, von Gunten U (2008) Ozonation of reverse osmosis concentrate: kinetics and efficiency of beta blocker oxidation. Water Res 42:3003–3012. https://doi. org/10.1016/j.watres.2008.04.002 Bennett JL, Mackie AL, Park Y, Gagnon GA (2018) Advanced oxidation processes for treatment of 17[beta]-estradiol and its metabolites in aquaculture wastewater. Aquacultural Eng 83:40 Benotti M, Trenholm R, Vanderford B, Holady J, Stanford B, Snyder S (2009a) Pharmaceuticals and endocrine disrupting compounds in U.S drinking water. Environ Sci Technol 43:597–603. https://doi.org/10.1021/es801845a Benotti MJ, Stanford BD, Wert EC, Snyder SA (2009b) Evaluation of a photocatalytic reactor membrane pilot system for the removal of pharmaceuticals and endocrine disrupting compounds from water. Water Res 43:1513–1522. https://doi.org/10.1016/j.watres.2008.12.049 Besha AT, Gebreyohannes AY, Tufa RA, Bekele DN, Curcio E, Giorno L (2017) Removal of emerging micropollutants by activated sludge process and membrane bioreactors and the effects of micropollutants on membrane fouling: a review. J Environ Chem Eng 5:2395–2414 Bhattacharyya KG, Gupta SS (2008) Adsorption of a few heavy metals on natural and modified kaolinite and montmorillonite: a review. Adv Colloid Int Sci 140:114–131. https://doi. org/10.1016/j.cis.2007.12.008

78

N. Morin-Crini et al.

Bhattarai B, Manickavachagam M, Suri R (2012) Development of novel cyclodextrin adsorbents for the removal of emerging contaminants from water. In: Abstracts of papers, 244th ACS national meeting and exposition, Philadelphia, 19–23 August 2012, ENVR-5 Bilal M, Adeel M, Rasheed T, Zhao Y, Iqbal HMN (2019a) Emerging contaminants of high concern and their enzyme-assisted biodegradation – a review. Environ Int 124:336–353. https:// doi.org/10.1016/j.envint.2019.01.011 Bilal M, Rasheed T, Nabeel F, Iqbal HMN, Zhao YP (2019b) Hazardous contaminants in the environment and their laccase-assisted degradation  – a review. J Environ Manag 234:253–264. https://doi.org/10.1016/j.jenvman.2019.01.001 Bilea F, Bradu C, Mandache NB, Magureanu M (2019) Characterization of the chemical activity of a pulsed corona discharge above water. Chemosphere 236:124302. https://doi.org/10.1016/j. chemosphere.2019.07.033 Bodzek M (2019) Membrane separation techniques - removal of inorganic and organic admixtures and impurities from water environment - review. Arch Environ Protection 45:4–19. https://doi. org/10.24425/aep.2019.130237 Boehm D, Curtin J, Cullen PJ, Bourke P (2018) Hydrogen peroxide and beyond-the potential of high-voltage plasma activated liquids against cancerous cells. Anti Cancer Agents Med Chem 18:815–823. https://doi.org/10.2174/1871520617666170801110517 Bolong N, Ismail AF, Salim MR, Matsura T (2009) A review of the effects of emerging contaminants in wastewater and options for their removal. Des 239:229–246. https://doi.org/10.2016/j. desal.2008.03.20 Borea L, Ensano BMB, Hasan SW, Balakrishnan M, Belgiorno V, de Lun DG, Ballesteros FC Jr, Naddeo V (2019) Are pharmaceuticals removal and membrane fouling in electromembrane bioreactor affected by current density? Sci Total Environ 692:732–740. https://doi.org/10.1016/j. scitotenv.2019.07.149 Botero-Coy AM, Martínez-Pachón D, Boix C, Rincón RJ, Castillo N, Arias-Marín NL, Manrique-­ Losada L, Torres-Palma R, Moncayo-Lasso A, Hernández F (2018) An investigation into the occurrence and removal of pharmaceuticals in wastewater Colombian. Sci Total Environ 642:842–853. https://doi.org/10.1016/j.scitotenv.2018.06.088 Bourgin M, Beck B, Boehler M, Borowska E, Fleiner J, Salhi E, Teichler R, von Gunten U, Siegrist H, McArdell CS (2018) Evaluation of a full-scale wastewater treatment plant upgraded with ozonation and biological post-treatments: abatement of micropollutants, formation of transformation products and oxidation by-products. Water Res 129:486–498. https://doi.org/10.1016/j. watres.2017.10.036 Bradu C, Magureanu M, Parvulescu VI (2017) Degradation of the chlorophenoxyacetic herbicide 2,4-D byplasma-ozonation system. J Hazard Mater 336:52–56. https://doi.org/10.1016/j. jhazmat.2017.04.050 Bradu C, Kutasi K, Magureanu M, Puač N, Živković S (2020) Reactive nitrogen species in plasma-­ activated water: generation, chemistry and application in agriculture. J Phys D: Appl Phys 53:223001. https://doi.org/10.1088/1361-­6463/ab795a Braga FG, Pinto S, Antunes MCG (2011) Comparative study of 17β-estradiol removal from aqueous solutions using pine bark and almond shell as adsorbents. Microchim Acta 173:111–117. https://doi.org/10.1007/s00604-­010-­0531-­x Brisset JL, Pawlat J (2016) Chemical effects of air plasma species on aqueous solutes in direct and delayed exposure modes: discharge, post-discharge and plasma activated water. Plasma Chem Plasma Process 36:355–381. https://doi.org/10.1007/s11090-­015-­9653-­6 Bruggeman P, Leys C (2009) Non-thermal plasmas in and in contact with liquids. J Phys D Appl Phys 42:053001. https://doi.org/10.1088/0022-­3727/42/5/053001 Bruggeman P, Schram DC (2010) On OH production in water containing atmospheric pressure plasmas. Plasma Sources Sci Technol 19:45025. https://doi.org/10.1088/0963-­0252/19/4/045025 Bruggeman PJ, Kushner MJ, Locke BR, Gardeniers JGE, Graham WG, Graves DB, Hofman-Caris RCHM, Maric D, Reid JP, Ceriani E, Fernandez Rivas D, Foster JE, Garrick SC, Gorbanev Y, Hamaguchi S, Iza F, Jablonowski H, Klimova E, Kolb J, Krcma F, Lukes P, Machala Z,

1  Remediation of Emerging Contaminants

79

Marinov I, Mariotti D, Mededovic Thagard S, Minakata D, Neyts EC, Pawlat J, ZLj P, Pflieger R, Reuter S, Schram DC, Schröter S, Shiraiwa M, Tarabová B, Tsai PA, JRR V, von Woedtke T, Wilson KR, Yasui K, Zvereva G (2016) Plasma-liquid interactions: a review and roadmap. Plasma Sources Sci Technol 25:053002. https://doi.org/10.1088/0963-­0252/25/5/053002 Buffle M, Schumacher J, Salhi E, Jekel M, von Gunten U (2006) Measurement of the initial phase of ozone decomposition in water and wastewater by means of continuous quench flow system: application of disinfection and pharmaceutical oxidation. Water Res 40:1884–1894. https://doi. org/10.1016/j.watres.2006.02.026 Burlica R, Kirkpatrick MJ, Finney WC, Clark RJ, Locke BR (2004) Organic dye removal from aqueous solution by glidarc discharges. J Electrost 62:309–321. https://doi.org/10.1016/j. elstat.2004.05.007 Bulloch DN, Lavado R, Forsgren KL, Beni S, Schlenk D, Larive CK (2012) Analytical and biological characterization of halogenated gemfibrozil produced through chlorination of wastewater. Environ Sci Technol 46:5583–5589. https://doi.org/10.1021/es3006173 Cadorin BM, Tralli D, Ceriani E, de Brito Benetoli LO, Marotta E, Ceretta C, Debacher NA, Paradisi C (2015) Treatment of methyl orange by nitrogen non-thermal plasma in a corona reactor: the role of reactive nitrogen species. J Hazard Mater 300:754–764. https://doi. org/10.1016/j.jhazmat.2015.08.009 Canonica S, Meunier L, von Gunten U (2008) Phototransformation of selected pharmaceuticals during UV treatment of drinking water. Water Res 42:121–128 Cantarella M, Carroccio S, Dattilo S, Avolio R, Castaldo R, Puglisi C, Privitera V (2019) Molecularly imprinted polymer for selective adsorption of diclofenac from contaminated water. Chem Eng J 367:180–188. https://doi.org/10.1016/j.cej.2019.02.146 Cantarella M, Impellizzeri G, Privitera V (2017) Functional nanomaterials for water purification. Riv Nuovo Cimento 40:595–632 Castiglioni S, Bagnati R, Fanelli R, Pomati F, Calamari D, Zuccato E (2006) Removal of pharmaceuticals in sewage treatment plants in Italy. Environ Sci Technol 40:357–363. https://doi. org/10.1021/es050991m Charles J, Crini G, Morin-Crini N, Badot PM, Trunfio G, Sancey B, de Carvalho M, Bradu C, Avramescu S, Winterton P, Gavoille S, Torri G (2014) Advanced oxidation (UV-ozone) and cyclodextrin sorption: effects of individual and combined action on the chemical abatement of organic pollutants in industrial effluents. J Taiwan Institute Chem Eng 45:603–608. https://doi. org/10.1016/j.jtice.2013.06.023 Chen PJ, Linden KG, Hinton DE, Kashiwada S, Rosenfeldt EJ, Kullman SW (2006) Biological assessment of bisphenol A degradation in water following direct photolysis and UV advanced oxidation. Chemosphere 65:1094–1102. https://doi.org/10.1016/j.chemosphere.2006.04.048 Chen HW, Liang CH, Wu ZM, Lin TF, Chiang PC, Wang GS (2013) Occurrence and assessment of treatment efficiency of nonyphenol, octylphenol and bisphenol A on drinking water in Taiwan. Sci Total Environ 449:20–28. https://doi.org/10.1016/j.scitotenv.2013.01.038 Chen W, Xue M, Xue F, Mu X, Xu Z, Meng Z, Zhu G, Shea KJ (2015) Molecularly imprinted hollow spheres for the solid phase extraction of estrogens. Talanta 140:68–72 Chen L, Wang X, Lu W, Wu X, Li J (2016a) Molecular imprinting: perspectives and applications. Chem Soc Rev 45:2137 Chen J, Wei XD, Liu YS, Ying GG, Liu SS, He LY, Su HC, Hu LX, Chen FR, Yang YQ (2016b) Removal of antibiotics and antibiotic resistance genes from domestic sewage by constructed wetlands: optimization of wetland substrates and hydraulic loading. Sci Total Environ 565:240–248. https://doi.org/10.1016/j.scitotenv.2016.04.176 Chung YC, Chen CY (2009) Competitive adsorption of a phthalate esters mixture by chitosan bead and α-cyclodextrin-linked chitosan bead. Environ Technol 30:1343–1350. https://doi. org/10.1080/09593330902858914 Coelho AD, Sans C, Agüera A, Gómez MJ, Esplugas S, Dezotti M (2009) Effects of ozone pre-­ treatment on diclofenac: intermediates, biodegradability and toxicity assessment. Sci Total Environ 407:3572–3578. https://doi.org/10.1016/j.scitotenv.2009.01.013

80

N. Morin-Crini et al.

Collivignarelli MC, Abbà A, Benigna I, Sorlini S, Torretta V (2018) Overview of the main disinfectin processes for wastewater and drinking water treatment plants. Sustainability 10:86. https://doi.org/10.3390/su10010086 Comes J, Costa R, Quinta-Ferreira RM, Martins RC (2017) Application of ozonation for pharmaceuticals and personal care products removal from water. Sci Total Environ 586:265–283. https://doi.org/10.1016/j.scitotenv.2017.01.216 Couto OM, Matos I, da Fonseca IM, Arroyo PA, da Silva EA, de Barros MASD (2015) Effect of solution pH and influence of water hardness on caffeine adsorption onto activated carbons. Can J Chem Eng 93:68–77. https://doi.org/10.1002/cjce.22104 Crini (2005) Recent developments in polysaccharide-based materials used as adsorbents in wastewater treatment. Prog Polym Sci 30:38–70. https://doi.org/10.1016/j.progpolymsci.2004.11.002 Crini G (2006) Non-conventional low-cost adsorbents for dye removal. Bioresour Technol 97:1061–1085. https://doi.org/10.1016/j.biortech.2005.05.001 Crini G (2010) Chapter 2: Wastewater treatment by sorption. In: Sorption processes and pollution. PUFC, Besançon, pp 39–78 Crini G (2014) Review: a history of cyclodextrins. Chem Rev 114:10940–10975. https://doi. org/10.1021/cr500081p Crini G (2015) Non-conventional adsorbents for dye removal. In: Sharma SK (ed) Green chemistry for dyes removal from wastewater. Scrivener Publishing LLC, pp 359–407 Crini G, Badot PM (2007) Traitement et épuration des eaux industrielles polluées (in French). Presses Universitaires de Franche-Comté, France, Besançon. ISBN 978-2-84867-197-0 Crini G, Badot PM (2010) Sorption processes and pollution. Presses Universitaires de Franche-­ Comté, France, Besançon. ISBN 978-2-84867-304-2 Crini G, Morcellet M (2002) Synthesis and applications of adsorbents containing cyclodextrins. J Sep Sci 25:789–813. https://doi.org/10.1002/1615-­9314(20020901)25:133.0.co;2-­j Crini G, Lichtfouse E (2018a) Chapter 1: Wastewater treatment: an overview. In: Crini G, Lichtfouse E (eds) Green adsorbents for pollutant removal. Springer, Cham, pp 1–21. https:// doi.org/10.1007/978-­3-­319-­92111-­2_1 Crini G, Lichtfouse E (eds) (2018b) Chitin and chitosan: applications in food, agriculture, pharmacy, medicine and wastewater treatment. Sustainable Agriculture Review 36. Springer. https://doi.org/10.1007/978-­3-­030-­16581-­9 Crini G, Lekchiri Y, Morcellet M (1995) Separation of structural isomers using cyclodextrin-­ polymers coated on silica beads. Chromatographia 40:296–302. https://doi.org/10.1007/ bf02290360 Crini G, Janus L, Morcellet M, Torri G, Naggi A, Bertini S, Vecchi C (1998) Macroporous polyamines containing cyclodextrin: synthesis, characterization, and sorption properties. J Appl Polym Sci 69:1419–1427. https://doi.org/10.1002/(sici)1097-­4628(19980815)69:73.0.co;2-­o Crini G, Euvrard E, Druart C, Morin-Crini N, Lagarrigue C, Gavoille S (2016) Alkylphénols et rejets de la filière traitement de surface (in French). GalvanoOrgano 845:53–55 Crini G, Fourmentin S, Fenyvesi É, Torri G, Fourmentin M, Morin-Crini N (2018) Cyclodextrins, from molecules to applications. Environ Chem Lett 16:1361–1375. https://doi.org/10.1007/ s10311-­018-­0763-­2 Crini G, Lichtfouse E, Wilson LD, Morin-Crini N (2019) Conventional and non-conventional adsorbents for wastewater treatment. Environ Chem Lett 17:145–155. https://doi.org/10.1007/ s10311-­018-­0785-­9 Cruz del Álamo A, Pariente MI, Vasiliadou I, Padrino B, Puyol D, Molina R, Martínez F (2018) Removal of pharmaceutical compounds from urban wastewater by an advanced bio-oxidation process based on fungi Trametes versicolor immobilized in a continuous RBC system. Environ Sci Pollut Res 25:34884–34892. https://doi.org/10.1007/s11356-­017-­1053-­4 Cruz-Morató C, Ferrando-Climent L, Rodriguez-Mozaz S, Barceló D, Marco-Urrea E, Vicent T, Sarrà M (2013) Degradation of pharmaceuticals in non-sterile urban wastewater by Trametes

1  Remediation of Emerging Contaminants

81

versicolor in a fluidized bed bioreactor. Water Res 47:5200–5210. https://doi.org/10.1016/j. watres.2013.06.007 Cruz-Morató C, Lucas D, Llorca M, Rodriguez-Mozaz S, Gorga M, Petrovic M, Barceló D, Vicent T, Sarrà M, Marco-Urrea E (2014) Hospital wastewater treatment by fungal bioreactor: removal efficiency for pharmaceuticals and endocrine disruptor compounds. Sci Total Environ 493:365–376. https://doi.org/10.1016/j.scitotenv.2014.05.117 Cruz-Tato P, Ortiz-Quiles EO, Vega-Figueroa K, Santiago-Martoral L, Flynn M, Diaz-Vázquez LM, Nicolau E (2017) Metalized nanocellulose composites as a feasible material for membrane supports: design and applications for water treatment. Environ Sci Technol 51:4585–4595 CycloLab (2020). www.cyclolab.hu Cyclopure (2020). www.cyclopure.com Dąbrowski A, Podkościelny P, Hubicki Z, Barczak M (2005) Adsorption of phenolic compounds by activated carbon – a critical review. Chemosphere 58:1049–1070. https://doi.org/10.1016/j. chemosphere.2004.09.067 Dantas R, Contreras S, Sans C, Esplugas S (2008) Sulfamethoxazole abatement by means of ozonation. J Hazard Mater 150:790–794 de Andrade JR, Oliveira MF, da Silva MGC, Vieira MGA (2018) Adsorption of pharmaceuticals from water and wastewater using nonconventional low-cost materials: a review. Ind Eng Chem Res 57:3103–3127. https://doi.org/10.1021/acs.iecr.7b05137 de Andrade JR, Vieira MGA, Silva MGC, Wang S (2020) Oxidative degradation of pharmaceutical losartan potassium with N-doped hierarchical porous carbon and peroxymonosulfate. Chem Eng J 382:1–11. https://doi.org/10.1016/j.cej.2019.122971 de Araújo KS, Antonelli R, Gaydeczka B, Granato AC, Malpass GRP (2016) Advanced oxidation processes: a review regarding the fundamentals and applications in wastewater treatment and industrial wastewater. Ambiente e Agua Int J Appl Sci 11. https://doi-­org.inee.bib.cnrs. fr/10.4136/1980-­993X de Godos I, Muñoz R, Guieysse B (2012) Tetracycline removal during wastewater treatment in high-rate algal ponds. J Hazard Mater 229–230:446–449. https://doi.org/10.1016/j. jhazmat.2012.05.106 de Jesus GV, Cardoso VV, Joao Benoliel M, Almeida CMM (2016) Chlorination and oxidation of sulfonamides by free chlorine: identification and behaviour of reaction products by UPLC-MS/ MS. J Environ Manag 166:466–477. https://doi.org/10.1016/j.jenvman.2015.10.048 de Oliveira M, Frihling BEF, Velasques J, Magalhaes FJG, Cavalheri PS, Migliolo L (2020) Pharmaceuticals residues and xenobiotics contaminants: occurrence, analytical techniques and sustainable alternatives for wastewater treatment. Sci Total Environ 705:135568. https://doi. org/10.1016/j.scitotenv.2019.135568 de Souza FM, Dos Santos OAA, Vieira MGA (2019) Adsorption of herbicide 2,4-D from aqueous solution using organo-modified bentonite clay. Environ Sci Pollut Res 26:18329–18342. https://doi.org/1010.1007/s11356-­019-­05196-­w de Souza FM, Lazarin AM, Vieira MGA, Santos OAA (2018) Kinetic, equilibrium and thermodynamic study on atrazine adsorption in organophilic clay. Des Water Treat 123:240–252. https:// doi.org/10.5004/dwt.2018.22767 de Wilt A, Butkovskyi A, Tuantet K, Leal LH, Fernandes TV, Langenhoff A, Zeeman G (2016) Micropollutant removal in an algal treatment system fed with source separated wastewater streams. J Hazard Mater 304:84–92. https://doi.org/10.1016/j.jhazmat.2015.10.033 Deborde M, Von Gunten U (2008) Reactions of chlorine with inorganic and organic compounds during water treatment  – kinetics and mechanism: a critical review. Water Res 42:13–51. https://doi.org/10.1016/j.watres.2007.07.025 Dehghani MH, Ghadermazi M, Bhatnagar A, Sadighara P, Jahed-Khaniki G, Heibati B, McKay G (2016) Adsorptive removal of endocrine disrupting bisphenol a from aqueous solution using chitosan. J Environ Chem Eng 4:2647–2655

82

N. Morin-Crini et al.

Dhaka S, Kumar R, Deep A, Kurade MB, Ji SW, Jeon BH (2019) Metal-organic frameworks (MOFs) for the removal of emerging contaminants from aquatic environments. Coordination Chem Rev 380:330–352. https://doi.org/10.1016/j.ccr.2018.10.003 Dias EM, Petit C (2015) Towards the use of metal-organic frameworks for water reuse: a review of the recent advances in the field of organic pollutants removal and degradation and the next steps in the field. J Mater Chem A 3:22484–22506. https://doi.org/10.1039/C5TA05440K Dobrin D, Bradu C, Magureanu M, Mandache NB, Parvulescu VI (2013) Degradation of diclofenac in water using a pulsed corona discharge. Chem Eng J 234:389e396. https://doi.org/10.1016/j. cej.2013.08.114 Dodd MC, Kohler HPE, von Gunten U (2009) Oxidation of antibacterial compounds by ozone and hydroxyl radical: elimination of biological activity during aqueous ozonation processes. Environ Sci Technol 43:2498–2504. https://doi.org/10.1021/es8025424 Doll TE, Frimmel FH (2003) Fate of pharmaceuticals  – photodegradation by simulated solar UV-light. Chemosphere 52:1757–1769. https://doi.org/10.1016/S0045-­6535(03)00446-­6 Dojčinović BP, Roglić GM, Obradović BM, Kuraica MM, Kostić MM, Nešić J, Manojlović DD (2011) Decolorization of reactive textile dyes using water falling film dielectric barrier discharge. J Hazard Mater 192:763–771 Domańska M, Boral A, Hamal K, Kuśnierz M, Łomotowski J, Płaza-Ożóg P (2007) Efficiency of municipal wastewater treatment with membrane bioreactor. J Water Land Dev. https://doi. org/10.2478/jwld-­2019-­0026 Dordio AV, Miranda S, Prates Ramalho JP, Palace Carvalho AJ (2016) Mechanisms of removal of three widespread pharmaceuticals by two clay materials. J Hazard Mater 323:575–583. https:// doi.org/10.1016/j.jhazmat.2016.05.091 El Shaer M, Eldaly M, Heikal G, Sharaf Y, Diab H, Mobasher M, Rousseau A (2020) Antibiotics degradation and bacteria inactivation in water by cold atmospheric plasma discharges above and below water surface. Plasma Chem Plasma Process. https://doi.org/10.1007/ s11090-­020-­10076-­0 EMA (European Medicines Agency) (2017) Background review for cyclodextrins used as excipient. EMA/CHMP/333892/2013. https://www.ema.europa.eu/en/documents/scientific-­guideline/ questions-­answers-­cyclodextrins-­used-­excipients-­medicinal-­products-­human-­use_en.pdf Escapa C, Coimbra RN, Paniagua S, García AI, Otero M (2017a) Paracetamol and salicylic acid removal from contaminated water by microalgae. J Environ Manag 203:799–806. https://doi. org/10.1016/j.jenvman.2016.06.051 Escapa C, Coimbra RN, Paniagua S, García AI, Otero M (2017b) Comparison of the culture and harvesting of Chlorella vulgaris and Tetradesmus obliquus for the removal of pharmaceuticals from water. J Appl Phycol 29:1179–1193. https://doi.org/10.1007/s10811-­016-­1010-­5 Fan X, Restivo J, Órfão JJM, Pereira MFR, Lapkin AA (2014) The role of multiwalled carbon nanotubes (MWCNTs) in the catalytic ozonation of atrazine. Chem Eng J 241:66–76. https:// doi.org/10.1016/j.cej.2013.12.023 Fenyvesi É, Zsadon B, Szejtli J, Tüdős F (1979) Preparation of cyclodextrin polymers in bead form and control of their characteristics by the parameters of preparation. Ann Univ Sci Budap Rolando Eotvos Nominatae Sect Chim 15:13–22 Fenyvesi É, Ujházy A, Szejtli J, Putter S, Gan TG (1996) Controlled release of drugs from CD polymers substituted with ionic groups. J Incl Phenom Mol Recogn Chem 25:185–189. https:// doi.org/10.1007/bf01041566 Fenyvesi É, Puskás I, Szente L (2019) Applications of steroid drugs entrapped in cyclodextrins. Environ Chem Lett 17:375–391. https://doi.org/10.1007/s10311-­018-­0807-­7 Fenyvesi É, Barkács K, Gruiz K, Kenyeres I, Záray G, Szente L (2020) Removal of hazardous micropollutants from treated wastewater using cyclodextrin bead polymer – a pilot demonstration case. J Hazar Mater 383:121181. https://doi.org/1016/j.jhazmat.2019.121181 Fernández C, González-Doncel M, Pro J, Carbonell G, Tarazona JV (2010) Occurrence of pharmaceutically active compounds in surface waters of the henares-jarama-tajo river system

1  Remediation of Emerging Contaminants

83

(Madrid, Spain) and a potential risk characterization. Sci Total Environ 408:543–551. https:// doi.org/10.1016/j.scitotenv.2009.10.009 Fernández-Alvarez P, Le Noir M, Guieysse B (2009) Removal and destruction of endocrine disrupting contaminants by adsorption with molecularly imprinted polymers followed by simultaneous extraction and phototreatment. J Hazard Mater 163:1107–1112 Fernández-Castro P, Vallejo M, San Román MF, Ortiz I (2015) Insight on the fundamentals of advanced oxidation processes. Role and review of the determination methods of reactive oxygen species. J Chem Technol Biotechnol 90:796–820. https://doi.org/10.1002/jctb.4634 Ferreira RC, De Lima HHC, Cândido AA, Couto Junior OM, Arroyo PA, De Carvalho KQ, Gauze GF, Barros MASD (2015) Adsorption of paracetamol using activated carbon of Dende and Babassu coconut mesocarp. World Academy of Science, Engineering and Technology, International Science Index. Chem Mol Eng 9:717–722 Fomina M, Gadd GM (2014) Biosorption: current perspectives on concept, definition and application. Bioresour Technol 160:3–14. https://doi.org/10.1016/j.biortech.2013.12.102 Foo KY, Hameed BH (2010) Detoxification of pesticide waste via activated carbon adsorption process. J Hazard Mater 175:1–11 Fountoulakis MS, Daskalakis G, Papadaki A, Kalogerakis N, Manios T (2017) Use of halophytes in pilot-scale horizontal flow constructed wetland treating domestic wastewater. Environ Sci Pollut Control Ser 24:16682e16689. https://doi.org/10.1007/s11356-­017-­9295-­8 Freitas ED, Carmo ACR, Almeida Neto AF, Vieira MGA (2017) Binary adsorption of silver and copper on Verde-lodo bentonite: kinetic and equilibrium study. Appl Clay Sci 137:69–76. https://doi.org/10.1016/j.clay.2016.12.016 Frömming KH, Szejtli J (1994) Cyclodextrins in pharmacy. Kluwer Academic Publishers, Dordrecht Gagnon C, Lajeunesse A, Cejka P, Gagné F, Hausler R (2008) Degradation of selected acidic and neutral pharmaceutical products in a primary-treated wastewater by disinfection processes. Ozone Sci Eng 30:387–392. https://doi.org/10.1080/01919510802336731 Gallard H, Leclercq A, Croué JP (2004) Chlorination of bisphenol A: kinetics and by-products formation. Chemosphere 56:465–473. https://doi.org/10.1016/j.chemosphere.2004.03.001 Geankoplis CJ (1993) Transport and unit operations, vol 3. Prentice-Hall International Editions, New Jersey Gehr R, Wagner M, Veerasubramanian P, Payment P (2003) Disinfection efficiency of peracetic acid, UV and ozone after enhanced primary treatment of municipal wastewater. Water Res 37:4573–4586. https://doi.org/10.1016/S0043-­1354(03)00394-­4 Genç N, Dogan EC, Yurtseve M (2013) Bentonite for ciprofloxacin removal from aqueous solution. Water Sci Technol 68:848–855 Gentili FG, Fick J (2017) Algal cultivation in urban wastewater: an efficient way to reduce pharmaceutical pollutants. J Appl Phycol 29:255–262. https://doi.org/10.1007/s10811-­016-­0950-­0 Gerrity D, Stanford BD, Trenholm RA, Snyder SA (2010) An evaluation of a pilot-scale nonthermal plasma advanced oxidation process for trace organic compound degradation. Water Res 44:493–504. https://doi.org/10.1016/j.watres.2009.09.029 Ghuge SP, Saroha AK (2018) Catalytic ozonation for the treatment of synthetic and industrial effluents  – application of mesoporous materials: a review. J Environ Manag 211:83–102. https://doi.org/10.1016/j.jenvman.2018.01.052 Gilabert-Alarcón C, Salgado-Méndez SO, Walter Daesslé L, Mendoza-Espinosa LG, Villada-­ Canela M (2018) Regulatory challenges for the use of reclaimed water in Mexico: a case study in Baja California. Water 10:1432. https://doi.org/10.3390/w10101432 Giraldo AL, Erazo-Erazo ED, Flórez-Acosta OA, Serna-Galvis T-PRA (2015) Degradation of the antibiotic oxacillin in water by anodic oxidation with Ti/IrO2 anodes: evaluation of degradation routes, organic by-products and effects of water matrix components. Chem Eng J 279:103–114. https://doi.org/10.1016/j.cej.2015.04.140

84

N. Morin-Crini et al.

Githinji LJM, Musey MK, Ankumah RO (2011) Evaluation of the fate of ciprofloxacin and amoxicillin in domestic wastewater. Water Air Soil Pollut 219:191–201. https://doi.org/10.1007/ s11270-­010-­0697-­1 Gogoi A, Mazumder P, Tyagi VK, Tushara Chaminda GGT, An AK, Kumar M (2018) Occurrence and fate of emerging contaminants in water environment: a review. Groundw Sustain Dev 6:169–180. https://doi.org/10.1016/j.gsd.2017.12.009 Goldstein S, Lind J, Merenyi G (2005) Chemistry of peroxynitrites as compared to peroxynitrates. Chem Rev 105:2457–2470. https://doi.org/10.1021/cr0307087 Goméz JM, Mezcua M, Martinez MJ, Fernández-Alba AR, Agüera A (2006) A new method for monitoring estrogens, N-octylphenol, and bisphenol A in wastewater treatment plants by solid-phase extraction-gas chromatography-tandem mass spectrometry. Int J Environ Anal Chem 86:3–13 Gonçalves AG, Figueiredo JL, Órfão JJM, Pereira MFR (2010) Influence of the surface chemistry of multi-walled carbon nanotubes on their activity as ozonation catalysts. Carbon 48:4369–4381. https://doi.org/10.1016/j.carbon.2010.07.051 Gonçalves AG, Órfão JJM, Pereira MFR (2013) Ozonation of sulfamethoxazole promoted by MWCNT. Cat Com 35:82–87. https://doi.org/10.1016/j.catcom.2013.02.012 González S, Petrovic M, Barceló D (2008) Evaluation of two pilot scale membrane bioreactors for the elimination of selected surfactants from municipal wastewaters. J Hydrol 356:46–55. https://doi.org/10.1016/j.jhydrol.2008.03.023 Gonzalez S, Lopez-Roldan R, Cortina JL (2012) Presence and biological effects of emerging contaminants in Llobregat River basin: a review. Environ Poll 161:83–92. https://doi.org/10.1016/j. envpol.2011.10.002 Gorito AM, Ribeiro AR, Almeida CMR, Silva AMT (2017) A review on the application of constructed wetlands for the removal of priority substances and contaminants of emerging concern listed in recently launched EU legislation. Environ Pollut 227:428–443. https://doi. org/10.1016/j.envpol.2017.04.060 Gorito AM, Ribeiro AR, Gomes CR, Almeida CMR, Silva AMT (2018) Constructed wetland microcosms for the removal of organic micropollutants from freshwater aquaculture effluents. Sci Total Environ 644:1171–1180. https://doi.org/10.1016/j.scitotenv.2018.06.371 Gouveia L, Graça S, Sousa C, Ambrosano L, Ribeiro B, Botrel EP, Neto PC, Ferreira AF, Silva CM (2016) Microalgae biomass production using wastewater: treatment and costs. Scale-up considerations. Algal Res 16:167–176. https://doi.org/10.1016/j.algal.2016.03.010 Grabowski LR, van Veldhuizen EM, Pemen AJM, Rutgers WR (2007) Breakdown of methylene blue and methyl orange by pulsed corona discharge. Plasma Sources Sci Technol 16:226–232. https://doi.org/10.1088/0963-­0252/16/2/003 Greyshock AE, Vikesland PJ (2006) Triclosan reactivity in chloraminated waters. Environ Sci Technol 40:2615–2622. https://doi.org/10.1021/es051952d Gruiz K, Molnár M, Fenyvesi É, Cs H, Atkári A, Barkács K (2011) Cyclodextrins in innovative engineering tools for risk-based environmental management. J Incl Phenom Macrocycl Chem 70:299–306. https://doi.org/10.1007/s10847-­010-­9909-­y Grundgeiger E, Lim YH, Frost RL, Ayoko GA, Yunfei X (2015) Application of organo-­beidellites for the adsorption of atrazine. Appl Clay Sci 105-106:252–258. https://doi.org/10.1016/j. clay.2015.01.003 Gupta SS, Bhattacharyya KG (2014) Adsorption of metal ions by clays and inorganic solids. RSC Adv 4:28537–28586. https://doi.org/10.1039/C4RA03673E Guo H, Jiang N, Wang H, Lu N, Shang K, Li J, Wu Y (2019a) Pulsed discharge plasma assisted with graphene-WO3 nanocomposites for synergistic degradation of antibiotic enrofloxacin in water. Chem Eng J 372:226–240. https://doi.org/10.1016/j.cej.2019.04.119 Guo H, Jiang N, Wang H, Shang K, Lu N, Li J, Wu Y (2019b) Enhanced catalytic performance of graphene-TiO2 nanocomposites for synergetic degradation of fluoroquinolone antibiotic in pulsed discharge plasma system. Appl Catal B Environ 248:552–566. https://doi.org/10.1016/j. apcatb.2019.01.052

1  Remediation of Emerging Contaminants

85

Guo H, Jiang N, Wang H, Lu N, Shang K, Li J, Wu Y (2019c) Degradation of antibiotic chloramphenicol in water by pulsed discharge plasma combined with TiO2/WO3 composites: mechanism and degradation pathway. J Haz Mater 371:666–676. https://doi.org/10.1016/j. jhazmat.2019.03.051 Gurung K, Ncibi MC, Sillanpää M (2017) Assessing membrane fouling and the performance of pilot-scale membrane bioreactor (MBR) to treat real municipal wastewater during winter season in Nordic regions. Sci Total Environ 579:1289–1297 Hama Aziz KH, Miessner H, Mueller S, Mahyar A, Kalass D, Moeller D, Khorshid I, MAM R (2018) Comparative study on 2,4-dichlorophenoxyacetic acid and 2,4-­ dichlorophenol removal from aqueous solutions via ozonation, photocatalysis and non-thermal plasma using a planar falling film reactor. J Hazard Mater 343:107–115. https://doi.org/10.1016/j. jhazmat.2017.09.025 Hansen R, Andersen HR (2012) Energy effectiveness of direct UV and UV/H2O2 treatment of estrogenic chemicals in biologically treated sewage. Int J Photoenergy 270320. https://doi. org/10.1155/2012/270320 Haque E, Lee JE, Jang IT, Hwang YK, Chang JS, Jegal J, Jhung SH (2010) Adsorptive removal of methyl orange from aqueous solution with metal-organic frameworks, porous chromium-­ benzenedicarboxylates. J Hazard Mater 181:535–542 He J, Ding L, Deng J, Yang W (2012) Oil-absorbent beads containing beta-cyclodextrin moieties: preparation via suspension polymerization and high oil absorbency. Polym Adv Technol 23:810–816. https://doi.org/10.1002/pat.1975 He D, Sun Y, Xin L, Feng J (2014) Aqueous tetracycline degradation by non-thermal plasma combined with nano-TiO2. Chem Eng J 258:18–25. https://doi.org/10.1016/j.cej.2014.07.089 He D, Sun Y, Li S, Feng J (2015) Decomposition of tetracycline in aqueous solution by corona discharge plasma combined with a Bi2MoO6 nanocatalyst. J Chem Technol Biotechnol 90:2249–2256. https://doi.org/10.1002/jctb.4540 Hemine K, Skwierawska A, Kernstein A, Kozłowska-Tylingo K (2020) Cyclodextrin polymers as efficient adsorbents for removing toxic non-biodegradable pimavanserin from pharmaceutical wastewaters. Chemosphere 250:126250. https://doi.org/10.1016/j.chemosphere.2020.126250 Herrera-Morales J, Morales K, Ramos D, Ortiz-Quiles EO, Lopez-Encarnación JM, Nicolau E (2017) Examining the use of nanocellulose composites for the sorption of contaminants of emerging concern: an experimental and computational study. ACS Omega 2:7714–7722. https://doi.org/10.1021/acsomega.7b01053 Herrera-Morales J, Turley TA, Betancourt-Ponce M, Nicolau E (2019) Nanocellulose-block copolymer films for the removal of emerging organic contaminants from aqueous solutions. Materials 12:230. https://doi.org/10.3390/ma12020230 Hey C, Grabi R, Ledin A, la Cour JJ, Andersen HR (2012a) Oxidation of pharmaceuticals by chlorine dioxide in biologically treated wastewater. Chem Eng J 185-186:236–242. https://doi. org/10.1016/j.cej.2012.01.093 Hey G, Ledin A, la Cour JJ, Andersen HR (2012b) Removal of pharmaceuticals in biologically treated wastewater by chlorine dioxide or peracetic acid. Environ Technol 33:1041–1047. https://doi.org/10.1080/09593330.2011.606282 Hijosa-Valsero M, Molina R, Schikora H, Müller M, Bayona JM (2013) Removal of priority pollutants from water by means of dielectric barrier discharge atmospheric plasma. J Hazard Mater 262:664–667. https://doi.org/10.1016/j.jhazmat.2013.09.022 Hokkanen S, Repo E, Sillanpää M (2013) Removal of heavy metals from aqueous solutions by succinic anhydride modified mercerized nanocellulose. Chem Eng J 223:40–47 Hu JY, Aizawa T, Ookubo S (2002) Products of aqueous chlorination of bisphenol A and their estrogenic activity. Environ Sci Technol 36:1980–1987 Hu J, Cheng S, Aizawa T, Terao Y, Kunikane S (2003) Products of aqueous chlorination of 17a-estradiol and their estrogenic activities. Environ Sci Technol 37:5665–5670. https://doi. org/10.1021/es034324+

86

N. Morin-Crini et al.

Hu Y, Bai Y, Yu H, Zhang C, Chen J (2013) Degradation of selected organophosphate pesticides in wastewater by dielectric barrier discharge plasma. Bull Environ Contam Toxicol 91:314–319. https://doi.org/10.1007/s00128-­013-­1048-­x Hu Y, Peng Y, Liu W, Zhao D, Fu J (2017) Chapter 4: Removal of emerging contaminants from water and wastewater using nanofiltration technology. In: Joo SH (ed) Applying nanotechnology for environmental sustainability. https://doi.org/10.4018/978-­1-­5225-­0585-­3.ch004 Hu S, Liu X, Xu Z, Wang J, Li Y, Shen J, Lan Y, Cheng C (2019) Degradation and mineralization of ciprofloxacin by gas-liquid discharge nonthermal plasma. Plasma Sci Technol 21:015501. https://doi.org/10.1088/2058-­6272/aade82 Huang X, Xiao K, Shen Y (2010) Recent advances in membrane bioreactor technology for wastewater treatment in China. Front Environ Sci Eng China 4:245–271. https://doi.org/10.1007/ s11783-­010-­0240-­z Huang DL, Wang RZ, Liu YG, Zeng GM, Lai C, Xu P, Lu BA, Xu JJ, Wang C, Huang C (2015) Application of molecularly imprinted polymers in wastewater treatment: a review. Environ Sci Pollut Res 22:963–977 Huang X, Ye G, Yi N, Lu L, Zhang L, Yang L, Xiao L, Liu J (2019) Effect of plant physiological characteristics on the removal of conventional and emerging pollutants from aquaculture wastewater by constructed wetlands. Ecol Eng 135:45–53. https://doi.org/10.1016/j. ecoleng.2019.05.017 Huang Q, Chai K, Zhou L, Ji H (2020) A phenyl-rich β-cyclodextrin porous crosslinked polymer for efficient removal of aromatic pollutants: insight into adsorption performance and mechanism. Chem Eng J 387:124020. https://doi.org/10.1016/j.cej.2020.124020 Huber M, Canonica S, Park G, von Gunten U (2003) Oxidation of pharmaceuticals during ozonation and advanced oxidation processes. Environ Sci Technol 37:1016–1024 Huber MM, Ternes TA, von Guten U (2004) Removal of estrogenic activity and formation of oxidation products during ozonation of 17a-ethynylestradiol. Environ Sci Technol 38:5177–5186. https://doi.org/10.1021/es035205x Huber MM, Korhonen S, Ternes TA, von Gunten U (2005a) Oxidation of pharmaceuticals during water treatment with chlorine dioxide. Water Res 39:3607–3617 Huber M, Göbel A, Joss A, Hermann N, Löffler D, McArdell C, Ried A, Siegrist H, Ternes T, von Gunten U (2005b) Oxidation of pharmaceuticals during ozonation of municipal wastewater effluents: a pilot study. Environ Sci Technol 39:4290–4299 Iervolino G, Vaiano V, Palma V (2019) Enhanced removal of water pollutants by dielectric barrier discharge nonthermal plasma reactor. Sep Purif Technol 215:155–162. https://doi. org/10.1016/j.seppur.2019.01.007 Ikawa S, Tani A, Nakashima Y, Kitano K (2016) Physicochemical properties of bactericidal plasma-­ treated water. J Phys D-Appl Phys 49:425401. https://doi.org/10.1088/0022-­3727/49/42/425401 Ikehata K, Naghashkar NJ, Gamal El-Din M (2006) Degradation of aqueous pharmaceuticals by ozonation and advanced oxidation processes: a review. Ozone Sci Eng 28:353–414. https:// doi-­­org.inee.bib.cnrs.fr/10.1080/01919510600985937 Ikehata K, Gamal El-Din M, Snyder SA (2008) Ozonation and advanced oxidation treatment of emerging organic pollutants in water and wastewater. Ozone Sci Eng 30:21–26. https://doi-­org. inee.bib.cnrs.fr/10.1080/01919510701728970 Iorhemen OT, Hamza RA, Tay JH (2016) Membrane bioreactor (MBR) technology for wastewater treatment and reclamation: membrane fouling. Membranes 6:33. https://doi.org/10.3390/ membranes6020033 Irmak S, Erbatur O, Akgerman A (2005) Degradation of 17b-estradiol and bisphenol A in aqueous medium by using ozone and ozone/UV techniques. J Hazard Mater B 126:54–62 Jalilnejad E, Sadeghpour P, Ghasemzadeh K (2020) Advances in membrane bioreactor technology. In: current trends and future developments on (bio-)membranes. Ceramic membranes bioreactors. Basile A, Ghasemzadeh and Jalilnejad E (Eds.). Elsevier. Chapitre 1:1–29. https://doi. org/10.1016/B978-­0-­12-­816822-­6.00001-­X

1  Remediation of Emerging Contaminants

87

Janus L, Crini G, El-Rezzi V, Morcellet M, Cambiaghi A, Torri G, Naggi A, Vecchi C (1999) New sorbents containing beta-cyclodextrin. Synthesis, characterization, and sorption properties. React Funct Polym 42:173–180. https://doi.org/10.1016/s1381-­5148(98)00066-­2 Jeirani Z, Niu CH, Soltan J (2017) Adsorption of emerging pollutants on activated carbon. Rev Chem Eng 33:491–522. https://doi.org/10.1515/revce-­2016-­0027 Jelic A, Gros M, Ginebreda A, Cespedes-Sánchez R, Ventura F, Petrovic M, Barcelo D (2011) Occurrence, partition and removal of pharmaceuticals in sewage water and sludge during wastewater treatment. Water Res 45:1165–1176. https://doi.org/10.1016/j.watres.2010.11.010 Jiang B, Zheng J, Qiu S, Wu M, Zhang Q, Yan Z, Xue Q (2014) Review on electrical discharge plasma technology for wastewater remediation. Chem Eng J 236:348–368. https://doi. org/10.1016/j.cej.2013.09.090 Jiang L, Liu Y, Liu S, Li M, Hu X, Zeng G, Hu X, Liu S, Liu S, Huang B, Li M (2016) Fabrication of β-cyclodextrin/poly(l-glutamic acid) supported magnetic graphene oxide and its adsorption behavior for 17β-estradiol. Chem Eng J 308:597–605. https://doi.org/10.1016/j.cej.2016.09.067 Jiang Q, Ngo HH, Nghiem LD, Hai FI, Price WE, Zhang J, Guo W (2018) Effect of hydraulic retention time on the performance of a hybrid moving bed biofilm reactor-membrane bioreactor system for micropollutants removal from municipal wastewater. Bioresour Technol 247:1228–1232. https://doi.org/10.1016/j.biortech.2017.09.114 Johnson AC, Sumpter JP (2001) Removal of endocrine-disrupting chemicals in activated sludge treatment works. Environ Sci Technol 35:4697–4703 Jojoa-Sierra SD, Silva-Agredo J, Herrera-Calderon E, Torres-Palma RA (2017) Elimination of the antibiotic norfloxacin in municipal wastewater, urine and seawater by electrochemical oxidation on IrO2 anodes. Sci Total Environ 575:1228–1238. https://doi.org/10.1016/j. scitotenv.2016.09.201 Jović M, Manojlović D, Stanković D, Dojčinović B, Obradović B, Gašić U, Roglić G (2013) Degradation of triketone herbicides, mesotrione and sulcotrione, using advanced oxidation processes. J Hazard Mater 260:1092–1099. https://doi.org/10.1016/j.jhazmat.2013.06.073 Jović MS, Dojčinović BP, Kovačević VV, Obradović BM, Kuraica M, Gašić UM, Roglić GM (2014) Effect of different catalysts on mesotrione degradation in water falling film DBD reactor. Chem Eng J 248:63–70. https://doi.org/10.1016/j.cej.2014.03.031 Judd S (2008) The status of membrane bioreactor technology. Trends Biotechnol 26:109–116. https://doi.org/10.1016/j.tibtech.2007.11.005 Jurecska L, Dobosy P, Barkács K, Fenyvesi É, Záray G (2014) Characterization of cyclodextrin containing nanofilters for removal of pharmaceutical residues. J Pharm Biomed Anal 98:90–93. https://doi.org/10.1016/j.jpba.2014.05.007 Jux U, Baginski RM, Arnold HG, Krönke M, Seng PN (2002) Detection of pharmaceutical contaminations of river, pond, and tap water from Cologne (Germany) and surroundings. Int J Hyg Environ Health 205:393–398 Kanazawa S, Furuki T, Nakaji T, Akamine S, Ichiki R (2013) Application of chemical dosimetry to hydroxyl radical measurement during underwater discharge. J Phys Conf Ser 418. https://doi. org/10.1088/1742-­6596/418/1/012102 Kang J, Liu H, Zheng YM, Qu J, Chen JP (2010) Systematic study of synergistic and antagonistic effects on adsorption of tetracycline and copper onto a chitosan. J Colloid Interf Sci 344:117–125. https://doi.org/10.1016/j.jcis.2009.11.049 Karigar CS, Rao SS (2011) Role of microbial enzymes in the bioremediation of pollutants: a review. Enzyme Res 805187. https://doi.org/10.4061/2011/805187 Karthikeyan KG, Meyer MT (2006) Occurrence of antibiotics in wastewater treatment facilities in Wisconsin, USA.  Sci Total Environ 361:196–207. https://doi.org/10.1016/j. scitotenv.2005.06.030 Kasprzyk-Hordern B, Dinsdale RM, Guwy AJ (2008) The occurrence of pharmaceuticals, personal care products, endocrine disruptors and illicit drugs in surface water in South Wales, UK. Water Res 42:3498–3518

88

N. Morin-Crini et al.

Katsigiannis A, Noutsopoulos C, Mantziaras J, Gioldasi M (2015) Removal of emerging pollutants through granular activated carbon. Chem Eng J 280:49–57. https://doi.org/10.1016/j. cej.2015.05.109 Khan M, Tahir MN, Adil SF, Khan HU, Siddiqui MRH, Al-Warthan AA, Tremel W (2015) Graphene based metal and metal oxide nanocomposites: synthesis, properties and their applications. J Mater Chem A 3:18753–18808. https://doi.org/10.1039/C5TA02240A Khan NA, Khan SU, Ahmed S, Farooqi IH, Yousefi M, Mohammadi AA, Changani F (2020) Recent trends in disposal and treatment technologies of emerging-pollutants – a critical review. Trends Anal Chem 122:115744. https://doi.org/10.1016/j.trac.2019.115744 Khalaf MN (2016) Green polymers and environmental pollution control. CRC Press; Apple Academic Press, Inc, Oakville, 436 p Kim I, Tanaka H (2009) Photodegradation characteristics of PPCPs in water with UV treatment. Environ Int 35:793–802. https://doi.org/10.1016/j.envint.2009.01.003 Kim SE, Yamada H, Tsuno H (2004) Evaluation of estrogenicity for 17β-estradiol decomposition during ozonation. Ozone-Sci Eng 26:563–571 Kim I, Yamashita N, Tanaka H (2009a) Performance of UV and UV/H2O2 processes for the removal of pharmaceuticals detected in secondary effluent of a sewage treatment plant in Japan. J Hazard Mater 166:1134–1140. https://doi.org/10.1016/j.jhazmat.2008.12.020 Kim I, Yamashita N, Tanaka H (2009b) Photodegradation of pharmaceuticals and personal care products during UV and UV/H2O2 treatments. Chemosphere 77:518–525. https://doi. org/10.1016/j.chemosphere.2009.07.041 Kim KS, Yang CS, Mok YS (2013) Degradation of veterinary antibiotics by dielectric barrier discharge plasma. Chem Eng J 219:19–27. https://doi.org/10.1016/j.cej.2012.12.079 Kim KS, Kamb SK, Mok YS (2015) Elucidation of the degradation pathways of sulfonamide antibiotics in a dielectric barrier discharge plasma system. Chem Eng J 271:31–42. https://doi. org/10.1016/j.cej.2015.02.073 Kim S, Chu KH, Al-Hamadani YAJ, Park CM, Jang M, Kim DH, Yu M, Heo J, Yon Y (2018) Removal of contaminants of emerging concern by membranes in water and wastewater: a review. Chem Eng J 335:896–914. https://doi.org/10.1016/j.cej.2017.11.044 Kim MS, Cha D, Lee KM, Lee HJ, Kim T, Lee C (2020) Modeling of ozone decomposition, oxidant exposures, and the abatement of micropollutants during ozonation processes. Water Res 169:115230. https://doi.org/10.1016/j.watres.2019.115230 Kolpin DW, Furlong ET, Meyer MT, Thurman EM, Zaugg SD, Barber LB, Buxton HT (2002) Pharmaceuticals, hormones, and other organic wastewater contaminants in US streams, 1999-2000: a national reconnaissance. Environ Sci Technol 36:1202–1211. https://doi. org/10.1021/es011055j Kono H, Onishi K, Nakamura T (2013) Characterization and bisphenol A adsorption capacity of beta- cyclodextrin-carboxymethylcellulose-based hydrogels. Carbohydr Polym 98:784–792. https://doi.org/10.1016/j.carbpol.2013.06.065 Korshin G, Kim J, Gan L (2006) Comparative study of reactions of endocrine disruptors bisphenol A and diethylstilbestrol in electrochemical treatment and chlorination. Water Res 40:1070–1078 Kovalova L, Siegrist H, von Gunten U, Eugster J, Hagenbuch M, Wittmer A, Moser R, McArdell CS (2013) Elimination of micropollutants during post-treatment of hospital wastewater with powdered activated carbon, ozone and UV.  Environ Sci Technol 47:7899–7908. https://doi. org/10.1021/es400708w Kunst B, Košutić K (2008) Removal of emerging contaminants in water treatment by nanofiltration and reverse osmosis. In: Hutzinger O, Barceló D, Kostianoy A (eds) Handbook of environmental chemistry. Volume 5: water Pollution, vol 5. Springer, Berlin, pp 103–125. https://doi. org/10.1007/698_5_100 Kurwadkar S, Hoang TV, Malwade K, Kanel SR, Harper WF Jr, Struckhoff G (2019) Application of carbon nanotubes for removal of emerging contaminants of concern in engineered water and wastewater treatment systems. Nanotechnol Environ Eng 4:12. https://doi.org/10.1007/ s41204-­019-­0059-­1

1  Remediation of Emerging Contaminants

89

Kruithof JC, Kamp PC, Martijn BJ (2007) UV/H2O2 treatment: a practical solution for organic contaminant control and primary disinfection. Ozone Sci Eng 29:273–280 Kyzas GZ, Kostoglou M (2014) Green adsorbents for wastewaters: a critical review. Materials 7:333–364. https://doi.org/10.3390/ma7010333 Lacson CFZ, de Luna MDG, Dong C, Garcia-Segura S, Lu MC (2018) Fluidized-bed Fenton treatment of imidacloprid: optimization and degradation pathway. Sustainable Environ Res 28:309–314. https://doi.org/10.1016/j.serj.2018.09.001 Lam E, Male KB, Chong JH, Leung ACW, Luong JHT (2012) Applications of functionalized and nanoparticle-modified nanocrystalline cellulose. Trends Biotechnol 30:283–290 Landy D, Mallard I, Ponchel A, Monflier E, Fourmentin S (2012) Remediation technologies using cyclodextrins: an overview. Environ Chem Lett 10:225–237 Larsen TA, Lienert J, Joos A, Siegrist H (2004) How to avoid pharmaceuticals in the aquatic environment. J Biotechnol 113:295–304. https://doi.org/10.1016/j.jbiotec.2004.03.033 Le Noir M, Lepeuple AS, Guieysse B, Mattiasson B (2007) Selective removal of 17beta-estradiol at trace concentration using a molecularly imprinted polymer. Water Res 41:2825–2831 Lee Y, von Gunten U (2010) Oxidative transformation of micropollutants during municipal wastewater treatment: comparison of kinetic aspects of selective (chlorine, chlorine dioxide, ferrate VI, and ozone) and non-selective oxidants (hydroxyl radical). Water Res 44:555–566 Lee Y, von Gunten U (2012) Quantitative structure-activity relationships (QSARs) for the transformation of organic micropollutants during oxidative water treatment. Water Res 46:6177–6195 Lee Y, von Gunten U (2016) Advances in predicting organic contaminant abatement during ozonation of municipal wastewater effluent: reaction kinetics, transformation products, and changes of biological effects. Environ Sci Water Res Technol 2:421–442. https://doi.org/10.1039/ C6EW00025H Lee Y, Gerrity D, Lee M, Bogeat AE, Salhi E, Gamage S, Trenholm RA, Wert EC, Snyder SA, von Gunten U (2013) Prediction of micropollutant elimination during ozonation of municipal wastewater effluents: use of kinetic and water specific information. Environ Sci Technol 47:5872–5881. https://doi.org/10.1021/es400781r Lee Y, Kovalova L, McArdell C, von Gunten U (2014) Prediction of micropollutant elimination during ozonation of a hospital wastewater effluent. Water Res 64:134–148. https://doi. org/10.1016/j.watres.2014.06.027 Lee D, Lee JC, Nam JY, Kim HW (2018) Degradation of sulfonamide antibiotics and their intermediates toxicity in an aeration-assisted non-thermal plasma while treating strong wastewater. Chemosphere 209:901–907. https://doi.org/10.1016/j.chemosphere.2018.06.125 Lei H, Snyder SA (2007) 3D QSPR models for the removal of trace organic contaminants by ozone and free chlorine. Water Res 41:4051–4060. https://doi.org/10.1016/j.watres.2007.05.010 Leitner N, Roshani B (2010) Kinetic of benzotriazole oxidation by ozone and hydroxyl radicals. Water Res 44:2058–2066. https://doi.org/10.1016/j.watres.2009.12.018 Li B, Zhang T (2011) Mass flows and removal of antibiotics in two municipal wastewater treatment plants. Chemosphere 83:1284–1289 Li SP, Jiang YY, Cao XH, Dong YW, Dong M, Xu J (2013) Degradation of nitenpyram pesticide in aqueous solution by low-temperature plasma. Environ Technol 34:1609–1616. https://doi. org/10.1080/09593330.2013.765914 Li G, Lu Y, Lu C, Zhu M, Zhai C, Du Y, Yang P (2015) Efficient catalytic ozonation of bisphenol-a over reduced graphene oxide modified sea urchin-like α-MnO2architectures. J Hazard Mater 294:201–208 Li X, Wang B, Cao Y, Zhao S, Wang H, Feng X, Zhou J, Ma X (2019a) Water contaminant elimination based on metal-organic frameworks and perspective on their industrial applications. Sustain ACS Chem Eng 7:4548–4563. https://doi.org/10.1021/acssuschemeng.8b05751 Li X, Wang B, Cao Y, Zhao S, Wang H, Feng X, Zhou J, Ma X (2019b) Water contaminant elimination based on metal-organic frameworks and perspective on their industrial applications. ACS Sustain Chem Eng 7:4548–4563. https://doi.org/10.1021/acssuschemeng.8b05751

90

N. Morin-Crini et al.

Li H, Li T, He S, Zhou J, Wang T, Zhu L (2020) Efficient degradation of antibiotics by non-thermal discharge plasma: highlight the impacts of molecular structures and degradation pathways. Chem Eng J 395:125091. https://doi.org/10.1016/j.cej.2020.125091 Lian F, Sun B, Song Z, Zhu L, Qi X, Xing B (2014) Physicochemical properties of herb-residue biochar and its sorption to ionizable antibiotic sulfamethoxazole. Chem Eng J 248:128–134. https://doi.org/10.1016/j.cej.2014.03.021 Lian L, Yao B, Hou S, Fang J, Yan S, Song W (2017) Kinetic study of hydroxyl and sulfate radical-mediated oxidation of pharmaceuticals in wastewater effluents. Environ Sci Technol 51:2954–2962. https://doi.org/10.1021/acs.est.6b05536 Lim M, Patureau D, Heran M, Lesage G, Kim J (2020) Removal of organic micropollutants in anaerobic membrane bioreactors in wastewater treatment: critical review. Environ Sci Water Res. https://doi.org/10.1039/C9EW01058K Lin Y, Peng Z, Zhang X (2009) Ozonation of estrone, estradiol, diethylstilbestrol in waters. Des 249:235–240. https://doi.org/10.1016/j.desal.2008.06.034 Lin Y, Baggett DW, Kim JW, Siochi EJ, Connell JW (2011) Instantaneous formation of metal and metal oxide nanoparticles on carbon nanotubes and graphene via solvent-free microwave heating. Appl Mater Interfaces 3:1652–1664. https://doi.org/10.1021/am200209e Lindberg R, Jarnheimer PA, Olsen B, Johonsson M, Tysklind M (2004) Determination of antibiotic substances in hospital sewage water using solid phase extraction and liquid chromatography/mass spectrometry and group analogue internal standards. Chemosphere 57:1479–1488. https://doi.org/10.1016/j.chemosphere.2004.09.015 Ling Y, Klemes MJ, Xiao L, Alsbaiee A, Dichtel WR, Helbling DE (2017) Benchmarking micropollutant removal by activated carbon and porous β-cyclodextrin polymers under environmentally relevant scenarios. Environ Sci Technol 51:7590–7598. https://doi.org/10.1021/acs. est.7b00906 Liu DHF, Liptak BG (eds) (2000) Wastewater treatment. CRC Press, Boca Raton Liu ZH, Kanjo Y, Mizutami S (2009) Removal mechanisms for endocrine disrupting compounds (EDCs) in wastewater treatment  – physical means, biodegradation, and chemical advanced oxidation: a review. Sci Total Environ 407:731–748 Liu N, Wang M, Liu M, Liu F, Weng L, Koopal LK, Tan W (2012a) Sorption of tetracycline on organo-montmorillonites. J Hazard Mater 225-226:28–35. https://doi.org/10.1016/j. jhazmat.2012.04.060 Liu P, Liu WJ, Jiang H, Chen JJ, Li WW, Yu HQ (2012b) Modification of bio-char derived from fast pyrolysis of biomass and its application in removal of tetracycline from aqueous solution. Bioresour Technol 121:235–240. https://doi.org/10.1016/j.biortech.2012.06.085 Liu JN, Chen Z, Wu QY, Li A, Hu HY, Yang C (2016) Ozone/graphene oxide catalytic oxidation: a novel method to degrade emerging organic contaminant N, N-diethyl-m-toluamide (DEET). Sci Rep 6:1–9 Locke BR, Shih KY (2011) Review of the methods to form hydrogen peroxide in electrical discharge plasma with liquid water. Plasma Sources Sci Technol 20:034006. https://doi. org/10.1088/0963-­0252/20/3/034006 Locke BR, Lukes P, Brisset JL (2012) Chapter 6: Elementary chemical and physical phenomena in electrical discharge plasma in gas-liquid environments and in liquids. In: Parvulescu VI, Magureanu M, Lukes P (eds) Plasma chemistry and catalysis in gases and liquids. Wiley VCH, Weinheim, pp 185–241 Lofrano G (2012) Emerging compounds removal from wastewater: natural and solar based treatments. Springer, Dordrecht. https://doi.org/10.1007/978-­94-­007-­3916-­1 Loftsson T, Jarho P, Másson M, Järvinen T (2005) Cyclodextrins in drug delivery (review). Expert Opin Drug Delivery 2:335–351. https://doi.org/10.1517/17425247.2.1.335 Lopez A, Anna B, Giuseppe M, John K (2003) Kinetic investigation on UV and UV/H2O2 degradations of pharmaceutical intermediates in aqueous solution. J.  Photochem Photobiol 156:121–126

1  Remediation of Emerging Contaminants

91

Lukes P, Appleton AT, Locke BR (2004) Hydrogen peroxide and ozone formation in hybrid gas-­ liquid electrical discharge reactors. IEEE Trans Ind Appl 40:60–67. https://doi.org/10.1109/ TIA.2003.821799 Lukes P, Clupek M, Babicky V, Janda V, Sunka P (2005) Generation of ozone by pulsed corona discharge over water surface in hybrid gase liquid electrical discharge reactor. J Phys D Appl Phys 38:409e416. https://doi.org/10.1088/0022-­3727/38/3/010 Lukes P, Locke BR, Brisset JL (2012) Chapter 7: Aqueous-phase chemistry of electrical discharge plasma in water and in gas-liquid environments. In: Parvulescu VI, Magureanu M, Lukes P (eds) Plasma chemistry and catalysis in gases and liquids. Wiley-VCH, Weinheim, pp 243–308 Ma JX, Dai RB, Chen M, Khan SJ, Wang ZW (2018) Applications of membrane bioreactors for water reclamation: micropollutant removal, mechanisms and perspectives. Bioresour Technol 269:532–543. https://doi.org/10.1016/j.biortech.2018.08.121 Macedo CC, de Andrade JR, da Silva MGC, Vieira MGA (2020) Removal of propranolol hydrochloride by batch biosorption using remaining biomass of alginate extraction from Sargassum filipendula algae. Environ Sci Pollut Res 28:1–13. https://doi.org/10.1007/s11356-­020-­08109-­4 Machado KC, Grassi MT, Vidal C, Pescara IC, Jardim WF, Fernandes AN, Sodré FF, Almeida FV, Santana JS, Canela CC, Nunes CRO, Bichinho KM, Severo FJR (2016) A preliminary nationwide survey of the presence of emerging contaminants in drinking and source waters in Brazil. Sci Total Environ 572:138–146. https://doi.org/10.1016/j.scitotenv.2016.07.210 Madikizela LM, Tavengwa NT, Chimuka L (2018) Applications of molecularly imprinted polymers for solid-phase extraction of non-steroidal anti-inflammatory drugs and analgesics from environmental waters and biological samples. J Pharm Biomed Anal 147:624–633 Maeng SK, Choi BG, Lee KT, Song KG (2013) Influences of solid retention time, nitrification and microbial activity on the attenuation of pharmaceuticals and estrogens in membrane bioreactors. Water Res 47:3151–3162 Magureanu M, Mandache NB, Parvulescu VI (2007) Degradation of organic dyes in water by electrical discharges. Plasma Chem Plasma Process 27:589–598., https://doi.org/. https://doi. org/10.1007/s11090-­007-­9087-­x Magureanu M, Piroi D, Gherendi F, Mandache NB, Parvulescu VI (2008) Decomposition of methylene blue in water by corona discharges. Plasma Chem Plasma Process 28:677–688. https:// doi.org/. https://doi.org/10.1007/s11090-­008-­9155-­x Magureanu M, Piroi D, Mandache NB, David V, Medvedovici A, Bradu C, Parvulescu VI (2011) Degradation of antibiotics in water by non-thermal plasma treatment. Water Res 45:3407–3416. https://doi.org/10.1016/j.watres.2011.03.057 Magureanu M, Dobrin D, Bradu C, Gherendi F, Mandache NB, Parvulescu VI (2016) New evidence on the formation of oxidizing species in corona discharge in contact with liquid and their reactions with organic compounds. Chemosphere 165:507e514. https://doi.org/10.1016/j. chemosphere.2016.09.073 Magureanu M, Parvulescu VI (2016) Plasma in liquids and gas-liquid environment. In: Shohet L (ed) Encyclopedia of plasma technology. Taylor and Francis Magureanu M, Bradu C, Parvulescu VI (2018) Plasma processes for the treatment of water contaminated with harmful organic compounds. J Phys D Appl Phys 51:313002. https://doi. org/10.1088/1361-­6463/aacd9c Mahfoudhi N, Boufi S (2017) Nanocellulose as a novel nanostructured adsorbent for environmental remediation: a review. Cellulose 24:1171–1197 Mahmood Q, Pervez A, Zeb BS, Zaffar H, Yaqoob H, Waseem M, Zahidullah AS (2013) Natural treatment systems as sustainable ecotechnologies for the developing countries. Biomed Res Int 796373. https://doi.org/10.1155/2013/796373 Maia GS, de Andrade JR, Oliveira MF, Vieira MGA, da Silva MGC (2017) Affinity studies between drugs and clays as adsorbent material. Chem Eng Trans 57:583–588. https://doi.org/10.3303/ CET1757098

92

N. Morin-Crini et al.

Maia GS, de Andrade JR, da Silva MGC, Vieira MGA (2019) Adsorption of diclofenac sodium onto commercial organoclay: kinetic, equilibrium and thermodynamic study. Powder Technol 345:140–150. https://doi.org/10.1016/j.powtec.2018.12.097 Mailler R, Gasperi J, Coquet Y, Buleté A, Vulliet E, Deshayes S, Zedek S, Mirande-Bret C, Eudes V, Bressy A, Caupos E, Moilleron R, Chebbo G, Rocher V (2016) Removal of a wide range of emerging pollutants from wastewater treatment plant discharges by micro-grain activated carbon in fluidized bed as tertiary treatment at large pilot scale. Sci Total Environ 542:983–996. https://doi.org/10.1016/j.scitotenv.2015.10.153 Malik MA, ur-Rehman U, Ghaffar A, Ahmed K (2002) Synergistic effect of pulsed corona discharges and ozonation on decolourization of methylene blue in water. Plasma Sources Sci Technol 11:236–240. https://doi.org/10.1088/0963-­0252/11/3/302 Malik MA (2010) Water purification by plasmas: which reactors are most energy efficient? Plasma Chem Plasma Process 30:21–31. https://doi.org/10.1007/s11090-­009-­9202-­2 Maniero MG, Bila DM, Dezotti M (2008) Degradation and estrogenic activity removal of 17b-­ estradiol and 17a-ethinylestradiol by ozonation and O3/H2O2. Sci Total Environ 407:105–115 Marco-Urrea E, Perez-Trujillo M, Vicent T, Caminal G (2009) Ability of White-rot fungi to remove selected pharmaceuticals and degradation products of ibuprofen by Trametes versicolor. Chemosphere 74:765–772 Margot J, Kienle C, Magnet A, Weil M, Rossi L, de Alencastro LF, Abegglen C, Thonney D, Chevre N, Scharer M, Barry DA (2013) Treatment of micropollutants in municipal wastewater: ozone or powdered activated carbon? Sci Total Environ 461-462:480–498 Marotta E, Schiorlin M, Ren X, Rea M, Paradisi C (2011) Advanced oxidation process for degradation of aqueous phenol in a dielectric barrier discharge reactor. Plasma Process Polym 8:867e875. https://doi.org/10.1002/ppap.201100036 Martínez-Pachón D, Espinosa-Barrera P, Rincón-Ortíz J, Moncayo-Lasso A (2019) Advanced oxidation of antihypertensives losartan and valsartan by photo-electro-Fenton at near-neutral pH using natural organic acids and a dimensional stable anode-gas diffusion electrode (DSA-GDE) system under light emission diode (LED) lighting. Environ Sci Pollut Res 26:4426–4437. https://doi-­org.inee.bib.cnrs.fr/10.1007/s11356-­018-­2645-­3 Matamoros V, Gutiérrez R, Ferrer I, García J, Bayona JM (2015) Capability of microalgae-based wastewater treatment systems to remove emerging organic contaminants: a pilot-scale study. J Hazard Mater 288:34–42. https://doi.org/10.1016/j.jhazmat.2015.02.002 Mautner A (2020) Nanocellulose water treatment membranes and filters: a review. Polym Int. https://doi.org/10.1002/pi.5993 Mawhinney DB, Vanderford BJ, Snyder SA (2012) Transformation of 1H-benzotriazole by ozone in aqueous solution. Environ Sci Technol 46:7102–7111. https://doi.org/10.1021/es300338e Meffe R, Bustamante I (2014) Emerging organic contaminants in surface water and groundwater: a first overview of the situation in Italy. Sci Total Environ 481:280–295 Mehran Abtahi S, Ilyas S, Joannis Cassan C, Albasi C, Vos WM (2018) Micropollutants removal from secondary-treated municipal wastewater using weak polyelectrolyte multilayer based nanofiltration membranes. J Membrane Sci 548:654–666 Meneghetti Campos J, Queiroz SCN, Roston DM (2019) Removal of the endocrine disruptors ethinyl estradiol, bisphenol A, and levonorgestrel by subsurface constructed wetlands. Sci Total Environ 693:133514. https://doi.org/10.1016/j.scitotenv.2019.07.320 Meneses-Jácome A, Diaz-Chavez R, Velásquez-Arredondo IH, Cárdenas-Chávez DL, Parra R, Ruiz-Colorado AA (2016) Sustainable energy from agro-industrial wastewaters in Latin America. Renew Sust Energ Rev 56:1249–1262. https://doi.org/10.1016/j.rser.2015.12.036 Mercurio M, Sarkar B, Langella A (2019) Modified clay and zeolite nanocomposite materials. Environmental and pharmaceuticals applications. Elsevier Science Publishing Co. Inc, pp 167–196. ISBN: 978-0-12-814617-0

1  Remediation of Emerging Contaminants

93

Mert BK, Ozengin N, Dogan EC, Aydiner C (2018) Efficient removal approach of micropollutants in wastewater using membrane bioreactor. In: Yonar T (ed) Wastewater and water quality. IntechOpen. https://doi.org/10.5772/intechopen.75183 Mezzelani M, Gorbi S, Regoli F (2018) Pharmaceuticals in the aquatic environments: evidence of emerged threat and future challenges for marine organisms. Marine Environ Res 140:41–60. https://doi.org/10.1016/j.marenvres.2018.05.001 Miklos DB, Remy C, Jekel M, Linden KG, Drewes JE, Hübner U (2018) Evaluation of advanced oxidation processes for water and wastewater treatment  – a critical review. Water Res 139:118–131. https://doi.org/10.1016/j.watres.2018.03.042 Mishra NS, Reddy R, Kuila A, Rani A, Mukherjee P, Nawaz A, Pichiah S (2017) A review on advanced oxidation processes for effective water treatment. Curr World Environ 12. https://doi. org/10.12944/CWE.12.3.02 Mocanu G, Miahi D, LeCerf D, Picton L, Moscovici M (2009) Cyclodextrins anionic polysaccharide hydrogels: synthesis, characterization, and interaction with some organic molecules (water pollutants, drugs, proteins). J Appl Polym Sci 112:1175–1183. https://doi.org/10.1002/ app.29580 Mohammed N, Grishkewich N, Tam KC (2018) Cellulose nanomaterials: promising sustainable nanomaterials for application in water/wastewater treatment processes. Environ Sci Nano 5:623–658 Mole RA, Good CJ, Stebel EK, Higgins JF, Pitel SA, Welch AR, Minarik TA, Schoenfuss HL, Edmiston PL (2019) Correlating effluent concentrations and bench-scale experiments to assess the transformation of endocrine active compounds in wastewater by UV or chlorination disinfection. Chemosphere 226:565–575. https://doi.org/10.1016/j.chemosphere.2019.03.145 Molu ZB, Yurdakoç K (2010) Preparation and characterization of aluminum pillared K10 and KSF for adsorption of trimethoprim. Microporous Mesoporous Mater 127:50–60. https://doi. org/10.1016/j.micromeso.2009.06.027 Mon M, Bruno R, Ferrando-Soria J, Armentano D, Pardo E (2018) Metal-organic framework technologies for water remediation: towards a sustainable ecosystem. J Mater Chem A 6:4912–4947. https://doi.org/10.1039/C8TA00264A Monteoliva-García A, Martín-Pascual J, Muñío MM, Poyatos JM (2019) Removal of a pharmaceutical mix from urban wastewater coupling membrane bioreactor with advanced oxidation processes. J Environ Eng 145. https://doi.org/10.1061/(ASCE)EE.1943-­7870.0001571 Moon RJ, Martini A, Nairn J, Simonsen J, Youngblood J (2011) Cellulose nanomaterials review: structure, properties and nanocomposites. Chem Soc Rev 40:3941–3994 Moreira NFF, Sousa JM, Macedo G, Ribeiro AR, Barreiros L, Pedrosa M, Faria JL, Pereira MFR, Castro-Silva S, Segundo MA, Manaia CM, Nunes OC, Silva AMT (2016) Photocatalytic ozonation of urban wastewater and surface water using immobilized TiO2 with LEDs: micropollutants, antibiotic resistance genes and estrogenic activity. Water Res 94:10–22. https://doi. org/10.1016/j.watres.2016.02.003 Moreira FC, Boaventura RAR, Brillas E, Vilar VJP (2017) Electrochemical advanced oxidation processes: a review on their application to synthetic and real wastewaters. Appl Catal B Environ 202:217–261. https://doi.org/10.1016/j.apcatb.2016.08.037 Morin-Crini N, Crini G (2013) Environmental applications of water-insoluble β-cyclodextrin-­ epichlorohydrin polymers. Progr Polym Sci 38:344–368. https://doi.org/10.1016/j. progpolymsci.2012.06.005 Morin-Crini N, Crini G (eds) (2017) Eaux industrielles contaminées (in French). PUFC, Besançon, 513 p Morin-Crini N, Winterton P, Fourmentin S, Wilson LD, Fenyvesi É, Crini G (2018) Water-insoluble β-cyclodextrin-epichlorohydrin polymers for removal of pollutants from aqueous solutions by sorption processes using batch studies: a review of inclusion mechanisms. Progr Polym Sci 78:1–23. https://doi.org/10.1016/j.progpolymsci.2017.07.004 Morin-Crini N, Lichtfouse E, Torri G, Crini G (2019) Applications of chitosan in food, pharmaceuticals, medicine, cosmetics, agriculture, textiles, pulp and paper, biotechnology,

94

N. Morin-Crini et al.

and environmental chemistry. Environ Chem Lett 17:1667–1692. https://doi.org/10.1007/ s10311-­019-­00904-­x Moussa D, Doubla A, Kamgang-Youbi G, Brisset JL (2007) Postdischarge long life reactive intermediates involved in the plasma chemical degradation of an azoic dye. IEEE Trans Plasma Sci 35:444–453. https://doi.org/10.1109/TPS.2007.892578 Murray A, Örmeci B (2012) Application of molecularly imprinted and non-imprinted polymers for removal of emerging contaminants in water and wastewater treatment: a review. Environ Sci Pollut Res 19:3820–3830. https://doi.org/10.1007/s11356-­012-­1119-­2 Nagy ZM, Molnár M, Fekete-Kertész I, Molnár Perl I, Fenyvesi E, Gruiz K (2014) Removal of emerging micropollutants from water using cyclodextrin. Sci Total Environ 485-486:711–719. https://doi.org/10.1016/j.scitotenv.2014.04.003 Nakamura H, Kuruto-Niwa R, Uchida M, Terao Y (2007) Formation of chlorinated estrones via hypochlorous disinfection of wastewater effluent containing estrone. Chemosphere 66:1441–1448. https://doi.org/10.1016/j.chemosphere.2006.09.011 Nakashima Y, Ikawa S, Tani A, Kitano K (2016) Ion-exchange chromatographic analysis of peroxynitric acid. J Chromatogr A 1431:89–93. https://doi.org/10.1016/j.chroma.2015.12.054 Nam S, Choi D, Kim S, Her N, Zoh K (2014) Adsorption characteristics of selected hydrophilic and hydrophobic micropollutants in water using activated carbon. J Hazard Mater 270:144–152 Nasuhoglu D, Isazadeh S, Westlund P, Neamatallah S, Yargeau V (2018) Chemical, microbial and toxicological assessment of wastewater treatment plant effluents during disinfection by ozonation. Chem Eng J 346:466–476. https://doi.org/10.1016/j.cej.2018.04.037 Navalon S, Alvaro M, Garcia H (2008) Reaction of chlorine dioxide with emergent water pollutants: product study of the reaction of three β-lactam antibiotics with ClO2. Water Res 42:1935–1942 Neamtu M, Frimmel F (2006) Photodegradation of endocrine disrupting chemical nonylphenol by simulated solar UV-irradiation. Sci Total Environ 369:295–306. https://doi.org/10.1016/j. scitotenv.2006.05.002 Nemoto J, Saito T, Isogai A (2015) Simple freeze-drying procedure for producing nanocellulose aerogel-containing, high performance air filters. ACS Appl Mater Interfaces 7:19809–19815 Nika MC, Bletsou A, Koumaki E, Noutsopoulos C, Mamais D, Stasinakis AS, Thomaidis NS (2016) Chlorination of benzothiazoles and benzotriazoles and transformation products identification by LC-HR-MS/MS.  J Hazard Mater 323:400–413. https://doi.org/10.1016/j. jhazmat.2016.03.035 Norvill ZN, Toledo-Cervantes A, Blanco S, Shilton A, Guieysse B, Muñoz R (2017) Photodegradation and sorption govern tetracycline removal during wastewater treatment in algal ponds. Bioresour Technol 232:35–43. https://doi.org/10.1016/j.biortech.2017.02.011 Noutsopoulos C, Mamais D, Thomaidis N, Koumaki E, Nika M, Bletsou A, Stasinakis A (2013a) Removal of emerging pollutants through wastewater disinfection. Proceedings of the 13th international conference on environmental science and technology. Athens, Greece, 5–7 September 2013 Noutsopoulos C, Mamais D, Bouras A, Kokkinidou D, Samaras V, Antoniou K, Gioldasi M (2013b) The role of activated carbon and disinfection on the removal of endocrine disrupting chemicals and non-steroidal anti-inflammatory drugs from wastewater. Environ Technol 35:698–708. https://doi.org/10.1080/09593330.2013.846923 Noutsopoulos C, Mamais D, Samaras V, Bouras T, Marneri M, Antoniou K (2013c) Effect of wastewater chlorination on endocrine disruptor removal. Water Sci Technol 67:1551–1556. https://doi.org/10.2166/wst.2013.025 Noutsopoulos C, Koumaki E, Mamais D, Nika MC, Bletsou A, Thomaidis N (2015) Removal of endocrine disruptors and non-steroidal anti-inflammatory drugs through wastewater chlorination: the effect of pH, total suspended solids and humic acids and identification of degradation by-­ products. Chemosphere 119:S109–S114. https://doi.org/10.1016/j.chemosphere.2014.04.107 Oishi K, Moriuchi A (2010) Removal of dissolved estrogen in sewage efuents by beta-cyclodextrin polymer. Sci Total Environ 409:112–115. https://doi.org/10.1016/j.scitotenv.2010.09.031

1  Remediation of Emerging Contaminants

95

Ojajuni O, Saroj D, Cavalli G (2015) Removal of organic micropollutants using membrane-­ assisted processes: a review of recent progress. Environ Technol Rev 4:17–37. https://doi.org/1 0.1080/21622515.2015.1036788 Oliveira MF, Souza VM, da Silva MGC, Vieira MGA (2018) Fixed-bed adsorption of caffeine onto thermally modified Verdelodo bentonite. Ind Eng Chem Res 57:17480–17487. https://doi. org/10.1021/acs.iecr.8b03734 Oliveira MF, da Silva MGC, Vieira MGA (2019) Equilibrium and kinetic studies of caffeine adsorption from aqueous solutions on thermally modified Verde-lodo bentonite. Appl Clay Sci 168:366–373. https://doi.org/10.1016/j.clay.2018.12.011 Omtvedt LA, Dalheim MØ, Nielsen TT, Larsen KL, Strand BL, Aachmann FL (2019) Efficient grafting of cyclodextrin to alginate and performance of the hydrogel for release of model drug. Sci Rep 9:9325. https://doi.org/10.1038/s41598-­019-­45761-­4 Ono R, Oda (2002) Dynamics and density estimation of hydroxyl radicals in a pulsed corona discharge. J Phys D 35:2133. https://doi.org/10.1088/0022-­3727/35/17/309 Orprecio R, Evans CH (2003) Polymer-immobilized cyclodextrin trapping of model organic pollutants in flowing water streams. J Appl Polym Sci 90:2103–2110. https://doi.org/10.1002/ app.12818 Osei AR, Konate Y, Abagale FK (2019) Pollutant removal and growth dynamics of macrophyte species for faecal sludge treatment with constructed wetland technology. Water Sci Technol 80:1145–1154. https://doi.org/10.2166/wst.2019.354 Oturan MA, Aaron JJ (2014) Advanced oxidation processes in water/wastewater treatment: principles and applications: a review. Critical Rev Environ Sci Technol 44:2577–2641. https:// doi-­org.inee.bib.cnrs.fr/10.1080/10643389.2013.829765 Pablos C, Marugán J, van Grieken R, Serrano E (2013) Emerging micropollutant oxidation during disinfection processes using UV-C, UV-C/H2O2, UV-A/TiO2 and UV-A/TiO2/H2O2. Water Res 47:1237–1245. https://doi.org/10.1016/j.watres.2012.11.041 Panorel I, Preis S, Kornev I, Hatakka H, Louhi-Kultanen M (2013) Oxidation of aqueous pharmaceuticals by pulsed corona discharge. Environ Technol 34:923–930. https://doi.org/10.108 0/09593330.2012.722691 Park JY, Kostyuk PV, Han SB, Kim JS, Vu CN, Lee HW (2006) Study on optical emission analysis of AC air-water discharges under He, Ar and N2 environments. J Phys D Appl Phys 39:3805e3813. https://doi.org/10.1088/0022-­3727/39/17/015 Park DP, Davis K, Gilani S, Alonzo C, Dobrynin D, Friedman G, Fridman A, Rabinovich A, Fridman G (2013) Reactive nitrogen species produced in water by non-equilibrium plasma increase plant growth rate and nutritional yield. Curr Appl Phys 13:S19–S29. https://doi. org/10.1016/j.cap.2012.12.019 Parlade E, Hom-Diaz A, Blanquez P, Martínez-Alonso M, Vicent T, Gaju N (2018) Effect of cultivation conditions on β-estradiol removal in laboratory and pilot-plant photobioreactors by an algal-bacterial consortium treating urban wastewater. Water Res 137:86–96. https://doi. org/10.1016/j.watres.2018.02.060 Parolo ME, Savini M, Valles J, Baschini M, Avena M (2008) Tetracycline adsorption on montmorillonite: effects of pH and ionic strength. Appl Clay Sci 40:179–186. https://doi.org/10.1016/j. clay.2007.08.003 Parvulescu VI, Magureanu M, Lukes P (2012) Plasma chemistry and catalysis in gases and liquids. Wiley-VCH, Weinheim. https://doi.org/10.1002/9783527649525 Patel M, Kumar R, Kishor K, Mlsna T, Pittman CU Jr, Mohan D (2019) Pharmaceuticals of emerging concern in aquatic systems: chemistry, occurrence, effects, and removal methods. Chem Rev 119:3510–3673. https://doi.org/10.1021/acs.chemrev.8b00299 Pathak N, Tran VH, Phuntsho S, Shon HY (2020) Chapter 10: Membrane reactors for the removal of organic micro-pollutants. In: Varjani S, Pandey A, Tyagi RD, Ngo HH, Larroche C (eds) Current developments in biotechnology and bioengineering. Emerging organic micro-­ pollutants. Elsevier, Amsterdam, pp 231–252

96

N. Morin-Crini et al.

Patiño Y, DíazE OS, Gallegos-Suarez E, Guerrero-Ruiz A, Rodríguez-Ramos I (2015) Adsorption of emerging pollutants on functionalized multiwall carbon nanotubes. Chemosphere 136:174–180. https://doi.org/10.1016/j.chemosphere.2015.04.089 Paucar NE, Kim I, Tanaka H, Sato C (2018) Ozone treatment process for the removal of pharmaceuticals and personal care products in wastewater. Ozone Sci Eng 41:3–16. https://doi.org/1 0.1080/01919512.2018.1482456 Pereira VJ, Linden KG, Weinberg HS (2007a) Evaluation of UV irradiation for photolytic and oxidative degradation of pharmaceutical compounds in water. Water Res 41:4413–4423. https:// doi.org/10.1016/j.watres.2007.05.056 Pereira VJ, Weinberg HS, Linden KG, Singer PC (2007b) UV degradation of pharmaceutical compounds in surface water via direct and indirect photolysis at 254  nm. Environ Sci Technol 41:1682–1688. https://doi.org/10.1021/es061491b Petrie B, Barden R, Kasprzk-Hordern B (2015) A review on emerging contaminants in wastewaters and the environment: current knowledge, understudied areas and recommendations for future monitoring. Water Res 72:3–27. https://doi.org/10.1016/j.watres.2014.08.053 Petrovic M, Diaz A, Ventura F, Barceló D (2003) Occurrence and removal of estrogenic short-­ chain ethoxy nonylphenolic compounds and their halogenated derivatives during drinking water production. Environ Sci Technol 37:4442–4448 Phan HV, Hai FI, McDonald JA, Khan SJ, Van De Merwe JP, Leusch FD, Zhang R, Price WE, Broeckmann A, Nghiem LD (2015) Impact of hazardous events on the removal of nutrients and trace organic contaminants by an anoxic-aerobic membrane bioreactor receiving real wastewater. Bioresour Technol 192:192–201 Pichon V, Chapuis-Hugon F (2008) Role of molecularly imprinted polymers for selective determination of environmental pollutants – a review. Anal Chim Acta 622:48–61 Picos-Corrales LA, Sarmiento-Sánchez JI, Ruelas-Leyva JP, Crini G, Hermosillo-Ochoa E, Gutierrez-Montes JA (2020) Environment-friendly approach toward the treatment of raw agricultural wastewater and river water via flocculation using chitosan and bean straw flour as bioflocculants. ACS Omega 5:3943–3951. https://doi.org/10.1021/acsomega.9b03419 Pinkston KE, Sedlak DL (2004) Transformation of aromatic ether- and amine-containing pharmaceuticals during chlorine disinfection. Environ Sci Technol 38:4019–4025. https://doi. org/10.1021/es035368l Polatti S, Angioi S, Gianotti V, Gosetti F, Gennaro M (2006) Sorption of pesticides on kaolinite and montmorillonite as a function of hydrophilicity. J Environ Sci Health B 41:333–344 Polo-López MI, Castro-Alférez M, Oller I, Fernández-Ibáñez P (2014) Assessment of solar photo-­ Fenton, photocatalysis, and H2O2 for removal of phytopathogen fungi spores in synthetic and real effluents of urban wastewater. Chem Eng J 257:122–130. https://doi.org/10.1016/j. cej.2014.07.016 Portinho R, Zanella O, Féris LA (2017) Grape stalk application for caffeine removal through adsorption. J Environ Manag 202:178–187. https://doi.org/10.1016/j.jenvman.2017.07.033 Preis S, Panorel IC, Kornev I, Hatakka H, Kallas J (2013) Pulsed corona discharge: the role of ozone and hydroxyl radical in aqueous pollutants oxidation. Water Sci Technol 68:1536–1542. https://doi.org/10.2166/wst.2013.399 Priac A, Morin-Crini N, Druart C, Gavoille S, Bradu C, Lagarrigue C, Torri G, Winterton P, Crini G (2017) Alkylphenol and alkylphenol polyethoxylates in water and wastewater: a review of options for their elimination. Arab J Chem 10:S3749–S3773. https://doi.org/10.1016/j. arabjc.2014.05.011 Putra EK, Pranowo R, Sunarso J, Indraswati N, Ismadji S (2009) Performance of activated carbon and bentonite for adsorption of amoxicillin from wastewater: mechanisms, isotherms and kinetics. Water Res 43:2419–2430. https://doi.org/10.1016/j.watres.2009.02.039 Putro JN, Kurniawan A, Ismadji S, Ju YH (2017) Nanocellulose based biosorbents for wastewater treatment: study of isotherm, kinetic, thermodynamic and reusability. Environ Nanotechnol Monitoring Manage 8:134–149. https://doi.org/10.1016/j.enmm.2017.07.002

1  Remediation of Emerging Contaminants

97

Qin S, Su L, Wang P, Gao Y (2015) Rapid and selective extraction of multiple sulfonamides from aqueous samples based on Fe3O4-chitosan molecularly imprinted polymers. Anal Methods 7:874–8713 Qu RJ, Xu BG, Meng LJ, Wang LS, Wang ZY (2015) Ozonation of indigo enhanced by carboxylated carbon nanotubes: performance optimization, degradation products, reaction mechanism and toxicity evaluation. Water Res 68:316–327. https://doi.org/10.1016/j.watres.2014.10.017 Quintana JB, Rodil R, Lopez-Mahia P, Muniategui-Lorenzo S, Prada-Rodriguez P (2010) Investigating the chlorination of acidic pharmaceuticals and by-product formation aided by an experimental design methodology. Water Res 44:243–255 Quintana JB, Rodil R, Cela R (2012) Reaction of β-blockers and β-agonist pharmaceuticals with aqueous chlorine. Investigation of kinetics and by-products by liquid chromatography quadrupole time-of-flight mass spectrometry. Anal Bioanal Chem 403:2385–2395. https://doi. org/10.1007/s00216-­011-­5707-­7 Racar M, Dolar D, Karadakić K, Čavarović N, Glumac N, Ašperger D, Košutić K (2020) Challenges of municipal wastewater reclamation for irrigation by MBR and NF/RO: physicochemical and microbiological parameters, and emerging contaminants. Sci Total Environ 722:137959. https://doi.org/10.1016/j.scitotenv.2020.137959 Radha KV, Sirisha K (2018) Chapter 11: Electrochemical oxidation processes. In: Advanced oxidation processes for waste water treatment, pp  359–373. https://doi.org/10.1016/ B978-­0-­12-­810499-­6.00011-­5 Radjenovic J, Petrovic M, Barceló D (2007) Analysis of pharmaceuticals in wastewater and removal using a membrane bioreactor. Anal Bioanal Chem 387:1365–1377. https://doi. org/10.1007/s00216-­006-­0883-­6 Radich JG, Krenselewski AL, Zhu JD, Kamat PV (2014) Is graphene a stable platform for photocatalysis? Mineralization of reduced graphene oxide with UV-irradiated TiO2 nanoparticles. Chem Mater 26:4662–4668. https://doi.org/10.1021/cm5026552 Real FJ, Benitez FJ, Acero JL, Sagasti JJP, Casas F (2009) Kinetics of the chemical oxidation of the pharmaceuticals primidone, ketoprofen, and diatrizoate in ultrapure and natural waters. Ind Eng Chem Research 48:3380–3388 Rehman MSU, Munir M, Asfaq M, Rashis N, Nazar MF, Danish M, Han JI (2013) Adsorption of Brilliant Green dye from aqueous solution onto red clay. Chem Eng J 228:54–62. https://doi. org/10.1016/j.cej.2013.04.094 Ren YM, Zhang HY, An HG, Zhao Y, Feng J, Xue L, Luan TZ, Fan ZG (2018) Catalytic ozonation of di-n-butyl phthalate degradation using manganese ferrite/reduced graphene oxide nanofiber as catalyst in the water. J Colloid Interface Sci 526:347–355. https://doi.org/10.1016/j. jcis.2018.04.073 Renew JE, Huang CH (2004) Simultaneous determination of flurorquinolone, sulfonamide, and trimethoprim antibiotics in wastewater using tandem solid phase extraction and liquid chromatography-electrospray mass spectrometry. J Chromatogr A 1042:113–121. https://doi. org/10.1016/j.chroma.2004.05.056 Restivo J, Órfão JJM, Armenise S, Garcia-Bordejé E, Pereira MFR (2012) Catalytic ozonation of metolachlor under continuous operation using nanocarbon materials grown on a ceramic monolith. J Hazard Mater 239-240:249–256 Restivo J, Órfão JJM, Pereira MFR, Garcia-bordejé E, Roche P, Bourdin D, Houssais B, Coste M, Derrouiche S (2013) Catalytic ozonation of organic micropollutants using carbon nanofibers supported on monoliths. Chem Eng J 230:115–123. https://doi.org/10.1016/j.cej.2013.06.064 Restivo J, Garcia-bordejé E, Órfão JJM, Pereira MFR (2016) Carbon nanofibers doped with nitrogen for the continuous catalytic ozonation of organic pollutants. Chem Eng J 293:102–111. https://doi.org/10.1016/j.cej.2016.02.055 Reungoat J, Escher BI, Macova M, Argaud FX, Gernjak W, Keller J (2012) Ozonation and biological activated carbon filtration of wastewater treatment plant effluents. Water Res 46:863–872

98

N. Morin-Crini et al.

Ribeiro AR, Nunes OC, Pereira MFR, Silva AMT (2015) An overview on the advanced oxidation processes applied for the treatment of water pollutants defined in the recently launched Directive 2013/39/EU. Environ Int 75:33–51. https://doi.org/10.1016/j.envint.2014.10.027 Riva V, Mapelli F, Syranidou E, Crotti E, Choukrallah R, Kalogerakis N, Borin S (2019) Root bacteria recruited by Phragmites australis in constructed wetlands have the potential to enhance azo-­ dye phytodepuration. Microorganisms 7:384. https://doi.org/10.3390/microorganisms7100384 Riva V, Riva F, Vergani L, Crotti E, Borin S, Mapelli F (2020) Microbial assisted phytodepuration for water reclamation: environmental benefits and threats. Chemosphere 241:124843. https:// doi.org/10.1016/j.chemosphere.2019.124843 Rivera-Utrilla J, Gómez-Pacheco CV, Sánchez-Polo M, López-Peñalver JJ, Ocampo-Pérez R (2013) Tetracycline removal from water by adsorption/bioadsorption on activated carbons and sludgederived adsorbents. J Environ Manag 131:16–24 Rizzi V, Romanazzi F, Gubitosa J, Fini P, Romita R, Agostiano A, Petrella A, Cosma P (2019) Chitosan film as eco-friendly and recyclable bio-adsorbent to remove/recover diclofenac, ketoprofen, and their mixture from wastewater. Biomol Ther 9:571. https://doi.org/10.3390/ biom9100571 Roca Jalil ME, Vieira R, Azevedo D, Baschini M, Sapag K (2013) Improvement in the adsorption of thiabendazole by using aluminum pillared clays. Appl Clay Sci 71:55–63. https://doi. org/10.1016/j.clay.2012.11.005 Roca Jalil ME, Baschini M, Sapag K (2015) Influence of pH and antibiotic solubility on the removal of ciprofloxacin from aqueous media using montmorillonite. Appl Clay Sci 114:69–76. https:// doi.org/10.1016/j.clay.2015.05.010 Rocha RP, Gonçalves AG, Pastrana-Martínez LM, Bordoni BC, Soares OSGP, Órfão JJM, Faria JL, Figueiredo JL, Silva AMT, Pereira MFR (2015) Nitrogen-doped graphene-based materials for advanced oxidation processes. Catal Today 249:192–198. https://doi.org/10.1016/j. cattod.2014.10.046 Rodriguez-Narvaez OM, Peralta-Hernandez JM, Goonetilleke A, Bandala ER (2017) Treatment technologies for emerging contaminants in water: a review. Chem Eng J 323:361–380. https:// doi.org/10.1016/j.cej.2017.04.106 Rodríguez-Rodríguez CE, Cambronero JC, Beita-Sandi W, Duran JE (2019) Chapter 7: Removal of emerging pollutants by fungi. In: Fungal bioremediation: fundamentals and applications. CRC Press, Taylor & Francis Group, pp 186–237. https://doi.org/10.1201/9781315205984-­7 Rojas S, Horcajada P (2020) Metal-organic frameworks for the removal of emerging organic contaminants in water. Chem Rev. https://doi.org/10.1021/acs.chemrev.9b00797 Romo A, Penas FJ, Sevillano X, Isasi JR (2006) Application of factorial experimental design to the study of the suspension polymerization of beta-cyclodextrin and epichlorohydrin. J Appl Polym Sci 100:3393–3402. https://doi.org/10.1002/app.23778 Rong S, Sun Y (2014) Wetted-wall corona discharge induced degradation of sulfadiazine antibiotics in aqueous solution. J Chem Technol Biotechnol 89:1351–1359. https://doi.org/10.1002/ jctb.4211 Rong SP, Sun YB, Zhao ZH (2014) Degradation of sulfadiazine antibiotics by water falling film dielectric barrier discharge. Chinese Chem Lett 25:187–192. https://doi.org/10.1016/j. cclet.2013.11.003 Rosal R, Rodrıguez A, Perdigón-Melón JA, Petre A, Garcıa-Calvo E, Gomez MJ, Aguera A, Fernandez-Alba A (2010a) Occurrence of emerging pollutants in urban wastewater and their removal through biological treatment followed by ozonation. Water Res 44:578–588. https:// doi.org/10.1016/j.watres.2009.07.004 Rosal R, Gonzalo M, Rodríguez A, García-Calvo E (2010b) Catalytic ozonation of fenofibric acid over alumina-supported manganese oxide. J Hazard Mater 183:271–278 Rosario-Ortiz FL, Wert EC, Snyder SA (2010) Evaluation of UV/H2O2 treatment for the oxidation of pharmaceuticals in wastewater. Water Res 44:1440–1448. https://doi.org/10.1016/j. watres.2009.10.031

1  Remediation of Emerging Contaminants

99

Rosenfeldt EJ, Linden KG (2004) Degradation of endocrine disrupting chemicals bisphenol A, ethinyl estradiol, and estradiol during UV photolysis and advanced oxidation. Environ Sci Technol 38:5476–5483. https://doi.org/10.1021/es035413p Rossner A, Snyder SA, Knappe DR (2009) Removal of emerging contaminants of concern by alternative adsorbents. J Water Res 43:3787–3796. https://doi.org/10.1016/j.watres.2009.06.009 Sakairi N, Nishi N, Tokura S (1999) Cyclodextrin-linked chitosan: synthesis and inclusion complexation abilities. ACS Symp Ser 737:68–84. https://doi.org/10.1021/bk-­1999-­0737.ch005 Salgado R, Marques R, Noronha JP, Carvalho G, Oehmen A, Reis MAM (2011) Assessing the removal of pharmaceuticals and personal care products in a full-scale activated sludge plant. Environ Sci Pollut Res. https://doi.org/10.1007/s11356-­011-­0693-­z Salgado R, Marques R, Noronha JP, Carvalho G, Oehmen A, Reis MAM (2012) Assessing the removal of pharmaceuticals and personal care products in a full-scale activated sludge plant. Environ Sci Pollut Res Int 19:1818–1827. https://doi.org/10.1007/s11356-­011-­0693-­z Salgado R, Pereira VJ, Carvalho G, Soeiro R, Gaffney V, Almeida C, Vale Cardoso V, Ferreira E, Benoliel MJ, Ternes TA, Oehmen A, Reis MAM, Noronha JP (2013) Photodegradation kinetics and transformation products of ketoprofen, diclofenac and atenolol in pure water and treated wastewater. J Hazard Mater 244-245:516–527 Sampaio MJ, Bacsa RR, Benyounes A, Axet R, Serp P, Silva CG, Silva AMT, Faria JL (2015) Synergistic effect between carbon nanomaterials and ZnO for photocatalytic water decontamination. J Catalysis 331:172–180. https://doi.org/10.1016/j.jcat.2015.08.011 Sanford S, Singh KS, Chaini S, LeClair G (2012) Study of natural adsorbent chitosan and derivatives for the removal of caffeine from water. Water Quality Res J 47:80–90. https://doi. org/10.2166/wqrjc.2012.021 Sanguanpak S, Chiemchaisri C, Chiemchaisri W, Yamamoto K (2015) Effects of mixed liquor pH on membrane fouling and micro-pollutant removals in membrane bioreactors for municipal landfill leachate treatment. Water Sci Technol 72:770–778 Santos-Silva TG, Montagner CC, Martinez CBR (2018) Evaluation of caffeine effects on biochemical and genotoxic biomarkers in the neotropical freshwater teleost Prochilodus lineatus. Environ Toxicol Pharmacol 58:237–242. https://doi.org/10.1016/j.etap.2018.02.002 Sarangapani C, Ziuzina D, Behan P, Boehm D, Gilmore BF, Cullen PJ, Bourke P (2019) Degradation kinetics of cold plasma-treated antibiotics and their antimicrobial activity. Sci Rep 9:3955. https://doi.org/10.1038/s41598-­019-­40352-­9 Sauvêtre A, Schröder P (2015) Uptake of carbamazepine by rhizomes and endophytic bacteria of Phragmites australis. Frontiers Plant Sci 6:83. https://doi.org/10.3389/fpls.2015.00083 Schaar H, Clara M, Gans O, Kreuzinger N (2010) Micropollutant removal during biological wastewater treatment and a subsequent ozonation step. Environ Pollut 158:1399–1404. https://doi. org/10.1016/j.envpol.2009.12.038 Secondes MFN, Naddeo V, Belgiorno V, Ballesteros F Jr (2014) Removal of emerging contaminants by simultaneous applications of membrane ultrafiltration, activated carbon adsorption, and ultrasound irradiation. J Hazard Mater 264:342–349. https://doi.org/10.1016/j. jhazmat.2013.11.039 Serna-Galvis EA, Silva-Agredo J, Giraldo-Aguirre AL, Flórez-Acosta OA, Torres-Palma RA (2016) High frequency ultrasound as a selective advanced oxidation process to remove penicillinic antibiotics and eliminate its antimicrobial activity from water. Ultrasonics Sonochem 31:276–283. https://doi.org/10.1016/j.ultsonch.2016.01.007 Serna-Galvis EA, Botero-Coy AM, Martínez-Pachón D, Moncayo-Lasso A, Ibáñez M, Hernández F, Torres-Palma RA (2019) Degradation of seventeen contaminants of emerging concern in municipal wastewater effluents by sonochemical advanced oxidation processes. Water Res 154:349–360. https://doi.org/10.1016/j.watres.2019.01.045 Shak KPY, Pang YL, Mah SK (2018) Nanocellulose: recent advances and its prospects in environmental remediation. Beilstein J Nanotechnol 9:2479–2498. https://doi.org/10.3762/ bjnano.9.232

100

N. Morin-Crini et al.

Sharma VK (2008) Oxidative transformations of environmental pharmaceuticals by Cl2, ClO2, O3 and Fe(VI): kinetics assessment. Chemosphere 73:1379–1386. https://doi.org/10.1016/j. chemosphere.2008.08.033 Shen X, Xu C, Ye L (2013) Molecularly imprinted polymers for clean water: analysis and purification. Ind Eng Chem Res 52:13890–13899 Shiono T, Yamamoto K, Yotsumoto Y, Yoshida A (2017) Caffeine adsorption of montmorillonite in coffee extracts. Biosci Biotechnol Biochem 81:1591–1597. https://doi.org/10.1080/0916845 1.2017.1340087 Shreve MJ, Brockman A, Hartleb M, Prebihalo S, Dorman FL, Brennan RA (2016) The white-rot fungus Trametes versicolor reduces the estrogenic activity of a mixture of emerging contaminants in wastewater treatment plant effluent. Int Biodeterior Biodegrad 109:132–140. https:// doi.org/10.1016/j.ibiod.2016.01.018 Silva A, Delerue-Matos C, Figueiredo SA, Freitas OM (2019) The use of algae and fungi for removal of pharmaceuticals by bioremediation and biosorption processes: a review. Water 11:1555. https://doi.org/10.3390/w11081555 Simazaki D, Fujiwara J, Manabe S, Matsuda M, Asami M, Kunikane S (2008) Removal of selected pharmaceuticals by chlorination, coagulation-sedimentation and powdered activated carbon treatment. Water Scie Technol 58:1129–1135 Singh RN, Sharma S (2012) Development of suitable photobioreactor for algae production -a review. Renew Sust Energ Rev 16:2347–2353. https://doi.org/10.1016/j.rser.2012.01.026 Singh RK, Philip L, Ramanujam S (2016) Rapid removal of carbofuran from aqueous solution by pulsed corona discharge treatment: kinetic study, oxidative, reductive degradation pathway, and toxicity assay. Ind Eng Chem Res 55:7201–7209. https://doi.org/10.1021/acs.iecr.6b01191 Singh RK, Philip L, Ramanujam S (2017) Removal of 2,4-dichlorophenoxyacetic acid in aqueous solution by pulsed corona discharge treatment: effect of different water constituents, degradation pathway and toxicity assay. Chemosphere 184:207–214. https://doi.org/10.1016/j. chemosphere.2017.05.134 Singhal N, Perez-Garcia O (2016) Degrading organic micropollutants: the next challenge in the evolution of biological wastewater treatment processes. Front Environ Sci 4:36. https://doi. org/10.3389/fenvs.2016.00036 Slade R, Bauen A (2013) Micro-algae cultivation for biofuels: cost, energy balance, environmental impacts and future prospects. Biomass Bioenergy 53:29–38. https://doi.org/10.1016/j. biombioe.2012.12.019 Smith JB, Adams I, Ji H-F (2018) Mechanism of ampicillin degradation by non-thermal plasma treatment with FE-DBD. Plasma 1:1–11. https://doi.org/10.3390/plasma1010001 Snyder SA, Wert EC, Rexing DJ, Zegers RE, Drury DD (2006) Ozone oxidation of endocrine disruptors and pharmaceuticals in surface water and wastewater. Ozone Sci Eng 28:445–460. https://doi-­org.inee.bib.cnrs.fr/10.1080/01919510601039726 Snyder SA, Adham S, Redding AM, Cannon FS, DeCarolis J, Oppenheimer J, Wert EC, Yoon Y (2007) Role of membranes and activated carbon in the removal of endocrine disruptors and pharmaceuticals. Des 202:156–181 Soares OSGP, Gonçalves AG, Delgado JJ, Órfão JJM, Pereira MFR (2015) Modification of carbon nanotubes by ball-milling to be used as ozonation catalysts. Catal Today 249:199–203 Sole M, Lopez de Alda MJ, Castillo M, Porte C, Ladegaard-Pedersen K, Barceló D (2000) Estrogenicity determination in sewage treatment plants and surface waters from the Catalonian area (NE Spain). Environ Sci Technol 34:5076–5083 Song Z, Wang MG, Wang Z, Wang YF, Li RY, Zhang YT, Liu C, Liu Y, Xu BB, Qi F (2019a) Insights into heteroatom-doped graphene for catalytic ozonation: active centers, reactive oxygen species evolution, and catalytic mechanism. Environ Sci Technol 53:5337–4538 Song Z, Zhanf YT, Liu C, Xu BB, Qi F, Yuan DG, Pu SG (2019b) Insight into OH and O2-formation in heterogeneous catalytic ozonation by delocalized electrons and surface oxygen-containing functional groups in layered-structure nanocarbons. Cheml Eng J 357:655–666

1  Remediation of Emerging Contaminants

101

Soufan M, Deborde M, Legube B (2012) Aqueous chlorination of diclofenac: kinetic study and transformation products identification. Water Res 46:3377–3386. https://doi.org/10.1016/j. watres.2012.03.056 Spina F, Anastasi A, Prigione V, Tigini V, Varese GC (2012) Biological treatment of industrial wastewaters: a fungal approach. Chem Eng Trans 27:175–180. https://doi.org/10.3303/ CET1227030 Stackelberg PE, Gibs J, Furlong ET, Meyer MT, Zaugg SD, Lippincott RL (2007) Efficiency of conventional drinking-water-treatment processes in removal of pharmaceuticals and other organic compounds. Sci Total Environ 377:255–272. https://doi.org/10.1016/j.scitotenv.2007.01.095 Stadlmair LF, Letzel T, Graßmann J (2018) Monitoring enzymatic degradation of emerging contaminants using a chip-based robotic nano-ESI-MS tool. Anal Bioanal Chem 410:7–32. https:// doi.org/10.1007/s00216-­017-­0729-­4 Stalter D, Magdeburg A, Oehlmann J (2010a) Comparative toxicity assessment of ozone and activated carbon treated sewage effluents using an in vivo test battery. Water Res 44:2610–2620. https://doi.org/10.1016/j.watres.2010.01.023 Stalter D, Magdeburg A, Weil M, Knacker T, Oehlmann J (2010b) Toxication or detoxication? In vivo toxicity assessment of ozonation as advanced wastewater treatment with the rainbow trout. Water Res 44:439–448. https://doi.org/10.1016/j.watres.2009.07.025 Stalter D, Magdeburg A, Wagner M, Oehlmann J (2011) Ozonation and activated carbon treatment of sewage effluents: removal of endocrine activity and cytotoxicity. Water Res 45:1015–1024. https://doi.org/10.1016/j.watres.2010.10.008 Stará Z, Krčma F, Nejezchleb M, Skalný JD (2009) Organic dye decomposition by DC diaphragm discharge in water: effect of solution properties on dye removal. Des 239:283–294. https://doi. org/10.1016/j.desal.2008.03.025 Stofela SKF, de Andrade JR, Vieira MGA (2017) Adsorption of benzene, toluene, and xylene (BTX) from binary aqueous solutions using commercial organoclay. Can J Chem Eng 95:1034–1044. https://doi.org/10.1002/cjce.22748 Suárez S, Reif R, Lema JM, Omil F (2012) Mass balance of pharmaceutical and personal care products in a pilot-scale single-sludge system: influence of T SRT and recirculation ratio. Chemosphere 89:164–171 Sui M, Xing S, Sheng L, Huang S, Guo H (2012) Heterogeneous catalytic ozonation of ciprofloxacin in water with carbon nanotube supported manganese oxides as catalyst. J Hazard Mater 227-228:227–236 Sugiarto AT, Ito S, Ohshima SM, Skalny JD (2003) Oxidative decoloration of dyes by pulsed discharge plasma in water. J Electrost 58:135–145. https://doi.org/10.1016/S0304-­3886 (02)00203-­6 Sun J, Wang J, Zhang R, Wei D, Long Q, Huang Y, Xie X, Li A (2017) Comparison of different advanced treatment processes in removing endocrine disruption effects from municipal wastewater secondary effluent. Chemosphere 168:1–9. https://doi.org/10.1016/j. chemosphere.2016.10.031 Syam Babu D, Srivastava V, Nidheesh PV, Suresh Kumar M (2019) Detoxification of water and wastewater by advanced oxidation processes. Sci Total Environ 696:133961. https://doi. org/10.1016/j.scitotenv.2019.133961 Syranidou E, Christofilopoulos S, Politi M, Weyens N, Venieri D, Vangronsveld J, Kalogerakis N (2017) Bisphenol-A removal by the halophyte Juncus acutus in a phytoremediation pilot, characterization and potential role of the endophytic community. J Hazard Mater 323:350–358. https://doi.org/10.1016/j.jhazmat.2016.05.034 Szejtli J (1998) Introduction and general overview of cyclodextrin chemistry. Chem Rev 98:1743–1753. https://doi.org/10.1021/cr970022c Szejtli J, Szente L (2005) Elimination of bitter, disgusting tastes of drugs and foods by cyclodextrins. Eur J Pharm Biopharm 61:115–125. https://doi.org/10.1016/j.ejpb.2005.05.006 Tang S, Yuan D, Rao Y, Li N, Qi J, Cheng T, Sun Z, Gu J (2018a) Huang H Persulfate activation in gas phase surface discharge plasma for synergetic removal of antibiotic in water. Chem Eng J 337:446–454. https://doi.org/10.1016/j.cej.2017.12.117

102

N. Morin-Crini et al.

Tang S, Yuan D, Rao Y, Zhang J, QumY GJ (2018b) Evaluation of antibiotic oxytetracycline removal in water using a gas phase dielectric barrier discharge plasma. J Environ Manag 226:22–29. https://doi.org/10.1016/j.jenvman.2018.08.022 Tang S, Yuan D, Rao Y, Li M, Shi G, Gu J, Zhang T (2019) Percarbonate promoted antibiotic decomposition in dielectric barrier discharge plasma. J Hazard Mater 366:669–676. https://doi. org/10.1016/j.jhazmat.2018.12.056 Tay K, Rahman N, Abas M (2010) Ozonation of parabens in aqueous solution: kinetics and mechanism of degradation. Chemosphere 81:1446–1453. https://doi.org/10.1016/j. chemosphere.2010.09.004 Teijón G, Candela L, Tamoh K, Molina-Diáz A, Fernández-Alba AR (2010) Occurrence of emerging contaminants, priority substances (2008/105/EC) and heavy metals in treated wastewater and groundwater at Depurbaix facility (Barcelona, Spain). Sci Total Environ 408:3584–3595. https://doi.org/10.1016/j.scitotenv.2010.04.041 Ternes TA, Stuber J, Herrmann N, McDowell D, Ried A, Kampmann M, Teiser B (2003) Ozonation: a tool for removal of pharmaceuticals, contrast media and musk fragrances from wastewater? Water Res 37:1976–1982. https://doi.org/10.1016/S0043-­1354(02)00570-­5 Thanekar P, Murugesan P, Gogate PR (2018) Improvement in biological oxidation process for the removal of dichlorvos from aqueous solutions using pretreatment based on hydrodynamic cavitation. J Water Proc Eng 23:20–26. https://doi.org/10.1016/j.jwpe.2018.03.004 Thomas B, Raj MC, Athira KB, Rubiyah MH, Joy J, Moores A, Drisko GL, Sanchez C (2018) Nanocellulose, a versatile green platform: from biosources to materials ant their applications. Chem Rev 118:11575–11625. https://doi.org/10.1021/acs.chemrev.7b00627 Thompson J, Eaglesham G, Reungoat J, Poussade Y, Bartkow M, Lawrence M, Mueller JF (2011) Removal of PFOS, PFOA and other perfluoroalkyl acids at water reclamation plants in South East Queensland Australia. Chemosphere 82:9–17. https://doi.org/10.1016/j. chemosphere.2010.10.040 Thurman M (2006) Accurate-mass identification of chlorinated and brominated products of 4-­ nonylphenol, nonylphenol dimers, and other endocrine disrupters. J Mass Spectrom 41:1287–1297 Tian W, Kushner MJ (2014) Atmospheric pressure dielectric barrier discharges interacting with liquid covered tissue. J Phys D Appl Phys 47:165201. https://doi.org/10.1088/ 0022-­3727/47/16/165201 Tolboom SN, Carrillo-Nieves D, de Jesús R-AM, de la Cruz QR, Barceló D, Iqbal HMN, Parra-Saldivar R (2019) Algal-based removal strategies for hazardous contaminants from the environment  – a review. Sci Total Environ 665:358–366. https://doi.org/10.1016/j. scitotenv.2019.02.129 Tomasini A, León-Santiesteban HH (2019) Fungal bioremediation. Fundamentals and applications. CRC Press Taylor & Francis Group. ISBN: 9781138636408 Torres-Palma RA, Serna-Galvis EA (2018) Chapter 7: Sonolysis. In: Advanced oxidation processes for waste water treatment, pp 117–213. https://doi.org/10.1016/B978-­0-­12-­810499-­6.00007-­3 Uekama K, Hirayama F, Arima H (2006) Recent aspect of cyclodextrin-based drug delivery system. J Incl Phenom Macrocycl Chem 56:3–8. https://doi.org/10.1007/s10847-­006-­9052-­y Valero P, Verbel M, Silva-Agredo J, Mosteo R, Ormad MP, Torres-Palma RA (2017) Electrochemical advanced oxidation processes for Staphylococcus aureus disinfection in municipal WWTP effluents. J Environ Manag 198:256–265. https://doi.org/10.1016/j.jenvman.2017.04.070 Vandenkoornhuyse P, Quaiser A, Duhamel M, Le Van A, Dufresne A (2015) The importance of the microbiome of the plant holobiont. New Phytol 206:1196–1206. https://doi.org/10.1111/ nph.13312 Vanraes P, Willems G, Nikiforov A, Surmont P, Lynen F, Vandamme J, Van Durme J, Verheust YP, Van Hulle SWH, Dumoulin A, Leys C (2015) Removal of atrazine in water by combination of activated carbon and dielectric barrier discharge. J Hazard Mater 299:647–655. https://doi. org/10.1016/j.jhazmat.2015.07.075 Vanraes P, Ghodbane H, Davister D, Wardenier N, Nikiforov A, Verheust YP, Van Hulle SWH, Hamdaoui O, Vandamme J, Van Durme J, Surmont P, Lynen F, Leys C (2017) Removal of

1  Remediation of Emerging Contaminants

103

several pesticides in a falling water film DBD reactor with activated carbon textile: energy efficiency. Water Res 116:1–12. https://doi.org/10.1016/j.watres.2017.03.004 Vasiliadou IA, Sánchez-Vázquez R, Molina R, Martínez F, Melero JA, Bautista LF, Iglesias J, Morales G (2016) Biological removal of pharmaceutical compounds using white-rot fungi with concomitant FAME production of the residual biomass. J Environ Manag 180:228–237. https:// doi.org/10.1016/j.jenvman.2016.05.035 Vergani L, Mapelli F, Zanardini E, Terzaghi E, Di Guardo A, Morosini C, Raspa G, Borin S (2017) Phyto-rhizoremediation of polychlorinated biphenyl contaminated soils: an outlook on plant-­ microbe beneficial interactions. Sci Total Environ 575:1395–1406. https://doi.org/10.1016/j. scitotenv.2016.09.218 Verlicchi P, Al Aukidy M, Zambelle E (2012) Occurrence of pharmaceutical compounds in urban wastewater: removal, mass load and environmental risk after a secondary treatment – a review. Sci Total Environ 429:123–155. https://doi.org/10.1016/j.scitotenv.2012.04.028 Vieira WT, de Farias MB, Spaolonzi MP, da Silva MGC, MGA V (2020) Removal of endocrine disruptors in waters by adsorption, membrane filtration and biodegradation. A review. Environ Chem Lett 2020. https://doi.org/10.1007/s10311-­020-­01000-­1 Vijayaraghavan K, Yun YS (2008) Bacterial biosorbents and biosorption. Biotechnol Adv 26:266–291. https://doi.org/10.1016/j.biotechadv.2008.02.002 Villegas-Guzman P, Silva-Agredo J, Florez O, Giraldo-Aguirre AL, Pulgarin C, Torres-Palma RA (2017) Selecting the best AOP for isoxazolyl penicillins degradation as a function of water characteristics: effects of pH, chemical nature of additives and pollutant concentration. J Environ Manag 190:72–79. https://doi.org/10.1016/j.jenvman.2016.12.056 Vogna D, Marotta R, Andreozzi R, Napolitano A, d’Ischia M (2004) Kinetic and chemical assessment of the UV/H2O2 treatment of antiepileptic drug carbamazepine. Chemosphere 54:497–505. https://doi.org/10.1016/S0045-­6535(03)00757-­4 Voisin H, Bergström L, Liu P, Mathew AP (2017) Nanocellulose-based materials for water purification. Nano 7:57. https://doi.org/10.3390/nano7030057 von Sonntag C, von Gunten U (2012) Chemistry of ozone in water and wastewater treatment  – from basic principles to applications. IWA Publishing, London. https://doi. org/10.2166/9781780400839 Vyas A, Saraf S, Saraf S (2008) Cyclodextrin based novel drug delivery systems. J Incl Phenom Macrocycl Chem 62:23–42. https://doi.org/10.1007/s10847-­008-­9456-­y Wang D (2019) A critical review of cellulose-based nanomaterials for water purification in industrial processes. Cellulose 26:687–701. https://doi.org/10.1007/s10570-­018-­2143-­2 Wang JL, Wang SZ (2016) Removal of pharmaceuticals and personal care products (PPCPs) from wastewater: a review. J Environ Manag 182:620–640. https://doi.org/10.1016/j. jenvman.2016.07.049 Wang JH, Zhuan R (2020) Degradation of antibiotics by advanced oxidation processes: an overview. Sci Total Environ 701:135023. https://doi.org/10.1016/j.scitotenv.2019.135023 Wang CJ, Li Z, Jiang WT (2011) Adsorption of ciprofloxacin on 2:1 dioctahedral clay minerals. Appl Clay Sci 53:723–728. https://doi.org/10.1016/j.clay.2011.06.014 Wang CL, Zou XQ, Zhao YF, Li BJ, Song QC, Li YL, Yu WW (2016a) Distribution, sources, and ecological risk assessment of polycyclic aromatic hydrocarbons in the water and suspended sediments from the middle and lower reaches of the Yangtze River, China. Environ Sci Pollut Res 23:17158–17170 Wang Y, Xie Y, Sun H, Xiao J, Cao H, Wang S (2016b) 2D/2D Nano-hybrids of γ-MnO2 on reduced graphene oxide for catalytic ozonation and coupling peroxymonosulfate activation. J Hazard Mater 301:56–64. https://doi.org/10.1016/j.jhazmat.2015.08.031 Wang Y, Xie Y, Sun H, Xiao J, Cao H, Wang S (2016c) Efficient catalytic ozonation over reduced graphene oxide for p-hydroxylbenzoic acid (PHBA) destruction: active site and mechanism. Appl Mater Interfaces 8:9710–9720. https://doi.org/10.1021/acsami.6b01175

104

N. Morin-Crini et al.

Wang T, Huang ZX, Miao HF, Ruan WQ, Ji XP, Sun FB, Ren HY (2018a) Insights into influencing factor, degradation mechanism and potential toxicity involved in aqueous ozonation of oxcarbazepine. Chemosphere 201:189–196. https://doi.org/10.1016/j.chemo sphere.2018.02.062 Wang J, Chen S, Quan X, Yu HG (2018b) Fluorine-doped carbon nanotubes as an afficient metal-free catalyst for destruction of organic pollutants in catalytic ozonation. Chemosphere 190:135–143. https://doi.org/10.1016/j.chemosphere.2017.09.119 Wang Y, Cao H, Chen L, Chen C, Duan X, Xie Y, Song W, Sun H, Wang S (2018c) Tailored synthesis of active reduced graphene oxides from waste graphite: structural defects and pollutant-­ dependent reactive radicals in aqueous organics decontamination. Appl Catal B Environ 229:71–80. https://doi.org/10.1016/j.apcatb.2018.02.010 Wang C, Qu G, Wang T, Deng F, Liang D (2018d) Removal of tetracycline antibiotics from wastewater by pulsed corona discharge plasma coupled with natural soil particles. Chem Eng J 346:159–170. https://doi.org/10.1016/j.cej.2018.03.149 Wang J, Quan X, Chen S, Yu H, Liu G (2019a) Enhanced catalytic ozonation by highly dispersed CeO2 on carbon nanotubes for mineralization of organic pollutants. J Hazard Mater 368:621–629. https://doi.org/10.1016/j.jhazmat.2019.01.095 Wang Y, Cao H, Chen C, Xie Y, Sun H, Duan X, Wang S (2019b) Metal-free catalytic ozonation on surface-engineered graphene: microwave reduction and heteroatom doping. Chem Eng J 355:118–129. https://doi.org/10.1016/j.cej.2018.08.134 Wang B, Wang C, Yao S, Peng Y, Xu Y (2019c) Plasma-catalytic degradation of tetracycline hydrochloride over Mn/γ-Al2O3 catalysts in a dielectric barrier discharge reactor. Plasma Sci Technol 21:065503. https://doi.org/10.1088/2058-­6272/ab079c Warsinger DM, Chakraborty S, Tow EW, Plumlee MH, Bellona C, Loutatidou S, Karimi L, Mikelonis AM, Achilli A, Ghassemi A, Padhye LP, Snyder SA, Curcio S, Vecitis CD, Arafat HA, Lienhard VJH (2018) A review of polymeric membranes and processes for potable water reuse. Prog Polym Sci 81:209–237. https://doi.org/10.1016/j.progpolymsci.2018.01.004 Watkinson AJ, Murby EJ, Kolpin DW, Costanzo SD (2009) The occurrence of antibiotics in an urban watershed: from wastewater to drinking water. Sci Total Environ 407:2711–2723 Wert EC, Rosario-Ortiz FL, Drury DD, Snyder SA (2007) Formation of oxidation byproducts from ozonation of wastewater. Water Res 41:1481–1490. https://doi.org/10.1016/j. watres.2007.01.020 Westerhoff P, Yoon Y, Snyder S, Wert E (2005) Fate of endocrine-disruptor, pharmaceutical, and personal care product chemicals during simulated drinking water treatment processes. Environ Sci Technol 39:6649–6663. https://doi.org/10.1021/es0484799 Wiedenhof N, Lammers JNJJ, Van Panthaleon Van Eck CL (1969) Properties of cyclodextrins. III Cyclodextrin-epichlorhydrin resins Preparation and analysis Stärke 21:119–123. https://doi. org/10.1002/star.19690210504 Wimmer T, Kreuzer FH, Staudinger G, Nussstein P (1992) Synthesis of novel insoluble cyclodextrin polymers. In: Hedges AR (ed) Minutes Int. Symp. Cyclodextrins, 6th (1992), 106–9. Editions de Santé, Paris, France Xiang Y, Fang J, Shang C (2016) Kinetics and pathways of ibuprofen degradation by the UV/ chlorine advanced oxidation process. Water Res 90:301–308 Xie M, Chen W, Xu Z, Zheng S, Zhu D (2014) Adsorption of sulfonamides to demineralized pine wood biochars prepared under different thermochemical conditions. Environ Pollut 186:187–194. https://doi.org/10.1016/j.envpol.2013.11.022 Xin L, Sun Y, Feng J, Wang J, He D (2016) Degradation of triclosan in aqueous solution by dielectric barrier discharge plasma combined with activated carbon fibers. Chemosphere 144:855–863. https://doi.org/10.1016/j.chemosphere.2015.09.054 Xiong Q, Yang Z, Bruggeman PJ (2015) Absolute OH density measurements in an atmospheric pressure dc glow discharge in air with water electrode by broadband UV absorption spectroscopy. J Phys D Appl Phys 48:424008. https://doi.org/10.1088/0022-­3727/48/42/424008

1  Remediation of Emerging Contaminants

105

Xiong JQ, Kurade MB, Abou-Shanab RAI, Ji MK, Choi JJO, Jeon BH (2016) Biodegradation of carbamazepine using freshwater microalgae Chlamydomonas mexicana and Scenedesmus obliquus and the determination of its metabolic fate. Bioresour Technol 205:183–190 Xu P, Janex ML, Savoye P, Cockx A, Lazarova V (2002) Wastewater disinfection by ozone: main parameters for process design. Water Res 36:1043–1055. https://doi.org/10.1016/ S0043-­1354(01)00298-­6 Xu G, Xie X, Qin L, Hu X, Zhang D, Xu J, Li D, Ji X, Huang Y, Tu Y, Jiang L, Wei D (2019) Simple synthesis of a swellable porous β-cyclodextrin-based polymer in the aqueous phase for the rapid removal of organic micro-pollutants from water. Green Chem 21:6062–6072. https:// doi.org/10.1039/C9GC02422K Xu Z, Xue X, Hu S, Li Y, Shen J, Lan Y, Zhou R, Yang F, Cheng C (2020) Degradation effect and mechanism of gas-liquid phase dielectric barrier discharge on norfloxacin combined with H2O2 or Fe2+. Sep Purif Technol 230:115862. https://doi.org/10.1016/j.seppur.2019.115862 Yadav AN, Mishra S, Singh S, Gupta (2019) Recent advancement in white biotechnology through fungi. Springer, Cham. https://doi.org/10.1007/978-­3-­030-­10480-­1 Yamasaki H, Makihata Y, Fukunaga K (2008) Preparation of crosslinked β-cyclodextrin polymer beads and their application as a sorbent for removal of phenol from wastewater. J Chem Technol Biotechnol 83:991–997. https://doi.org/10.1002/jctb.1904 Yamasaki H, Odamura A, Makihata Y, Fukunaga K (2017) Preparation of new photo-crosslinked β-cyclodextrin polymer beads. Polym J 49:377–383. https://doi.org/10.1038/pj.2016.127 Yang TR (2003) Adsorbents: fundamentals and applications. Wiley-Interscience, Hoboken, 424 p Yang W, Zhou H, Cicek N (2013) Treatment of organic micropollutants in water and wastewater by UV-based processes: a literature review. Critical Rev Environ Sci Technol 44:1443–1476. https://doi-­org.inee.bib.cnrs.fr/10.1080/10643389.2013.790745 Yang Z, Wu G, LimQ AH, Yao X, Ji H (2020) Removal of various pollutants from wastewaters using an efficient and degradable hypercrosslinked polymer. Sep Sci Technol. https://doi.org/1 0.1080/01496395.2020.1745239 Yin RL, Guo WQ, Du JS, Zhou XJ, Zheng H, Wu QL, Chang J, Ren NQ (2017) Heteroatoms doped graphene for catalytic ozonation of sulfamethoxazole by metal-free catalysis: performances and mechanisms. Chem Eng J 317:632–639. https://doi.org/10.1016/j.cej.2017.01.038 Ying GG, Kookana RS, Kolpin DW (2009) Occurrence and removal of pharmaceutically active compounds in sewage treatment plants with different technologies. J Environ Monitoring 11:1498–1505. https://doi.org/10.1039/B904548A Yu Z, Peldszus S, Huck PM (2009) Adsorption of selected pharmaceuticals and an endocrine disrupting compound by granular activated carbon. I. Adsorption capacity and kinetics. Environ Sci Technol 43:–1467, 1473 Yu X, Tong S, Ge M, Wu L, Zuo J, Cao C, Song W (2013) Adsorption of heavy metal ions from aqueous solution by carboxylated cellulose nanocrystals. J Environ Sci 25:933–943 Yu Z, Sun Y, Zhang G, Zhang C (2017) Degradation of DEET in aqueous solution by water falling film dielectric barrier discharge: effect of three operating modes and analysis of the mechanism and degradation pathway. Chem Eng J 317:90–102. https://doi.org/10.1016/j.cej.2017.02.068 Yuan F, Hu C, Hu X, Qu J, Yang M (2009) Degradation of selected pharmaceuticals in aqueous solution with UV and UV/H2O2. Water Res 43:1766–1774. https://doi.org/10.1016/j. watres.2009.01.008 Zha SX, Zhou Y, Jin X, Chen Z (2013) The removal of amoxicillin from wastewater using organobentonite. J Environ Manag 129:569–576. https://doi.org/10.1016/j.jenvman.2013.08.032 Zhang S, Grimm D (2005) Naproxen removal from water by chlorination and biofilm processes. Water Res 39:668–676 Zhang Y, Geiβen SU (2012) Elimination of carbamazepine in a non-sterile fungal bioreactor. Bioresour Technol 112:221–227 Zhang Z, Feng Y, Liu Y, Sun Q, Gao P, Ren N (2010) Kinetic degradation model and estrogenicity changes of EE2 (17α-ethinylestradiol) in aqueous solution by UV and UV/H2O2 technology. J Hazard Mater 181:1127–1133. https://doi.org/10.1016/j.jhazmat.2010.05.132

106

N. Morin-Crini et al.

Zhang Q, Deng S, Yu G, Huang J (2011) Removal of perfluorooctane sulfonate from aqueous solution by crosslinked chitosan beads: sorption kinetics and uptake mechanism. Bioresour Technol 102:2265–2271 Zhang Y, Xie J, Liu M, Tian Z, He Z, van Nostrand JD, Ren L, Zhou J, Yang M (2013) Microbial community functional structure in response to antibiotics in pharmaceutical wastewater treatment systems. Water Res 47:6298–6308 Zhang X, Guo W, Ngo HH, Wen H, Li N, Wu W (2016a) Performance evaluation of powdered activated carbon for removing 28 types of antibiotics from water. J Environ Manag 172:193–200. https://doi.org/10.1016/j.jenvman.2016.02.038 Zhang Y, Lv T, Carvalho PN, Arias CA, Chen Z, Brix H (2016b) Removal of the pharmaceuticals ibuprofen and iohexol by four wetland plant species in hydroponic culture: plant uptake and microbial degradation. Environ Sci Pollut Res 23:2890–2898. https://doi.org/10.1007/ s11356-­015-­5552-­x Zhang Q, Zhang H, Zhang Q, Huang Q (2018) Degradation of norfloxacin in aqueous solution by atmospheric pressure non-thermal plasma: mechanism and degradation pathways. Chemosphere 210:433–439. https://doi.org/10.1016/j.chemosphere.2018.07.035 Zhao D, Zhao L, Zhu CS, Huang WQ, Hu JL (2009) Water-insoluble beta-cyclodextrin polymer crosslinked by citric acid: synthesis and adsorption properties toward phenol and methylene blue. J Incl Phenom Macrocyclic Chem 63:195–201. https://doi.org/10.1007/ s10847-­008-­9507-­4 Zhongbo Z, Hu J (2008) Selective removal of estrogenic compounds by molecular imprinted polymer (MIP). Water Res 42:4101–4108 Zhou AJ, Chen WG, Liao L, Xie PC, Zhang TC, Wu XM, Feng XN (2019) Comparative adsorption of emerging contaminants in water by functional designed magnetic poly(N-­ isopropylacrylamide)/chitosan hydrogels. Sci Total Environ 671:377–387. https://doi. org/10.1016/j.scitotenv.2019.03.183 Zhu XX, Brizard F, Wen CC, Brown GR (1997) Binding of bile acids by polymerized cyclodextrin. J Macromol Sci Pure Appl Chem A34:335–347. https://doi.org/10.1080/10601329708014959 Zhu R, Chen Q, Zhou Q, Xi Y, Zhu J, He H (2016) Adsorbents based on montmorillonite for contaminant removal from water: a review. Appl Clay Sci 123:239–258. https://doi.org/10.1016/j. clay.2015.12.024 Zimmermann SG, Wittenwiler M, Hollender J, Krauss M, Ort C, Siegrist H, von Gunten U (2011) Kinetic assessment and modeling of an ozonation step for full-scale municipal wastewater treatment: micropollutant oxidation, by-product formation and disinfection. Water Res 45:605–617. https://doi.org/10.1016/j.watres.2010.07.080 Zraunig A, Estelrich M, Gattringer H, Kisser J, Langergraber G, Radtke M, Rodriguez-Roda I, Buttiglieri G (2019) Long term decentralized greywater treatment for water reuse purposes in a tourist facility by vertical ecosystem. Ecol Eng 138:138–147. https://doi.org/10.1016/j. ecoleng.2019.07.003

Chapter 2

Electrochemical Treatments for the Removal of Emerging Contaminants Borislav N. Malinović, Jernej Markelj, Helena Prosen, Andreja Žgajnar Gotvajn, and Irena Kralj Cigić

Abstract  Emerging contaminants are a diverse group of predominantly anthropogenic pollutants which are present in the environment in the concentrations that can cause known or suspected adverse environmental and health effects. The common denominator of these diverse groups of contaminants is the fact that there is a lack of legislative regulation, as well as lack of monitoring programs or risk assessment studies. For these reasons, their removal from the sources of pollution and from the environment is of key importance. Among possible approaches for the remediation of wastewater as the main source of pollution, electrochemical treatments have recently gained increasing attention as promising, feasible, and environmentally friendly removal techniques. In this chapter, we have reviewed and critically presented the applications of electrochemical treatments for the removal of selected groups of emerging contaminants from wastewater and similar samples. These contaminants are: bisphenol A, phthalic acid esters, and benzotriazoles, chosen because they are linked to intensive production of plastics, as well as on the basis of a relative lack of review papers. First, we present and compare various available electrochemical treatment approaches, their combinations, and the most frequently applied electrodes. In the main part, the available studies for the chosen contaminants from the last 5 years are reviewed with the emphasis on the effectiveness of the treatment; identification of the transformation products; and elucidation of the degradation mechanisms. We identify the main problems of the reviewed electrochemical treatments: the lack of B. N. Malinović Faculty of Technology, University of Banja Luka, Banja Luka, Bosnia and Herzegovina e-mail: [email protected] J. Markelj · H. Prosen (*) · A. Žgajnar Gotvajn · I. Kralj Cigić Faculty of Chemistry and Chemical Technology, University of Ljubljana, Ljubljana, Slovenia e-mail: [email protected]; [email protected]; andreja.zgajnar@fkkt. uni-lj.si; [email protected] © The Author(s), under exclusive license to Springer Nature Switzerland AG 2021 N. Morin-Crini et al. (eds.), Emerging Contaminants Vol. 2, Environmental Chemistry for a Sustainable World 66, https://doi.org/10.1007/978-3-030-69090-8_2

107

108

B. N. Malinović et al.

identification of the transformation products, as well as the general lack of toxicity assessment of the treated samples. Although electrochemical treatments are viable and effective processes for the removal of selected emerging contaminants from wastewater as shown by the high degree of mineralization in several reviewed papers, further studies need to be done with the emphasis on the identified shortcomings. Keywords  Emerging contaminants · Additives in plastic products · Phthalates · Bisphenol A · Benzotriazoles · Electrochemical treatment

Abbreviations AOP AOPs BDD BOD COD GC-MS HPLC-UV LC-MS/MS SHE TOC

advanced oxidation process advanced oxidation processes boron-doped diamond electrode biological oxygen demand chemical oxygen demand gas chromatography coupled with mass spectrometry high-performance liquid chromatography with UV detection liquid chromatography coupled with tandem mass spectrometry standard hydrogen electrode total organic carbon

2.1  Introduction The contemporary times are arguably more than ever before in the history of mankind characterized by widespread concerns regarding our environment. Media reports are full of global issues related to changed weather patterns and warming of the atmosphere, with the ensuing polarized and frequently ill-informed debate on the causes for the obviously changing climate, as well as pros and contras of fossil fuels vs. renewable energy sources. Politicians talk about these topics; there are climate protests and marches. Yet, there is a sinister environmental issue that is not so obviously seen and recognized by the public: widespread chemical pollution of all environmental compartments: atmosphere, waters, soil, and biota. Human activity has influenced the distribution of chemicals in the environment for centuries, but the only real chemical pollution, as we understand it now, was for a long time caused only by mining and smelting, which resulted in increased concentrations of metals and metalloids in the whereabouts of these sites, and sometimes the pollution spread further downstream along the rivers. Industrial revolution increased the pollution with these chemicals, although it was recognized as such only much later. Metals and metalloids could therefore be named “old” pollutants or “old” contaminants.

2  Electrochemical Treatments for the Removal of Emerging Contaminants

109

Another type of chemical pollution first appeared with the synthetic organic chemicals in the nineteenth century. Again, they were recognized as pollutants only much later, with the recognition that pesticides such as dichlorodiphenyltrichloroethane (DDT) can cause adverse effects on wildlife (Carson 2002). DDT and other chlorinated pesticides were soon banned from widespread use and are now considered “old” contaminants but are unfortunately, together with their transformation products, still present in the environment because of their persistence. As an opposite to “old” or “traditional” contaminants, “new” or emerging contaminants are those that have more recently came in the spotlight of the environmentalists and legislative bodies. Environmental chemists publishing papers on this issue most often resort to the definition that these are environmental pollutants not included in monitoring programs or not regulated by any legislation (Zedda and Zwiener 2012; Geissen et  al. 2015; Bieber et  al. 2018; Carmona and Picó 2018; Bunke et al. 2019; García-Córcoles et al. 2019; Gaston et al. 2019), as opposed to “traditional” contaminants that are well regulated and sometimes even banned from production (Richardson and Kimura 2017). Emerging contaminants are principally released into the environment via wastewater treatment plants using activated sludge, because they are inadequately removed by the traditional aerobic digestion processes (Richardson and Kimura 2017; Taheran et al. 2018). Therefore, several novel methodologies have been proposed to remove or degrade emerging contaminants in wastewater, and electrochemical treatments are one of the less applied, but nevertheless viable options for their removal. In the present chapter, we have focused on the electrochemical treatments of emerging contaminants in wastewater. Based on our current research interests and a relative lack of review papers, we have reviewed the electrochemical treatments of anticorrosive agents and of contaminants related to the production and environmental degradation of plastic materials. Specifically, three distinct groups of compounds were considered: bisphenol A, o-phthalic acid esters, and benzotriazoles. We have focused on the research papers from the last 5 years, but when necessary, older literature has been cited as well. The search for papers was done in Web of Science, using the keywords of the current review. Because of bisphenol A’s now already notorious xenoestrogenic effects, a high number of research papers were found. Other bisphenol analogues are gaining in importance since the phasing-out of bisphenol A, but their electrochemical removal hasn’t as yet been studied. Although phthalates are high-production plastic additives and well known as endocrine disruptors, a significantly lower number of papers were found. Compared to the previous two groups, a very small number of papers on benzotriazoles’ removal were found. The definition, properties and available treatment options for emerging contaminants are first described. Then, we describe and compare the various available electrochemical treatment approaches, their combinations, and the most frequently applied electrodes as the most important part of the electrochemical system. In the next three sections, we present the review of papers dealing with the applications of electrochemical treatments on the selected three groups of emerging contaminants.

110

B. N. Malinović et al.

Fig. 2.1  Graphical representation of the contents of the chapter

Special emphasis in reviewing was put on the monitoring of the treatment effectiveness; as well as the identification and determination of the transformation products and the elucidation of the degradation mechanisms. The problems of the electrochemical treatments for the selected groups of emerging contaminants are then summarized, i.e. the problem of transformation products, the toxicity assessment of the treated samples, as well as the operational and engineering problems associated with the application of electrochemical treatments on a large scale. Finally, a short overview of the other available treatment options for the removal of emerging contaminants from wastewater is described. Figure 2.1 summarizes the contents of the chapter.

2.2  Emerging Contaminants A more widespread human awareness of chemical pollution can be said to begin in the sixties of the previous century, with the recognition of chlorinated pesticides’ adverse effects on wildlife, published in a book ‘Silent Spring’ by Rachel Carson in 1962 (Carson 2002). Soon, environmental agencies, such as Environmental Protection Agency in USA; corresponding environmental agencies of European countries, Canada, Australia, etc. started to be established and devoted their efforts

2  Electrochemical Treatments for the Removal of Emerging Contaminants

111

to the quantification of chemical pollutants in the environment and research on their adverse effects. Thus, in the sixties and seventies of the twentieth century, pesticides, metals, and metalloids, started to be considered as environmental pollutants for the first time. The latter two groups originate from centuries-long mining and smelting activities, but were not considered as pollutants before. These pollutants could therefore be named “emerging contaminants” if the term had existed at that time. Yet they were present in the environment for at least a few decades or even centuries. What, then, defines an emerging contaminant?

2.2.1  Definition of Emerging Contaminants Sauvé and Desrosiers (2014) make an interesting case on what an emerging contaminant is. This term, now in widespread use by environmental chemists, actually appeared around year 2000 and gained wide popularity perhaps mainly due to a series of highly influential review papers by Susan D.  Richardson from USA Environmental Protection Agency (Richardson 2002, 2003, 2004, 2006, 2007, 2008, 2009, 2010, 2012; Richardson and Kimura 2016; Richardson and Ternes 2005, 2011, 2014, 2018). Other influential researchers of environmental pollution use the term “emerging pollutants” (Lopez de Alda et al. 2003; NORMAN). Both names suggest a sudden appearance of the contaminant/pollutant in the environment, but it is rarely so. As shown above for chlorinated pesticides and metals, they can be present in the environment long before the attention of environmental chemists shifts to them. In her review papers from 2002 to 2008, Richardson has included arsenic as an emerging contaminant (Richardson 2002, 2003, 2004, 2006, 2007, 2008; Richardson and Ternes 2005). Arsenic is present in the environment either naturally or as a consequence of human activity; its toxicity has been known since antiquity, yet it has again become the topic of many scientific papers because of the fact that it was significantly mobilized in the environment due to changed land use and consequently again became a major health problem in certain geographical areas. Also, certain microorganisms (Richardson 2002, 2003, 2004, 2006, 2007; Richardson and Ternes 2005) can become emerging contaminants if they suddenly cause previously unexpected outbreaks of illness. Therefore, as Sauvé and Desrosiers (2014) pointed out, it would probably be better to use the term “contaminants of emerging concern”. Yet other names are used, e.g. “anthropogenic contaminants of high concern” (López-Pacheco et al. 2019), or “substances of very high concern” (ECHA). In the present review, we will continue to use the term emerging contaminants because it is the most widespread in the scientific literature. But what do most environmental chemists understand when talking about emerging contaminants? We’ve already made a case that any chemical in the environment or even something of a higher organization order, such as a microorganism, can be considered an emerging contaminant if it suddenly becomes a recognized environmental problem. It is

112

B. N. Malinović et al.

probably safe to say that a chemical not causing any problems at either low or high concentration could never be considered an emerging contaminant; although, such a substance probably doesn’t exist. Another interesting point is our ability to detect and quantify a substance in very low concentrations present in the environmental samples. This has become possible only with the development of sophisticated analytical instrumentation, especially liquid chromatography-(tandem)mass spectrometry (LC-MS, LC-MS/MS), some 25 years ago (Noguera-Oviedo and Aga 2016). Before that, we were not able to detect most emerging contaminants in the environment, even if they were already present. However, the general understanding of the meaning of the term “emerging contaminant” is narrower in most cases. Environmental chemists usually apply the definition that these are environmental pollutants not included in monitoring programs or not regulated by any legislation (Zedda and Zwiener 2012; Geissen et al. 2015; Bieber et  al. 2018; Carmona and Picó 2018; Bunke et  al. 2019; García-Córcoles et al. 2019; Gaston et al. 2019). Besides the lack of regulation, emerging contaminants are those substances that also exhibit some adverse effects for human health or for the environment. Human exposure to emerging contaminants via the environment is currently very low, but the effects of these substances on other living organisms, especially aquatic wildlife, are concerning (Richardson and Kimura 2017; Nilsen et al. 2019). NORMAN network, an initiative of laboratories and authorities for the exchange of information on emerging contaminants, uses the following definition: “pollutants that are currently not included in routine monitoring programs at the European level and which may be candidates for future regulation, depending on research on their (eco)toxicity, potential health effects and public perception and on monitoring data regarding their occurrence in the various environmental compartments” (NORMAN). UNESCO uses a similar definition: “Emerging pollutants can be understood in a broad sense as any synthetic or naturally-occurring chemical or any microorganism that is not commonly monitored or regulated in the environment with potentially known or suspected adverse ecological and human health effects” (UNESCO). The current list of substances that are defined as emerging contaminants is very long; the last updated list from the NORMAN network in 2016 gives a total of 1036 individual substances organized into more than 20 categories or groups (NORMAN). Most belong to the following groups: perfluoroalkyl substances; nanomaterials; pharmaceuticals and illicit drugs; constituents of personal care products, such as antimicrobials, UV filters, preservatives, and fragrances; hormones; flame retardants, such as brominated diphenyl ethers, phosphate esters, and other; disinfection byproducts in drinking water or in swimming pools; artificial sweeteners; anticorrosives, such as benzotriazoles and benzothiazoles; and plasticizers, such as phthalates and bisphenol A (Geissen et  al. 2015; Petrie et  al. 2015; Carmona and Picó 2018). Additionally, 1,4-dioxane, gasoline additives, new pesticides, algal and cyanobacterial toxins, wood preservatives, naphthenic acids, microorganisms, and antibiotic resistance genes are sometimes included (Geissen et al. 2015; Noguera-Oviedo and Aga 2016; Richardson and Ternes 2018). More recently, microplastics, siloxanes, prions, ionic liquids, and halogenated methanesulfonic acids were recognized

2  Electrochemical Treatments for the Removal of Emerging Contaminants

113

as novel emerging contaminants (Richardson and Kimura 2016; Richardson and Ternes 2018). Others will be added to the list as society evolves and new substances are introduced as replacements for those restricted or banned from production (Bunke et al. 2019). Those emerging contaminants that are recognized as the most harmful have already been included in the watchlists in the European Union (Sousa et al. 2018; Jurado et al. 2019), USA (Richardson and Kimura 2017), and elsewhere (Bieber et al. 2018), but most remain unregulated.

2.2.2  Properties and Distribution of Emerging Contaminants There are several features that are common to most of the above-mentioned groups of emerging contaminants: (i) the concentrations in various environmental matrices are usually very low (Taheran et al. 2018; García-Córcoles et al. 2019); (ii) human health issues are therefore not related to their acute toxicity, but to long-term exposure with possible links to carcinogenicity and reproductive health risks (Lei et al. 2015); (iii) environmental issues are likewise related to long-term exposure effects, e.g. endocrine disruption in aquatic wildlife (Richardson and Kimura 2017), and to the fact that organisms in the environment are exposed to a cocktail of emerging contaminants (Petrie et al. 2015; Nilsen et al. 2019); (iv) only some emerging contaminants are bioaccumulative, but (aquatic) organisms are nevertheless continuously exposed (Richardson and Kimura 2017); (v) most emerging contaminants are (bio)degradable, but due to the continuous release into the environment, they are considered “pseudopersistent” (Taheran et al. 2018); (vi) in the environment, many emerging contaminants are transformed by microbial degradation, photolysis, and other abiotic processes to transformation products that may be equally or more problematic than the parent compounds (Bletsou et  al. 2015; Richardson and Kimura 2017; Kotthoff et al. 2019); (vii) a majority of emerging contaminants are released into the environment via the wastewater treatment plants based on biological treatment – Fig. 2.2 (Richardson and Kimura 2017; Taheran et al. 2018), but also from stock farming, agriculture, and industry (Geissen et al. 2015); and (viii) their environmental fate, spatial and temporal distribution are poorly understood (Petrie et al. 2015; Wilkinson et al. 2017). Moreover, these parameters are likely to change due to climate changes with altered rain and draught patterns (Richardson and Kimura 2017). With regard to the latter, it should be noted that emerging contaminants have been the most extensively studied in wastewater (García-Córcoles et al. 2019), surface waters including seawater (Bletsou et  al. 2015; Wilkinson et  al. 2017; Richardson and Ternes 2018; García-Córcoles et  al. 2019; López-Pacheco et  al. 2019), and groundwater (Sousa et al. 2018; Gaston et al. 2019; Jurado et al. 2019). Their transfer and occurrence in wastewater sludge (Petrie et al. 2015), sediments and soil (Careghini et  al. 2015; Vodyanitskii and Yakovlev 2016), atmosphere (Salgueiro-González et al. 2015; Barroso et al. 2019), plants (Christou et al. 2019;

114

B. N. Malinović et al.

Fig. 2.2  Municipal biological wastewater treatment plant. (Source: Matej Čehovin)

Matich et al. 2019), animals (Nilsen et al. 2019), and even food (Careghini et al. 2015), however,  have only recently gained attention. Besides these understudied samples, there are other areas of research on emerging contaminants which are not adequately covered. To name just a few: since many emerging contaminants are primarily pharmacologically active compounds, their chiral distribution and possible different behavior in the environment is an important issue (Petrie et al. 2015). Distribution of emerging contaminants in the environment is inseparable from their sorption to natural organic matter, solid particles in the atmosphere (Barroso et al. 2019), activated sludge in wastewater treatment plants, and even microplastic particles (Wilkinson et al. 2017). The toxicity of emerging contaminants for aquatic organisms has been widely studied, but the opposite is true for soil fauna (Petrie et al. 2015; Vodyanitskii and Yakovlev 2016). These issues remain a challenge. Another important issue is the distribution of emerging contaminants in different geographical areas of the world. Although emerging contaminants are present worldwide (Noguera-Oviedo and Aga 2016), their occurrence and concentrations are generally well documented in European Union countries, USA, Canada, and Australia (Bieber et al. 2018; Richardson and Ternes 2018; García-Córcoles et al. 2019; López-Pacheco et al. 2019), but significantly less information is available for the rest of the world. For example, a review by Tran et  al. (2018) compared the occurrence of emerging contaminants in wastewater treatment plants’ influents in Europe, North America, and Asia. Concentrations of most emerging contaminants were higher in the samples from the Asian region. Similarly, Philip et  al. (2018) reviewed the occurrence of emerging contaminants in samples from the Indian subcontinent and found alarmingly high concentrations of certain pollutants. In Brazil, Montagner et al. (2017) reported concentrations of certain classes of emerging contaminants in water samples. High concentrations of some emerging contaminants

2  Electrochemical Treatments for the Removal of Emerging Contaminants

115

were found even in drinking water. On the other side, Sorensen et al. (2015) measured concentrations of typical emerging contaminants in urban groundwater in Zambia and found that most personal care products, life-style compounds, and pharmaceuticals were absent. Only the insect repellent diethyltoluamide was present in all samples. These differences could be attributed to different usage, climate conditions, population density, and sanitation conditions. But some authors (Tran et al. 2018) also noted the possibility of inadequate analytical methods or unsuitable sampling, therefore these data should be treated with caution.

2.2.3  Removal Options for Emerging Contaminants Emerging contaminants are principally released into the environment via wastewater treatment plants (Richardson and Kimura 2017; Taheran et al. 2018). It is a well-­ recognized problem that these substances are inadequately removed by the traditional aerobic digestion processes in urban wastewater treatment plants using activated sludge (Noguera-Oviedo and Aga 2016; Richardson and Kimura 2017; Rodriguez-Narvaez et al. 2017; Gogoi et al. 2018; Taheran et al. 2018; Fischer et al. 2019; Rasheed et al. 2019; Rizzo et al. 2019). Therefore, several novel methodologies have been proposed to remove or degrade them in wastewater. These treatments can be roughly divided into several groups based on the prevalent mechanism of removal and/or degradation. The below order of treatments is based mainly on the frequency of their application and corresponding frequency of reviews: (i) adsorption to different solids (activated carbon, biochar, carbon nanotubes, clay minerals, and others) and coagulation-flocculation (Priac et  al. 2017; Rodriguez-Narvaez et al. 2017; Gogoi et al. 2018; Rasheed et al. 2019; Rizzo et al. 2019); (ii) membrane processes, i.e. micro-, nano-, or ultra-filtration, reverse and forward osmosis (Priac et al. 2017; Rodriguez-Narvaez et al. 2017; Rizzo et al. 2019); (iii) biological treatment: aerobic and anaerobic microorganisms, constructed wetlands, and enzyme-­ mediated (Priac et al. 2017; Rodriguez-Narvaez et al. 2017; Gogoi et al. 2018; Bilal et al. 2019; Rasheed et al. 2019); (iv) advanced oxidation processes (AOPs) with hydroxyl radicals, which are produced photochemically or chemically, e.g. by ozonation, O3/H2O2, UV/H2O2, and Fenton-based processes; or by other oxidants, e.g. ferrate(VI), permanganate, chlorine, and chlorine dioxide (Priac et  al. 2017; Richardson and Kimura 2017; Salimi et al. 2017; Rizzo et al. 2019); (v) photolysis and photocatalysis by UV or (rarely) visible light alone or in combination with H2O2, TiO2, and O3 (Richardson and Kimura 2017; Salimi et al. 2017; Gogoi et al. 2018; Rizzo et al. 2019); (vi) sonochemically by ultrasound (hydrodynamic) cavitation (Salimi et  al. 2017; Rizzo et  al. 2019); and (vi) electrochemical treatments (Rizzo et al. 2019). The treatments are frequently combined to exploit the advantages of several removal mechanisms.

116

B. N. Malinović et al.

2.3  Electrochemical Treatments The beginnings of the application of electrochemical treatments to wastewater treatment date back to the nineteenth century. Due to the lenient laws and regulations and relatively high investments costs, they were not widely used at that time. With rigorous environmental regulations on wastewater discharges, electrochemical treatments have globally regained their importance over the last decades and reached a level comparable to other treatments in efficiency and cost-effectiveness. On a laboratory scale, electrochemical treatments can be developed and their efficiency tested in either batch mode (Fig. 2.3) or continuous or recirculation mode (Fig. 2.4). However, wastewater treatment under realistic conditions requires scaling-­up which presents new challenges. These will be discussed separately in Sect. 2.7.3, while here, we present various available electrochemical treatments and the mechanism by which they degrade pollutants.

Fig. 2.3  Schematic of a laboratory batch reactor. (1: source of electric power; 2: anode; 3: cathode; 4: magnetic stir bar; 5: electrochemical cell; 6: magnetic stirrer/heater; 7: thermo/pH meter)

Fig. 2.4  Schematic of a laboratory batch recirculation reactor. (1: peristaltic pump; 2: tank; 3: reactor; 4: source of electric power)

2  Electrochemical Treatments for the Removal of Emerging Contaminants

117

2.3.1  Classical Electrochemical Treatments Some classical electrochemical treatments include electrodeposition/electroreduction, electrocoagulation, electroflotation, electrodialysis, and electrofiltration. While electrochemical oxidation is a long-known process, the discovery and application of new electrode materials have recently revived this process and given it increased importance. Therefore, electrooxidation is classified as a new and advanced electrochemical process. Electrodeposition is mainly related to the removal of heavy metals from wastewater or to the regeneration of precious metals from electrolytes or wastewater by cathodic deposition. The electrochemical mechanism for metal regeneration is very simple (Eq. 2.1).

M n + + ne - ® M

(2.1)

Electrochemical reduction or electroreduction at the cathode can remove many inorganic and organic pollutants from wastewater but is less prone to the formation of unwanted by-products. The latter may be the case in oxidation treatment processes. In such cases, electroreduction treatment works effectively to avoid these disadvantages and to detoxify the wastewater. For example, electroreduction is capable of transforming a wide range of organic pollutants such as organic halides under certain well defined conditions (Eq. 2.2) (Rondinini et al. 2018).

R - X + H + + 2e - ® RH + X - with X = F, Cl, Br, I

(2.2)

Electroreduction methods are effective for the dechlorination of a wide range of chlorinated organic compounds (Lei et al. 2019) and electrochemical reduction of nitrite and nitrate (Malinovic et al. 2015). Compared to other techniques, electrochemical reduction (dechlorination) is the preferred technique to degrade chlorinated compounds due to the mild reaction conditions and the avoidance of possible toxic by-products (Feng et al. 2016). The selection of materials for cathode can have a great influence on the efficiency of electroreduction processes. Electrocoagulation is the formation of coagulants by electrolytic dissolution of sacrificial electrodes, e.g. aluminum or iron, in an electrochemical reactor. The metal ions formed on the anode by hydrolysis create a large number of different compounds that are effective coagulants for the removal of pollutants. At the cathode, hydrogen is released which often carries flocculated particles to the solution surface, which is why the process is often combined with electroflotation. The electrocoagulation reactor primarily produces colloidal aggregation of coagulants (increased size) and gas bubbles which are small, if the applied current is low. The mechanism with the aluminum anode takes place according to the reactions (Eqs. 2.3, 2.4 and 2.5) (Barrera-Diaz et al. 2018).

Al  Al3+ + 3e _

(2.3)

118

B. N. Malinović et al.



2H 2 O + 2e -  H 2 ( g ) + 2OH -





Al3+ + 3OH -  Al ( OH )3 ( s )



(2.4) (2.5)

If the wastewater, which is an electrolyte in this case, does not have sufficient conductivity, a supporting electrolyte is used to increase the ionic conductivity. The typical field of electrocoagulation application is heavy metal removal, clean-up of tannery and textile industry wastewater, food industry wastewater, paper industry wastewater, refinery wastewater, and produced water (Moussa et al. 2017). The process requires simple equipment, which also leads to easier process control and flexibility of industrial operating conditions. It should be noted that using higher current densities increases removal efficiency but also increases specific energy consumption and the consumption of electrode material. The main advantage of electrocoagulation is that there is no need for temperature control (feasible at ambient temperature); it is a simple, fast, inexpensive, easily operable, and eco-friendly process. Besides, the purified water is palatable, clear, colorless, and odorless with low sludge production (Mollah et al. 2001). Bazrafshan et al. (2015) claimed that there was no chance of secondary contamination of water with these techniques. Similarly, other authors reported that since no chemicals are added, there is no chance of secondary contamination of water (Moussa et al. 2017). If there are any organic compounds in wastewater, some chlorinated organic compounds may be formed. Wastewater which contains humic and fulvic acids may be subject to the formation of trihalomethanes (Islam 2019). Under the influence of applied electric field this treatment can remove the smallest colloidal particles. Flocs formed by electrocoagulation were larger, more stable, contained less water than flocs formed by chemical coagulation, and electrocoagulation produced smaller amounts of settable and easily de-watered sludge (Mollah et al. 2001; Mollah et al. 2004). The sludge may contain some toxic organic and inorganic components; bacteria and viruses; oil and grease; nutrients such as nitrogen and phosphorus; and heavy metals (Adyel et al. 2013). The main disadvantages are the replacement of sacrificial anodes because the anode dissolves into the solution, and the phenomenon of passivation of the electrode in some cases. Gelatinous hydroxide may tend to solubilize when high conductivity of the wastewater is required (Mollah et al. 2001). The process is influenced mostly by the concentration of the supporting electrolyte, pollutant concentration, electrode material, current density, and reaction time (Malinovic and Pavlovic 2016; Malinovic et al. 2017). Electroflotation is an electrochemical version of flotation in which pollutants float on the water surface with the help of small bubbles of hydrogen and oxygen generated at the cathode and anode, which are created by the electrolysis of water (Eqs. 2.6 and 2.7). The first application of electroflotation was recorded in the early twentieth century for the concentration of minerals from ore. The electroflotation process is able to produce very small bubbles, which enhance the flotation of small particles (Santiago Santos et al. 2018).

2  Electrochemical Treatments for the Removal of Emerging Contaminants



2H 2 O + 2e -  H 2 ( g ) + 2OH -



2H 2 O  O 2 ( g ) + 4H + + 4e -





119

(2.6) (2.7)

The electroflotation process is desirable for removing oil from water or oil emulsions (Genc and Goc 2018), dyes, surfactants, ligands, biological pollutants (Kolesnikov et al. 2019), and printing wastewater (Safwat et al. 2019), among others. Many researches have combined electrocoagulation with the electroflotation process into a single system called electrocoagulation/flotation. An example of a combined process reactor is shown in Fig. 2.5. The removal efficiency of the pollutants depends on the size of the bubbles formed because the smaller bubbles provide a larger surface for particle adhesion. The size of the bubbles generated in the electrofloation reactor mostly depends on the pH value of the wastewater, electrode material, and current density (Chen 2004; Santiago Santos et al. 2018). The generation of hydrogen bubbles is very important for the electroflotation process, so special attention is given to the choice of cathodic materials. The energy consumption and process efficiency also depend on the design of the device, therefore in many studies special attention has been paid to the reactor design, as well as new combinations and methods of electrode placement. Usually the anodes are installed at the bottom of the reactor, but this arrangement does not allow for a good dispersion of bubbles in the wastewater (Kyzas and Matis 2016). On the vertically positioned electrodes, generated bubbles tend to move immediately to the electrode creating rapid coalescence and undesirably big bubbles. Chen et al. (2002) have proposed and tested a new electrode arrangement, where electrodes are on the same plane (open arrangement). The quick dispersion of tiny bubbles is responsible for electroflotation efficiency. In addition to electrocoagulation, electroflotation can be successfully combined with adsorption and membrane processes (Santiago Santos et al. 2018). Yang et al. (2014b) have successfully paired electrooxidation and electroflotation for sludge solubilization. Electroflotation presents several advantages, such as: easy and safe operation, conductivity of the effluent solution is not a crucial parameter, no need for harmful Fig. 2.5  3D model of a laboratory continuous reactor for coupled electrocoagulation/ electroflotation treatment with precipitation chamber and skimming part

120

B. N. Malinović et al.

chemicals, easy automation, and relatively low energy consumption in certain cases (Santiago Santos et al. 2018; Kolesnikov et al. 2019). The energy consumption and the cost of operation (electrode material) may be the disadvantages of electroflotation treatment. If sacrificial electrodes are used, secondary contamination with heavy metals is possible. The principle of electrodialysis was presented at the end of the nineteenth century. The first applications were related to the water desalination process. The basic device (dialyzer) is divided into three or more parts. The dialyzing solution (effluent) flows through the middle compartment, between the two ion exchange membranes. The electrodes are placed on the walls of the electrolyzer. Under the influence of an electric field, cations will move toward the cathode and anions to the anode. Thus, pure water accumulates along the electrodes and a concentrated solution remains between the membranes without any reaction or chemicals. Electrodialysis is a promising electrochemical technology for micropollutant preconcentration and recovery from wastewater. In the study published by Arola et al. (2019), the transport of pharmaceuticals during electrodialysis from synthetic wastewater was quantified. The results showed that the wastewater was up to five times more concentrated. The electrodialysis process was investigated to remove phenolic compounds in the form of phenoxide ions from salty wastewater. More than 90% of the phenol was removed. The specific energy consumption depended on phenol and salt concentration (Wu et al. 2019a). Although many researchers are trying to improve and optimize electrodialysis, there are still some problems in terms of capacity increase (scaling), membrane fouling, and permselectivity (Al-Amshawee et al. 2020). Similar to electrodialysis, electrofiltration is a membrane process that uses an electrical potential gradient to improve performance. The electric field (direct current) facilitates the transport of ions through the pores of the membrane. Practically all colloids and suspended solids, including microorganisms, have either negative or positive electrical charge. Applications of this process are mainly in the treatment of colloidal suspensions and sludge by mechanisms of gravity separation (i.e., sedimentation or flotation) and by filtration (Khosravanipour Mostafazadeh et al. 2016). Electrofiltration could be in media filtration, where the charges of particles would affect their deposition on the surface of collectors (e.g., sand filter), or in membrane filtration (usually inorganic membrane). Electrophoresis, electrolysis, and electroosmosis as main mechanisms improve the membrane efficiency by reducing the fouling phenomena. Important process parameters are the amount of energy for permeate suction, type of membrane, concentration of particles, and electrical field strength (Khosravanipour Mostafazadeh et al. 2016).

2.3.2  Advanced and Combined Electrochemical Treatments Over the last few decades, several new and more effective technologies have been developed for wastewater treatment, such as electrochemical advanced oxidation process or new combinations of separate methods. These processes include

2  Electrochemical Treatments for the Removal of Emerging Contaminants

121

electrochemical oxidation, electro-Fenton, photoelectrochemical and sono-­ electrochemical technologies, and capacitive deionization. The combination of two or more different processes often leads to greater process efficiency, reduced reaction time and increased process economy. In this sense, electrocoagulation has recently often been combined with other treatments. Coupled electrocoagulation-ozone is a treatment in which ozone is introduced into the electrocoagulation reactor. Ozone is a strong oxidant that interacts directly with organic pollutants, or indirectly, whereby it generates free-radical intermediates such as OH∙. The reaction of dissolved iron and ozone produces OH∙ while Fe2+ acts as a catalyst. The main reactions in the combined process are given in Eqs. 2.8, 2.9 and 2.10 (Barrera-Diaz et al. 2018).

Fe 2 + + O3 ( g )  FeO 2 + + O 2 ( g )



FeO 2+ +H 2 O  Fe3+ +OH · +OH -

(2.9)



FeO 2+ +Fe 2+ +2H +  2Fe3+ +H 2 O

(2.10)



(2.8)

The novel trielectrode electrocoagulation-Fenton treatment enables the effective removal of solid suspended materials and persistent organic pollutants with low energy consumption. Electrocoagulation is paired with the electro-Fenton process to allow Fe coagulants to precipitate organic carbon that unnecessarily consumes hydroxyl radicals. Electro-Fenton is more efficient than electrocoagulation in the removal of individual pollutants, but significantly higher costs make it less economical. Coupling microbial fuel cells with electrocoagulation cells was first reported by Safwat (2019) and claims were made that it significantly reduced the operational cost compared to electrocoagulation at treatment model wastewater (mixture of glucose and soluble starch) and real municipal wastewater. Microbial fuel cells can produce electric current through microorganisms which is still below the desired levels, but it can be stored in an external energy storage device and increase voltage for practical use. Graphite fiber felt electrodes were used for both electrodes in the microbial fuel cell, and the anode for electrocoagulation was made of Al, while the cathode was made of stainless steel. According to the research group Ensano et al. (2017), intermittent electrocoagulation process is a promising alternative for treatment of the emerging contaminants in the sense of a simple and highly efficient technology. The feasibility of treating emerging contaminants (i.e., pharmaceuticals) using an intermittent electrocoagulation process was studied in the last decade in order to prevent the electrode passivation phenomenon. A batch reactor with an Al anode and a stainless steel cathode was used. The phenomenon of electrode passivation has limited some applications of electrocoagulation treatment. Application of the “pulsed current” regime, also known as pulsed electrocoagulation (it uses the interactions of electrochemical technology and polarity reversal in an electrical field), yielded higher efficiency during the process (Malinovic et al. 2017).

122

B. N. Malinović et al.

Ensano et al. (2019) also investigated the removal of pharmaceuticals from synthetic wastewater using an electro-membrane bioreactor. A stainless steel cathode and perforated cylindrical Al anode were immersed inside the bioreactor. An ultrafiltration module was placed vertically at the bioreactor and air diffusers were located at the bottom of the reactor. Removal efficiency increases of up to 28% were achieved in the electro-membrane reactor with respect to the conventional membrane reactor due to the enhanced effects caused by the electrochemical processes. This integration of different processes improves membrane reactor performance in terms of pollutant removal and membrane fouling control. The first major research into the electrooxidation of wastewater began after the late 1970s and today it is the most popular electrochemical technique for wastewater treatment and is often named anodic oxidation. Electrooxidation can be direct or indirect. In the process of direct electrooxidation, the pollutants are degraded directly on the anode surface (direct electron transfer to the anode from the adsorbed pollutant) which yields a very poor efficiency. This is theoretically possible at lower potentials than those needed for oxygen evolution reaction (E0 = 1.229 V vs. SHE) (Feng et al. 2016). If the oxidation potential is lower than the potential of water oxidation reaction, the electrodes are likely to inhibit further oxidation due to adsorbed pollutants on the surface. During indirect electrooxidation pollutants are oxidized in the bulk solution through the mediation of some electrochemically redox reagents generated anodically such as chlorine and hypochlorite, hydrogen peroxide, ozone, persulfate and metal mediators. Electrooxidation of organic pollutants can occur indirectly at the anode by physisorbed reactive oxygen species or “active oxygen” (OH∙) that should cause predominantly the complete degradation of organic compounds (Eq.  2.11), and by chemisorbed reactive oxygen species or “active oxygen” (MOx + 1) that forms selectively oxidation products (Eq. 2.12) (Comninellis 1994).

(

R + MO x OH ·

)

z

® CO2 + zH + + ze - + MO x

R + MO x +1 ® RO + MO x



(2.11) (2.12)

Chemisorbed active oxygen modifies the electrode surface and the anode is covered with metallic oxides such as iridium oxides (IrO3/IrO2) and takes part in selective oxidation (complete mineralization is not possible). At the same time, the pollutant oxidation is in competition with oxygen formation. In general, the physisorbed OH∙ group is much more efficient for oxidation of pollutants than MOx + 1. The physisorbed OH∙ is a very strong oxidant with a high standard potential (E0 = 2.80 V vs. SHE) that provides complete mineralization of organic compounds. However, application of high anode potentials is needed, causing competition with the oxygen evolution reaction at the anode. In contrast, direct anodic oxidation is theoretically possible at low potential values (before oxygen evolution potential); however, the reaction rate is usually slow and strongly

2  Electrochemical Treatments for the Removal of Emerging Contaminants

123

dependent on the electrocatalytic activity of the anode (De Battisti and MartínezHuitle 2018). The consequence of this is the application of new innovative material for electrodes such as boron doped diamond (BDD). Many of the AOPs are used to remove emerging contaminants from wastewater because they are eco-friendly and based on the in situ production of the hydroxyl radical (OH∙). Among these emerging AOPs, electrochemical advanced oxidation processes have significant advantages. The electrocatalytic properties of the anodic material determine the mechanism of electrooxidation. An anode with a high oxygen evolution overpotential is required for oxidation of pollutants at high potentials. The anodes are divided into “active” and “inactive” electrodes (De Battisti and Martínez-Huitle 2018). “Active anodes” (low oxygen evolution overpotential) are good electrocatalysts for oxygen evolution (e.g., IrO2, Pt, and RuO2), and “inactive anodes” (high oxygen evolution overpotential) are poor electrocatalysts for the oxygen evolution overpotential. Electrode materials that favor chemisorption, such as Pt, IrO2, or RuO2, will therefore promote selective oxidation. On “inactive anodes” (e.g., SnO2, PbO2, and BDD) there is a weak connection between the electrode and the hydroxyl radical, so oxidation is mediated by OH∙ giving complete degradation. Quasi-free hydroxyl radicals on the BDD electrode surface will exhibit a fast and strong reaction, and favor complete mineralization (Groenen Serrano 2018). Also, direct electrooxidation can occur on these “inactive” anodes. In addition to the hydroxyl radical, other strong oxidation agents can be formed at the anode from dissolved salts or from water. These salts are naturally present or added to wastewater to increase conductivity (as supporting electrolytes) and to reduce the energy consumption. Reactive oxygen species are formed at the electrode such as hydrogen peroxide, ozone, or hydroxyl radical. Oxidants electrogenerated from salts are active chlorine and peroxo-compounds, such as peroxodisulfate and sulfate radicals (Groenen Serrano 2018). Ozone is a strong oxidant with high standard potential (E0 = 2.10 V vs. SHE) used for water treatment and disinfection purposes. Treatments with ozone prevent the formation of unwanted chlorinated by-products. Ozone application is important in the treatment of drinking water, as well as for the pharmaceutical industry, where it is used as a disinfectant in purified water loops. The possibility of direct and continuous production of ozone in water is the main advantage of electrochemical ozone production. There are two important parameters for electrochemical ozone production, and these are the anode material (high oxygen evolution overpotential) and composition of the solution (addition of fluoride as HBF2 or KPF6). The research of Nishiki et al. (2011) with a new electrochemical generator of ozone-dissolved water with a diamond electrode using tap water as a source solution has proven by medical tests that the treatment using ozone-dissolved water was clearly effective to combat the occurrence of inflammatory disease agents. According to Turkay et al. (2017), ozone can form OH. or degrade contaminants directly in water. Authors presented some disadvantages, such as high cost and on-­ site production, but the removal of synthetic dyes, phthalates and some pharmaceuticals has been inefficient when only ozone was used for treatment.

124

B. N. Malinović et al.

Hydrogen peroxide can be produced on both electrodes (anode and cathode) during the electrolysis of wastewater. The production on the cathode requires an oxygen gas supply (Eq. 2.13). O 2 ( g ) + 2H + + 2e - ® H 2 O 2



(2.13)



Moreira et al. (2017) have described H2O2 electrogeneration requirements for the electrode material. Good contact between the cathode, oxygen and wastewater are required. Therefore, porous cathodes like gas-diffusion electrodes and three-­ dimensional electrodes are preferred for H2O2 electrogeneration. Hydrogen peroxide is not a very strong oxidant (E0 = 1.77 V vs. SHE) in acidic medium and is a weak oxidant (E0 = 0.88 V vs. SHE) in alkaline medium. Hydrogen peroxide can only react with reduced compounds, for example sulfur, cyanides, and some organics (aldehydes, formic acid, nitro-organics, and sulfo-organics) (Brillas et al. 2009). Recently, some oxidants have often been used in combination with others or with other treatments. For example, O3 has been used in combination with other processes such as electrooxidation, ultrasound, H2O2, and TiO2. Combined processes usually have a synergistic effect on the degradation of organic and inorganic pollutants and minimize the disadvantages. The use of O3 and H2O2 together is referred to as the peroxone process and has a proven synergetic effect (Turkay et  al. 2017). While O3 and H2O2 oxidize the target pollutants, the reaction between O3 and H2O2 also takes place and forms OH∙ (Eq. 2.14). The peroxone process does not cause the formation of undesirable products as these oxidants do individually. Therefore, the peroxone process is marked as an eco-friendly and effective advanced oxidation process (AOP) in wastewater treatment (Turkay et al. 2017).

O3 ( g ) + H 2 O2 ® OH · + O2 -· + O2 ( g )

(2.14)



Hydroxyl radicals (OH∙) are the strongest oxidant for wastewater treatment in an aqueous solution. Electrogenerated adsorbed hydroxyl radical (M(OH∙)) onto the anode surface (M) can be unnecessarily consumed by inevitable competitive reactions (Eq. 2.15) and (Eq. 2.16) that consume the radical species leading to oxygen evolution (Garcia-Segura et al. 2018).

(

)

M OH · + H 2 O ® M + O2 ( g ) + 3H + + 3e -

(

)

2 M OH · ® 2 M + O2 ( g ) + 2H + + 2e -





(2.15) (2.16)

To achieve the production of greater amounts of M(OH∙), anodes with high overpotential for oxygen evolution should be used to avoid the reactions (Eq. 2.15) and (Eq. 2.16). Table 2.1 gives different oxygen overpotentials for several anode materials. Using BDD anodes, many synthetic and real effluents were completely mineralized with high efficiency, even up to 100% (Martínez-Huitle and Panizza 2018). Based on the achieved results, electrooxidation using a BDD anode is an efficient and effective process for the treatment of highly polluted wastewater.

2  Electrochemical Treatments for the Removal of Emerging Contaminants

125

Table 2.1  Potential for oxygen evolution reactions on different anodes. Anode RuO2 IrO2 Pt Graphite PbO2 SnO2 Pb–Sn (93:7) Ebonex (titanium oxide) Boron doped diamond

Potential (V) vs. SHE 1.47 1.52 1.6 1.7 1.9 1.9 2.5 2.2 2.8

Conditions 0.5 M H2SO4 0.5 M H2SO4 0.5 M H2SO4 0.5 M H2SO4 1 M HClO4 0.5 M H2SO4 0.5 M H2SO4 1 M H2SO4 0.5 M H2SO4

Chemisorption Selectivity ↓

↓

↓

↓

↓

↓

Physisorption

Mineralization

Chen (2004) and Panizza and Cerisola (2009)

Supporting electrolytes have a primary role to provide good conductivity of wastewater. The efficiency of electrooxidation for the treatment of wastewater can be enhanced by the addition of anions, which serve both to yield better conductivity and to provide strong oxidizing agents by electrogeneration, such as active chlorine species. Following the presence or addition of chloride ions to the wastewater to be treated by electrooxidation, the treatment of organic pollutants becomes much more effective. This is due to the additional chemical reactions with the active chlorine in solution and is often called electrochemical chlorination. The oxidation of chloride ions generates active chlorine (Eqs. 2.17, 2.18 and 2.19), which can be present in different forms, e.g. Cl2, ClO− or HClO, depending on the pH.

2Cl - ® Cl 2 ( aq ) + 2e -



Cl 2 ( aq ) + H 2 O ® HClO + H + + Cl -



HClO  H + + ClO -

(2.17)



(2.18) (2.19)

There are various applications of electrochemical chlorination, such as water disinfection, treatment of refractory organics and chloride ions, pharmaceuticals, pesticides, landfill leachates, dyes, etc. The reaction between organic compounds and HClO/OCl− is selective, unlike the reactions with OH. (Groenen Serrano 2018). In the review by Martínez-Huitle et al. (2015), it is reported that several parameters are important for the oxidation of organics by active chlorine: the type of electrochemical reactor, anode material, temperature, pH, current density or potential, flow rate, concentration of chlorides, as well as the nature and concentration of organic contaminants. The pH of wastewater is an important parameter for the electrogeneration of Cl2(aq) which has the highest standard reduction potential (E0 = 1.36 V vs. SHE) indicating that acidic pH conditions will promote a faster oxidation of organic compounds by chlorine active species (de Moura et al. 2014). In electrochemical oxidation mediated by chlorine active species, the “active” anode materials have better

126

B. N. Malinović et al.

performances than the “inactive” anode materials. The “inactive” anode results in the further oxidation of Cl2 and HClO/ClO− to undesired non-oxidizing chlorine species (Garcia-Segura et al. 2015). The next important parameter that influences the oxidation of organic contaminants is the concentration of chloride in the treated water. Increased amounts of chlorine could lead to the formation of organochlorinated species, e.g. chloramines, trihalomethanes, haloacetonitriles, and haloketones, which have been detected during electrooxidation and are formed because of the reaction of organic matter containing different functional groups with active chlorine species. Organochlorinated products exhibit increased toxicity and are usually more recalcitrant than the parent molecules (Moreira et al. 2017). Therefore, it is necessary to define a threshold chlorine concentration that would allow the removal of pollutants, but would generate less of the undesired organochlorinated products. The definition and parametrization of such a condition is probably one of the biggest challenges of electrochemical technologies prior to their scaled-up implementation (Garcia-Segura et  al. 2018). The risk of formation of organochlorinated compounds during electrolysis is the limiting factor for the wide application of this wastewater treatment. Sulfate ions, as well as chlorides, can be present or added to wastewater to provide good conductivity for most electrochemical wastewater treatments. In the case of sulfate, reactive sulfate species are formed and function as secondary oxidants that can enhance the efficiency of the process. Unlike chlorides, sulfates are not considered to be pollutants. According to Groenen Serrano (2018), peroxodisulfate, S2O82−, can be electrogenerated from the oxidation of sulfate at very high potentials (>2 V vs. SHE), as shown in (Eqs. 2.20 and 2.21).

2HSO 4 - ® S2 O8 2 - + 2H + + 2e -

(2.20)



2SO 4 2 - ® S2 O8 2 - + 2e -

(2.21)

The electrochemical reaction leading to the formation of peroxodisulfates strongly depends on the electrode material selected for the oxidation of sulfuric acid (high overpotential value for the oxygen evolution reaction). In sulfuric acid solutions only HSO4− and molecular H2SO4 can react with OH∙ radicals to form sulfate radicals, (Eqs. 2.22, 2.23 and 2.24).

HSO 4 - + OH · ® SO 4 -· + H 2 O

(2.22)



H 2 SO 4 + OH · ® SO 4 -· + H 3 O +

(2.23)



SO 4 -· + SO 4 -· ? S2 O8 2 -



(2.24)

Sulfate radical, SO4–∙, is capable of degrading most persistent organic pollutants because it has a very high standard redox potential (E0 = 2.6 V vs. SHE) compared

2  Electrochemical Treatments for the Removal of Emerging Contaminants

127

to persulfate, S2O82− (E0 = 2.01 V vs. SHE) and hydroxyl radical OH∙ (E0 = 2.8 V vs. SHE). Reaction of S2O82− with many organic pollutants is slow. Electrogenerated reactive sulfate species have a longer life span (30–40 × 10−3 s) in comparison to BDD(OH∙) (10−9  s) and a highly stable redox potential over a wide pH range (pH  2–9), thus they can diffuse into the bulk solution to oxidize pollutants. Electrogenerated SO4–∙ and other reactive sulfate species are considered to be intermediate species in the formation of oxidants, however, many studies have proven their participation in the electrooxidation of organic pollutants (Ganiyu and Martínez-Huitle 2019). The mechanisms of oxidation of organic pollutants by reactive sulfate species have not been extensively investigated. In most studies both hydroxyl radical (predominant oxidant) and SO4– ∙ are generated and participate in the oxidation of the organics. Also, a high concentration of sulfate ions (≥50 mM) are usually required as a prerequisite for the generation of sulfate reactive species, which indicate that a high chemical consumption is required for successful wastewater treatment by electrochemically generated sulfate reactive species (Ganiyu and Martínez-Huitle 2019). Beside the electrooxidation process, electro-Fenton processes are also very popular as an efficient and economical alternative to destroy emerging contaminants and persistent organic pollutants in water. The classical Fenton’s reagent is a mixture of H2O2 and Fe2+. The reaction between these two, i.e. the Fenton’s reaction, causes the formation of the highly reactive OH∙ species (Eq. 2.25) (Moreira et al. 2017). The Fenton process needs high amounts of reagents to produce enough OH∙ which negatively affects the economical aspect, as well as causes the formation of ferric sludge. The rate of wasting reactions (Eq. 2.26) and (Eq. 2.27) is increased as well at high reagent concentrations, resulting in low efficiency (Oturan and Oturan 2018). Other compounds such as HClO are also able to react with Fe2+ producing OH∙. H2O2 can be directly decomposed into OH∙ by other catalytic active metal ions as well: Cr, Ce, Cu, Co, Mn, and Ru (Bokare and Choi 2014).

Fe 2 + + H 2 O2 ® Fe 3+ + OH · + OH -

(2.25)



OH · + H 2 O2 ® H 2 O + HO2 ·

(2.26)



OH · + Fe 2 + ? Fe 3+ + OH -

(2.27)

The electro-Fenton process is based on the synthesis of H2O2 on the cathode in the solution and catalyzes the Fenton reaction by electrochemical regeneration of Fe2+. H2O2 is usually produced by reduction of O2 from compressed air (Eq. 2.28) in an electrochemical cell in an acidic medium. Approx. 10−4 M of ferrous (or ferric) iron should be added to the solution to initiate the Fenton reaction. The Fenton reaction produces ferric iron that can be reduced at the cathode (Eq. 2.29) at the same potential as for the electrogeneration of H2O2 (Oturan and Oturan 2018). In this process, Fenton’s reagent is constantly formed, and reaction (Eq. 2.25) also constantly produces OH∙. This avoids waste generation. The reactors for electro-Fenton

128

B. N. Malinović et al.

can be divided cell (Lin et al. 2014), and lately undivided and open cells have been used due to their simplicity of design and practical advantages (Thiam et al. 2015).

O 2 ( g ) + 2H + + 2e - ® H 2 O 2



Fe 3 + e - ® Fe 2 +



(2.28) (2.29)

Some important benefits of the electro-Fenton process are the reduction of waste compared to classic Fenton, on-site production of H2O2 which is dangerous for handling and storage, high efficiency in removing pollutants, relatively small costs (Oturan and Oturan 2018), affinity for acidic wastewater (pH 2–4), the possibility of treating large quantities of wastewater (Martínez-Huitle et al. 2015), and successful use for the treatment of non-biodegradable or refractory organic compounds (Ferrag-Siagh et al. 2014). The use of heterogeneous catalysts represents a novel electro-Fenton technique – the heterogeneous electro-Fenton process. The use of different heterogeneous catalysts has been investigated for the electro-Fenton process such as bifunctional catalyst with FeOx nanoparticles embedded into N-doped hierarchically porous carbon (Cao et al. 2020), carbon-Fe3O4 (Fernández-Sáez et al. 2019), Cu-doped Fe@ Fe2O3 core-shell nanoparticles (Luo et al. 2019), chalcopyrite, Fe2O3, Fe3O4 (Ltaïef et al. 2018; Labiadh et al. 2019), iron-rich Mobil-type five zeolites (Huong Le et al. 2019), and pyrite (Ouiriemmi et al. 2017). Using pyrite as a heterogeneous catalyst increases process efficiency, and the process is called electro-Fenton-pyrite process. An increase in the efficacy of the treatment of various pollutants with pyrite has been reported. The electro-Fenton-pyrite process recycles solid ferrous sulfide without the addition of iron salt at optimal pH (around 3) and shows slightly better performance than electro-Fenton. In the review by Poza-Nogueiras et al. (2018), the advantages of the heterogeneous electro-Fenton process are highlighted such as extended pH range of wastewater for treatment, reduced quantity of iron hydroxide sludge, the catalyst is easy to handle, safe, and recyclable, while the possible disadvantages of the heterogeneous electro-Fenton process are operating costs and catalyst lifetime. The electro-Fenton process could be used as pre- or post-treatment step coupled with biological treatment methods (Ganzenko et al. 2014). This bioelectro-Fenton process has lesser operation cost and lower energy consumption compared to traditional electro-Fenton technologies. Two chambers of bioelectro-Fenton are separated by a membrane. In the anaerobic anode chamber biodegradable organic matter is oxidized by microorganisms and electrons are released. Electrons move to the cathode by an electrical circuit where electro-Fenton processes occur (reduction of oxygen in the presence of ferrous ions). The bioelectro-Fenton systems are suitable for the treatment of emerging contaminants and persistent organic pollutants in wastewater but there are still some limitations: the high cost of electrodes and membranes, scaling up, existence of residual iron in treated wastewater, pH adjustment before and after treatment, and formation of iron sludge (Li et al. 2018a).

2  Electrochemical Treatments for the Removal of Emerging Contaminants

129

Photoelectrocatalysis belongs to electrochemical advanced oxidation processes and represents the synergy between photocatalytic and electrolytic processes. The process uses a semiconductor (photocatalyst) under light illumination (UV/solar) to generate electron-hole pairs (Eqs.  2.30 and 2.31) with the possibility to degrade many organic compounds by hydroxyl radical OH∙ at its surface. Photoelectrocatalysis is based on the ejection of an electron (eCB) from the valence band (VB) of a semiconductor (saturated), to the empty conductive band (CB), generating a positively charged vacancy or hole (hVB+) (Garcia-Segura and Brillas 2017). The solar photoelectrocatalysis process uses sunlight as an energy source.

Anode + hv ® e CB + h VB+

(2.30)



h VB+ + H 2 O ® OH · + H +

(2.31)

Photoelectrocatalytic systems for removing real pharmaceutical wastewater (Collivignarelli et al. 2019), pesticides (Fernández-Domene et al. 2018), nitrophenols (Fatima et al. 2019), and other organic pollutants (Antolini 2019) used different anode materials. The most widespread photo-anode is TiO2 because it shows good characteristics in terms of stability, low toxicity, and low cost (Collivignarelli et al. 2019; Fatima et al. 2019). UV or solar irradiation of the photoactive electrode is the basis of photoassisted electrochemical treatments. UV radiation sources may be UVA (λ 315–400  nm), UVB (λ 285–315 nm), and UVC (λ  300 nm). The electrochemical processes can lead to passivation of the electrode’s surface due to the adsorption of pollutants, intermediates or products, and they are often limited by mass transport in the system, which significantly reduces the efficiency of the process (Feng et al. 2016). To prevent this, ultrasound (−)))) is often introduced into the system, which most often has a synergistic effect with electrochemical treatment (sonoelectrochemical treatment) by increasing mass transport. According to the researchers Radha and Sirisha (2018), sonoelectrochemical treatment (sonoelectrolysis) causes a homolytic fracture of H2O leading to the formation of the reactive oxygen species (Eq. 2.35) and creates cavitation microbubbles which cause a high degree of mixing and the cleaning of the electrode surfaces.

H 2 O + -))) ® OH · + H ·



(2.35)

Shestakova et al. (2016) performed the optimization of ultrasound frequencies and applied currents. The authors concluded that sonoelectrolysis may be comparable or better than other AOPs for the degradation of some pollutants. Ultrasound paired with the electro-Fenton process is known as sonoelectro-­ Fenton whereby wastewater is constantly treated with ultrasound. Research from the authors Babuponnusami and Muthukumar (2012) examined the performance of Fenton, electro-Fenton, sonoelectro-Fenton, and photoelectro-Fenton processes on the degradation of phenol. The results showed better performance with coupled processes of UV irradiation and electrolysis with Fenton’s reagent. Degradation efficiency was: photoelectro-Fenton > sonoelectro-Fenton > electro-Fenton > Fenton. The main disadvantage of these paired treatments is their problematic transformation into full scale industrial application. Capacitive deionization is an electrochemical technique for removing dissolved, charged species (especially salts) from aqueous solutions using electroadsorption of ions at the surface of electrically charged electrodes. It is an alternative to membrane-­ based technologies with some advantages: low operational cost, energy efficiency, and less water rejection (Gupta et al. 2019). The most commonly used electrodes are porous carbon materials, carbon aerogel, carbon cloth, carbon nanotubes, graphene, and carbon fibers. The most common use of capacitive deionization is for the desalination processes. Studies on capacitive deionization have increased significantly in the last 10 years. Recently, it has been increasingly used to remove other pollutants such as: textile cationic dyes (Senoussi and Bouhidel 2018), valuable heavy metals and nutrients (nitrate/phosphate), phosphorus (Zhang et al. 2019), and nitrate (Pastushok et al. 2019).

2  Electrochemical Treatments for the Removal of Emerging Contaminants

131

2.3.3  E  lectrodes for Electrochemical Removal of Contaminants Electrode material selection has one of the most important roles in electrochemical treatment. General requirements for electrode material are good physical and chemical stability, electrical conductivity, catalytic activity and selectivity, lifetime, and reasonable price. In this section, we present the properties and efficacy of the most commonly used electrode materials for different electrochemical treatments described in the previous sections. The most used cathode materials for electrochemical reduction are Cu, Ag, Pt, Pd, Au, Ru, Rh, Ir, and Ni (Feng et  al. 2016; Rondinini et  al. 2018). Electrocoagulation is the most efficient and most studied with Al and Fe sacrificial anodes. The most commonly used cathodes for electroflotation are stainless steel, nickel and titanium, while graphite, PbO2, Pt, and dimensionally stable anode are used for anodes. In the process optimization, electrodes can be placed into different positions (Santiago Santos et al. 2018). The study of electrode material in the application of electrooxidation for wastewater treatment at the end of the twentieth century was largely concerned with platinum and dimensionally stable anode made of metal oxide. Platinum electrodes have good corrosion resistance in aggressive media and low oxygen evolution overpotential (Table 2.1). Some studies have shown their selective conversion of pollutants resulting in different by-products (Feng et al. 2016). Dimensionally stable anodes are made by coating a substrate, usually titanium, with metal oxide or mixed metal oxides (Ti, Ru, Ir, Sn, Ta, and/or Sb). The use of mixed RuO2 and TiO2 based anodes is the most significant in the chlor-alkali industry (from the late 1960s), but during the past two decades these anodes were used in wastewater treatment to achieve low current efficiencies and long electrolysis times to completely remove pollutants (Feng et al. 2016). The dimensionally stable anodes such as RuO2 and IrO2 can be used efficiently to produce active chlorine for pollutant degradation. Ti/IrOx − Ta2O5 electrodes do not achieve total mineralization of organics, and Ti/SnO2 anodes are more effective, but they have a limited service life (Martínez-Huitle et al. 2015). Carbon-based anodes have also been investigated (activated carbon, activated carbon fiber, glassy carbon, graphite) because they are cheap, have a large surface area, good conductivity and excellent adsorption capability. Their most significant application is in the form of three-dimensional electrodes (Martínez-Huitle et  al. 2015). The main drawback is the appearance of surface corrosion and passivation (Feng et al. 2016). A large number of studies have demonstrated better efficacy and chemical resistance of “inactive” anodes such as PbO2, SnO2, and BDD (Martínez-Huitle et al. 2015). PbO2 anodes have good conductivity and high oxygen evolution overpotential (Table 2.1) allowing hydroxyl radicals to be generated. Their use is limited due to the short life span and possible generation of highly toxic Pb2+, which raises concerns about secondary wastewater pollution. Pure SnO2 is not suitable for electrodes because of its low conductivity, so it is used as a coating on some base metal,

132

B. N. Malinović et al.

most commonly titanium. The electrochemical properties and the lifetime depend mostly on the method of preparation. Titanium anodes coated with antimony doped SnO2 have high oxygen evolution overpotential (1.9 V vs. SHE) and excellent performance in electrooxidation of organic pollutants. Ti/SnO2-Sb has a relatively short service life due to the formation of a passive layer between the substrate and the coating, and many improvement attempts such as doping with rare earth metals and carbon nanotubes have been reported (Feng et al. 2016). Novel carbon-based materials with outstanding electrochemical properties are diamond electrodes. The research and application of BDD electrodes in electrochemical advanced oxidation processes was very intensive over the last two decades and has been found to represent the best anode material and allow complete mineralization of many contaminants, for example ammonia, cyanide, phenols, aniline, hydrocarbons, dyes, surfactants, drugs, pesticides, etc. (Martínez-Huitle et al. 2015). Diamond is the hardest material available on Earth, has a high resistance to thermal shock, high thermal conductivity, wide band gap, high electronic mobility, and chemical inertia (Chen 2004). Introduction of dopant atoms (usually boron) in the structure improves electrical conductivity. Diamond electrodes are usually obtained by chemical vapor deposition or high pressure high temperature methods. Electrocatalytic properties of BDD extensively depend on the boron content, substrate material, sp3/sp2 ratio, morphologic factors and crystallographic orientation, the presence of impurities, and the thickness of the BDD layer. Silicon is the most suitable support, but diamond has been grown on metals, such as W, Mo, Ti, Nb, etc. (Nidheesh et al. 2019). The surface of the BDD anode is capable of producing large quantities of various reactive oxygen species, such as hydroxyl radicals, peroxodisulfate, ozone, peroxodicarbonate, and peroxodiphosphate. The reactive oxygen species most commonly produced on BDD is hydroxyl radical (BDD(OH.)) which allows for complete removal of organic matter (mineralization) whereby the reactions are irreversible and the reactor does not require the use of a separator. The electrochemical oxidation can proceed by the transfer of electrons or by the transfer of oxygen atoms, but both pathways can take place simultaneously. For generating reactive oxygen species, a high oxygen overpotential is required, otherwise a large amount of current will be unnecessarily consumed by a side-reaction of oxygen evolution, which reduces the efficiency of the process (Martínez-Huitle et al. 2015). Hydroxyl radicals are active only very close to the BDD surface (95% at neutral pH DT: 1.5 h

Degradation efficiency BPA: >90% TOC: lower than 50% after 1.5 h DT: 12 min for 10 mg/L; 1.5 h for 150 mg/L

Burgos-­ Castillo et al. (2018)

Zhao et al. (2018)

Cao et al. (2018)

Divyapriya et al. (2017)

References Chmayssem et al. (2017)

136 B. N. Malinović et al.

In-situ metal-free electrochemical advanced oxidation

10 mg/L in 0.05 M Na2SO4

Electrooxidation 0.02 to 0.1 mmol/L in 10 to 50 mmol/L NaCl Electrolyte: Electrooxidation and 17 mM chloride biodegradation

Electrocatalytic synergism Cr(VI)/BPA system

Process Electrocatalytic oxidation

20.00 mg/L BPA

Sample characteristic 15 mg/L BPA

NA

UC

UC

Three-­ dimensional electrode reactor

Electrodes/operational parameters Mode: GM Anode: DSA (20 cm2) Cathode: SS Particle electrodes: Nitrogen-­ doped graphene aerogel (NGA) pH: 3.0 Cell voltage 20 V Ratio of NGA mass to solution volume 2.25 mg/L NA Polypyrrole/reduced graphene oxide aerogel 80.00 mg/L Cr(VI) pH: 3.0 to 11.0 0.05 Mode: PM WE: BGA-GDE (3.0 cm × 2.0 cm) CE: Pt (2.5 cm × 2.5 cm) E: 0 to −1.4 V (vs. SCE) j: 4.5 mA/cm2 pH: 3.0 to 9.0 0.4 Mode: GM Anode: BDD Cathode: SS (1 cm2) pH: 3.0 to 11.0 0.3 Mode: GM Anode: Ti-TiO2/IrO2/RuO2 Cathode: Ti pH: 9.0 j: 20 mA/cm2

Electrochemical reactor V Configuration [L] 0.2 Three-­ dimensional electrode reactor

Zhang et al. (2018)

BPA: 98% DT: 0.5 h

COD: 91.9% TOC: 56% DT: 10 min

BPA: Complete DT: 1 h

(continued)

Aravind et al. (2019)

Li et al. (2017)

BPA: Almost complete to Wu et al. (2019b) 90% after five cycles TOC: 90% after five cycles DT: 1 h

References Chen et al. (2017)

Degradation efficiency BPA: 90% COD: 85% DT: 0.5 h

2  Electrochemical Treatments for the Removal of Emerging Contaminants 137

Electro-assisted catalytic wet air oxidation

10 mg/L in 0.2 M Photoelectrocatalysis NaCl

50 mg/L in 0.050 M Na2SO4

Electrooxidation 10 mg/L in nitrate/ bicarbonate/ carbonate; 0, 0.1, and 0.35 mM

Sample characteristic Process 100 mg/L in 1 M Electrooxidation Na2SO4

Table 2.2 (continued)

DC

UC

0.13 Mode: PM WE: MnO2/GF (5 mm diameter) CE: Pt (0.5 mm diameter) E: +1.0 V (vs. SCE) Room temperature pH: 7.0 NA Mode: GM Anode (photo): WO3 (5 × 6 cm, 2.2 mm) Cathode: dopamine modified carbon felts (CF-DPA 5 × 6 cm) LS: 300 W Xe pH: 10.8

Electrochemical reactor V Configuration [L] Electrodes/operational parameters NA 0.1 Mode: GM Anode: BDD with/without Ce(IV) addition Cathode: SS Current density 0.5 A/cm2 75 °C pH: 3.0 to 10.0 NA 1 Mode: GM Anode: Ti/IrO2 Cathode: Ti (130 cm2) j: 3 and 5 mA/cm2

BPA: 92% TOC: 92% DT: 2 h

BPA: almost complete in the absence of anions; 40% in the presence of NO3−, 42% in the presence of bicarbonate and 80% in the presence of carbonate DT: 5 and 20 min BPA: 90% TOC: 12% DT: 3 h

Degradation efficiency BPA: Complete DT: 20 min

Xiao et al. (2019)

Sun et al. (2019b)

Jo et al. (2016)

References Huang et al. (2017)

138 B. N. Malinović et al.

Process Photoelectro-chemical

Photoelectrocatalysis g-C3N4/ Fe2O3

Photoelectrocatalysis

0.5 mM in 0.05 M Na2SO4

Electrolyte: 150 mg/L Na2SO4

5 mg/L BPA Photoelectro-chemical pH adjusted with NaOH and H2SO4

Sample characteristic Electrolyte: 0.5 M Na2SO4

Electrochemical reactor V Configuration [L] Electrodes/operational parameters NA 0.2 Mode: PM (CV) WE: Cu-doped WO3 CE: Pt 0.0 to 1.5 V (vs. Ag/AgCl) LS: solar simulator 100 W NA 0.1 Mode: PM WE: Co3O4/BiVO4 based nanostructured photo-anode films CE: Pt LS: Xe lamp with a UV cut-off filter (λ > 420 nm) pH: 7.06 UC 0.25 Mode: GM Anode and cathode: Perforated graphite plates (5 cm × 5 cm × 1 cm) LS: 455 nm, 0.47 mW/cm2 j: 100 mA/cm2 pH: neutral NA Mode: PM Three-­ WE: Graphene oxide-polyaniline/ dimensional titanium dioxide composite network structure hydrogels CE: Pt LS: 320 nm Cui et al. (2018)

BPA: 100% DT: 4.5 h

(continued)

Yan et al. (2019)

Li et al. (2019)

BPA: 96% DT: 2 h

BPA: 92% (1 h) TOC: 70% (1 h); 91% (2 h) DT: 1 h/2 h

References Goulart et al. (2019)

Degradation efficiency BPA: 80% TOC: 75% DT: 8 h

2  Electrochemical Treatments for the Removal of Emerging Contaminants 139

Electrochemical filtration

Electrochemical microfiltration

Electro/Mn(II)/sulfite process

50 ppm in 0.1 mol/L Na2SO4

10 μM in 5 mM sulfate

Process Electrochemical filtration

0.5, 1 and 10 mg/L in 10 mM NaCl

Sample characteristic 1 to 100 mg/L in 1 to 100 mM NaCl

Table 2.2 (continued) Electrochemical reactor V Configuration [L] Electrodes/operational parameters NA NA Mode: GM Anode: MWNTs (20–30 nm OD, average length 10–20 mm and electrical conductivity >100 S/ cm) and BMWNTs (20–40 nm OD, average length 50 mm) Cathode: SS E: −1 to 1.5 V (vs. Ag/AgCl) pH: 3.0 and 9.0 E: 2 and 3 V DC NA Mode: GM Anode and cathode: MWNTs E (DC): 2 and 3 V pH: 5.7–6.0 NA 0.25 Mode: GM Anode: Tubular CCM Cathode: Ti 2.0 V DC 0.2 Mode: GM MMO electrodes pH: 6.0

Pan et al. (2019)

Jia et al. (2019)

BPA: 94%

Bakr and Rahaman (2019) BPA: 95% TOC: 92% DT: 2 h BPA: 97% COD: 90% DT: 0.88 min

References Bakr and Rahaman (2017)

Degradation efficiency BPA: Almost complete for 1 mg/L

140 B. N. Malinović et al.

Electrochemical and photocatalytic

Process Electro-enhanced heterogeneous activation of PDS

Electrochemical reactor V Configuration [L] Electrodes/operational parameters UC 0.2 Mode: GM Anode: Ti/IrO2-RuO2 Cathode: Ti Catalyst: Mn-Zn ferrite (Mn0.6Zn0.4Fe2O4) pH: from 3.0 to 9.0 1.0 mM peroxydisulfate (PDS), 0.5 g/L Mn0.6Zn0.4Fe2O4, and 3.36 mA/cm2 current density UC 0.5 Mode: GM Anode: Zn sacrificial Current intensities: 50–100 mA References Deng et al. (2018)

Alikarami et al. (2019)

Degradation efficiency BPA: 89%

BPA: 84% DT: 2 h

Abbreviations: BGA boron-doped graphene aerogel, BMWNT boron-doped multiwalled carbon nanotube, BPA bisphenol A, CCM coal-based carbon membrane, CE counter electrode, CF carbon fiber, CV cyclic voltammetry, DC divided cell, DT duration/time of experiment, GDE gas diffusion electrode, GM galvanostatic mode, LS light source, MMO mixed metal oxide, MWNT pristine graphite multi walled carbon nanotube, NA not available, NGA nitrogen-doped graphene aerogel, OD outer diameter, PM potentiostatic mode, QEEG quinone-functionalized electrochemically exfoliated graphene, SCE saturated calomel electrode, SS stainless steel, UC undivided cell, WE working electrode

10–30 mg/L in 0.03–0.1 M electrolyte

Sample characteristic 0.1 mM in 50 mM Na2SO4 2  Electrochemical Treatments for the Removal of Emerging Contaminants 141

142

B. N. Malinović et al.

was monitored by high-performance liquid chromatography with UV detection (HPLC-UV). Total organic carbon (TOC), COD, and biological oxygen demand (BOD) were measured as well. Hydroxyl radical production by electro-Fenton reaction was followed by its reaction with dimethylsulfoxide and subsequent determination of the product. In the optimal experimental conditions, the conversion of H2O2 to OH. was >90% and bisphenol A at low concentrations up to 10 mg/L was degraded in less than 12 min, however, the necessary time increased to over 1.5 h at bisphenol A concentrations up to 150  mg/L.  Authors identified some competing oxidation reactions of bisphenol A at dimensionally stable anode, but subsequent experiments showed that they had a negligible effect on bisphenol A removal. Although there was no attempt to identify the transformation products, some were obviously present even at >99% bisphenol A removal, as seen from the residual TOC.  Authors calculated the percentage of bisphenol A mineralization from the COD/TOC ratio and it was found to be 50% after 1.5 h of treatment at the optimal conditions and low bisphenol A concentration (10 mg/L). The remaining transformation products were speculated to be small carboxylic acids. Additionally, authors reported increased biodegradability of the treated bisphenol A solution, again at lower concentrations, as estimated from BOD/COD ratio. Based on this calculation, they concluded that a scaled-up system would be appropriate as the pre-treatment process before the biological treatment of wastewater. However, this is a far-fetched conclusion as typical wastewater contains an enormous number of organic substances that would consume hydroxyl radicals as well. Divyapriya et  al. (2017) prepared a quinone functionalized electrochemically exfoliated graphene and used it as an electrode material either alone or loaded with Fe3O4 nanoparticles to various weight ratios from 20:1 to 10:5. The prepared electrodes were used as working electrode (cathode) in a system consisting also of a Pt sheet counter electrode and Ag/AgCl reference electrode. Linear scanning voltammetric experiments were performed in the potential range of 0 to −1.5 V in Na2SO4 to study the formation of H2O2 and hydroxyl radicals at different working electrode materials and different pH of the solution (pH 3.0, 5.0, 7.0, 8.5, and 11.0). H2O2 was formed at all types of electrode material and the authors postulated the mechanism of its formation via the 2-electron mechanism at quinone moieties on the surface of electrode. However, H2O2 formation was increased in the strongly alkaline medium compared to acidic due to competitive reactions (H2O2 and H+ reduction) at acidic pH.  H2O2 formation was increased at higher (more negative) potentials but was diminished at composite materials with Fe3O4 in the order of increasing Fe3O4 ratio. The main reason were fewer available quinone moieties on the electrode surface, confirming the postulated mechanism of H2O2 formation. However, monitoring of OH. radicals revealed that the highest concentration was produced by the Fenton reaction on the composite electrode with a 10:1 (functionalized graphene:Fe3O4) ratio, approx. four-to-five-fold compared to functionalized graphene only. At higher Fe3O4 loading, OH. production was less efficient due to diminished H2O2 formation. The acidic medium favored the production of OH. because of higher concentration of dissolved Fe2+ ions, although the authors concluded from the H2O2:Fe2+ ratio that the Fenton reaction predominantly occurred at Fe3O4 nanoparticles. Finally,

2  Electrochemical Treatments for the Removal of Emerging Contaminants

143

bisphenol A was used as a test pollutant to study its removal with functionalized graphene working electrode or with composite working electrode. At the optimal conditions for OH. production, the removal of bisphenol A was >99.99% in 1.5 h, while a significantly longer time was needed for similar results with other electrode materials or pH.  Kinetic constants (pseudo-first order) were calculated as well. Authors followed the decomposition of bisphenol A by HPLC-UV and identified one degradation product at electro-Fenton conditions, 4-isopropenyl phenol, by liquid chromatography coupled with mass spectrometry (LC-MS) – Fig. 2.7. However, it is not reported if this transformation product is further degraded with the longer time of experiment. An interesting study in the electro-Fenton process was reported by Cao et  al. (2018), again using bisphenol A as a model pollutant. They applied spent graphite from the lithium-ion batteries as cathode material in the electrochemical system also containing a Pt sheet as anode. Spent graphite powder was used as a raw powder or leached either with acid, or with acid and alkali. These three types of materials were compared for their performance in an electro-Fenton system for the degradation of bisphenol A (50  mg/L), which was monitored by HPLC-UV and COD.  The best performance was achieved with the acid-leached spent graphite powder: 100% bisphenol A degradation in 70 min and 87% COD reduction in 4 h. However, there was no attempt to identify the transformation products as the focus of the research was on the utilization of solid waste, i.e. batteries.

Fig. 2.7  Currently known electrochemical degradation pathways of bisphenol A reported in literature. (Bakr and Rahaman 2017; Chen et al. 2017; Divyapriya et al. 2017; Li et al. 2017; Burgos-­ Castillo et al. 2018; Zhang et al. 2018; Bakr and Rahaman 2019; Aravind et al. 2019; Pan et al. 2019) Masses of individual compounds in Da account for stable isotopes with lowest masses, e.g. M 35 Da for Cl

144

B. N. Malinović et al.

Zhao et al. (2018) prepared an iron-copper-embedded carbon aerogel (FeCuC) and compared it with carbon aerogels embedded with Fe or Cu only, or without any addition. The main focus of the study was to investigate the mechanism of oxygen reduction reaction leading to the formation of H2O2 (2-electron process) or H2O (4-electron process). The calculation of kinetic constants for various electrode materials revealed that FeCuC was the most effective for the production of H2O2 and its subsequent cleavage to hydroxyl radicals by heterogeneous Fenton process taking place on the surface of cathode because of embedded Fe and Cu ions. The process was tested first by elucidation of efficiency and mechanism of dimethyl phthalate degradation and subsequently by degradation of bisphenol A and two other emerging contaminants as well. For bisphenol A, 97% removal was achieved in 4 h as deduced from TOC analysis. Finally, the process was applied to five real samples of sanitary wastewater and was shown to effectively lower TOC and total nitrogen content to meet the Chinese discharge standards in 2 h of treatment. Several electrochemical advanced oxidation processes were compared for the degradation of bisphenol A in the study by Burgos-Castillo et al. (2018), namely oxidation with electrochemically generated H2O2, electro-Fenton, photoelectro-­ Fenton, and solar photoelectro-Fenton. Two different electrolytes were studied: acidified solutions (pH 3) of Na2SO4 or NaCl+Na2SO4. For Fenton processes, FeSO4 was added as well. Experiments were performed in an undivided cell using a BDD anode and air-diffusion cathode on which H2O2 was generated. For photo-processes, UVA light at λmax 360 nm or direct irradiation with sun (summer, Mediterranean) was applied. Bisphenol A degradation was monitored by HPLC-UV and TOC measurements, while transformation products were followed either by ion-exclusion chromatography or by GC-MS (Fig. 2.7). The efficiency of the treatment was in the following order: electrochemically generated H2O2 95% removal of bisphenol A and >92% removal of TOC was achieved in 2 h. No identification of the remaining transformation products was done, although they are expected to be chlorinated species due to the confirmed production of Cl. and ClO. radicals on photo-anode. Goulart et  al. (2019) prepared a modified WO3 (Cu-doped) photo-anode with different percentages of Cu addition from 0.5 to 2.0% (w/w) and studied photoelectrocatalytic degradation of bisphenol A in a glass cell with Pt counter electrode and Ag/AgCl reference electrode. Working electrode was irradiated by visible light and H2O2 was used as an additional oxidant. Removal of bisphenol A was followed by UV/visible spectrophotometry and additionally by cyclic voltammetry with electrochemical sensor (NiO-multiwalled carbon nanotubes-glassy carbon), capable also of detecting phenolic intermediates of bisphenol A degradation. Moreover, TOC was measured. Different parameters were optimized: doping of photo-anode material, initial bisphenol A concentration, pH, current density, and H2O2 concentration.

2  Electrochemical Treatments for the Removal of Emerging Contaminants

149

The generated photocurrent was highest with the 1.0% doping with Cu, i.e., six times the current produced with the WO3-only electrode. At the optimal parameters, the photoelectrocatalytic system was capable of removing 80% of bisphenol A and 75% TOC in 8 h of operation. Li et al. (2019) investigated yet another photoelectrochemical system and tested its efficacy for bisphenol A degradation. They prepared Co3O4/BiVO4 nanostructured films by electrospinning. This film was used as photo-anode (working electrode) in a cell consisting also of Pt wire as counter electrode, saturated calomel reference electrode, and peroxymonosulfate solution as electrolyte. The cell was illuminated with a Xe lamp at λ > 420 nm and the removal of bisphenol A was monitored by HPLC-UV. The same system, but with a BiVO4 only photo-anode, was also tested and found to be much less effective: cca. 50% of bisphenol A removal in 2 h vs. 96% removal in the same time with the new photo-anode. Different modes of operation were tested as well: electrooxidation only, photochemical only, and combined with 21%, 55%, and 96% of bisphenol A removal, respectively. Electron spin resonance and tests with radical quenchers were used to elucidate the principal agent for degradation and the order hVB+  >  OH.  >  SO4–.  >  O2–. was established. Authors didn’t investigate the transformation products of bisphenol A or its mineralization (e.g. by TOC or COD). Yan et al. (2019) designed a photocatalytic – sulfate radical based AOP for the degradation of bisphenol A. The process was catalyzed by a visible light responsive catalyst (g-C3N4/Fe2O3) and enhanced by electro-assisted persulfate activation using a low-voltage bioelectricity supplied by a microbial fuel cell. Experiments were conducted in a glass photo-electrochemical reactor with two perforated graphite plates as electrodes. The reactor was surrounded by LED lamps (455 nm and broadband light at 500–600 nm). Under optimal conditions, 92.2% of bisphenol A was degraded in 1 h, but the process had to be prolonged to 2 h to achieve 91.0% TOC removal. Two other catalysts were tested as well (g-C3N4 and Fe2O3) at the otherwise identical conditions, and their performance was significantly lower: 21.9% and 68.5% of bisphenol A removal in 1 h, respectively. Since bisphenol A concentration was followed by HPLC-UV, no intermediates were identified. The g-C3N4/Fe2O3 catalyst could be recycled 5 times, afterwards its activity decreased significantly due to the iron leaching. Quenching experiments with different quenchers/scavengers revealed that the principal reactive species responsible for bisphenol A degradation was SO4-.; followed by singlet oxygen and the holes generated in the photocatalyst. This study is especially interesting because of the use of renewable sources such as visible light and microbial fuel cells. Another useful material for photo-electrodes was found to be 3D-structured reduced graphene oxide-polyaniline/TiO2 composite synthesized by Cui et  al. (2018). Prepared working electrode was used together with Pt counter electrode and saturated calomel reference electrode in an undivided cell with Na2SO4 as electrolyte. Working electrode was irradiated by a Hg lamp at λ 320 nm. Working electrodes made of TiO2 only or polyaniline/TiO2 were also tested, as well as operation without irradiation and without electric current. The system was applied for the degradation of different pollutants: phenol, bisphenol A, 2,4-dichlorophenol, and

150

B. N. Malinović et al.

coking wastewater. Individual pollutant removal was followed by HPLC-UV, and overall removal by TOC measurements. Removal of all three pollutants was the most effective with the working electrode made of composite material and in photoelectrochemical mode of operation. Under these conditions, removal of bisphenol A was 100% after 4.5 h. The addition of different radical scavengers showed that the active species were h+ and O2-., while OH. played a minor role. No transformation products were identified. Although electrofiltration is considered to be a classical electrochemical treatment and not an electrochemical advanced oxidation process, it is nevertheless still in the development and finds useful applications in pollutant treatment. Bakr and Rahaman (2017) applied pristine and boron-doped multiwalled carbon nanotubes for the removal of bisphenol A by electrofiltration. Nanomaterial was used as an anode electrochemical filter in a bench-scale electrochemical filtration cell, while the cathode was made of perforated stainless steel. Different process parameters were optimized: different direct current potentials or no potential (conventional filtration); flow rate; pH; and concentration of NaCl as electrolyte. Removal of bisphenol A was monitored by UV/visible spectrophotometry, but LC-MS analysis was performed as well to identify the transformation products (Fig. 2.7). By performing the process at different conditions, authors were able to identify superoxide species as the dominant factor in the degradation of bisphenol A. Moreover, they identified several transformation products such as hydroxylated phenol derivatives (i.e., hydroquinone) which further decomposed with longer residence time to short-chain aliphatic compounds. In the presence of NaCl, some phenolic intermediates and even final-stage aliphatic compounds were chlorinated, although it is claimed that the latter were non-toxic. However, no toxicity studies were done. The same research group also tested bisphenol A degradation (together with another emerging contaminant, ibuprofen) with an improved cross-flow electrochemical filtration system using a superconductive blended membrane made of multiwalled carbon nanotubes and buckypaper (Bakr and Rahaman 2019). Identical membranes were used as anodes and cathodes, separated by a porous Teflon rubber separator to prevent short-circuiting. The electrolyte was NaCl. UV/visible spectrophotometry and LC-MS were used to monitor the removal of pollutants and to identify the transformation products, respectively. The purpose of the study was to research the performance of the proposed system. Due to the cross-flow design, the residence time within the membrane was relatively long (18.3 s), which allowed for efficient degradation of both pollutants. Subsequently, the process was tested on synthetic wastewater as well. Both pollutants degraded to aliphatic organic acids as evidenced by LC-MS, with the common intermediates hydroquinone and p-­ benzoquinone (Fig. 2.7). No chlorinated transformation products were observed. Similarly, Pan et al. (2019) designed an electrochemical microfiltration process based on the coal-based carbon tubular membrane as anode and Ti plate as cathode. The system was tested on a solution of bisphenol A and Na2SO4 as electrolyte. Several process parameters were optimized: bisphenol A concentration, conductivity of the solution, applied voltage, residence time, and continuous or batch mode.

2  Electrochemical Treatments for the Removal of Emerging Contaminants

151

Bisphenol A removal was monitored by UV-visible spectrophotometry. Additionally, transformation products were identified by LC-MS, and COD was also measured. Under the optimized conditions, the removal of bisphenol A was 97% and COD reduction was 90% with residence time of only 0.88 min. Membrane fouling was still minimal after 7 h of operation. LC-MS analysis revealed the usual transformation products: p-benzoquinone and 4-isopropylphenol (Fig. 2.7). With further degradation, oxalic, maleic, and acetic acids were produced and subsequently mineralized. The mechanism of bisphenol A removal was elucidated by adding either hydroxyl or superoxide radical scavengers, while oxalic acid was used as the direct oxidation probe. Based on these experiments, authors concluded that the oxidation proceeded both directly at the anode (membrane) and indirectly via OH. radicals, while O2–. radicals played a minor role. Finally, use of an electrochemical system may not be applied per se to degrade pollutants, but as an aid to another AOP process. Three papers dealing with this type of application have been identified. Jia et al. (2019) enhanced the AOP of Mn(II)-catalyzed process for SO4–. radical production from sulfite by maintaining Mn(II) catalytic activity via constant oxygen production using the electrolysis of water. The electrolytic cell consisted of two mixed-metal oxide electrodes and was operated in galvanostatic mode. The combined electrolysis-AOP process was tested on bisphenol A as the model pollutant and compared to AOP without electrolytic production of O2. While in the latter, 46.3% of bisphenol A was removed in 40  min, combination with electrolysis increased the removal of bisphenol A to 94.2% in the same time. An experiment with divided cell (Nafion membrane) revealed that most of the bisphenol A removal occurred on the anodic side. Assays with radical scavengers showed that the SO4–. radical was the dominant species for bisphenol A removal. However, no transformation products were identified, because the removal of bisphenol A was monitored by HPLC-UV. The developed process was subsequently tested on several other water contaminants with equally good removal efficiencies. Similarly, Deng et  al. (2018) designed a combined electro-AOP process. A Mn0.6Zn0.4Fe2O4 catalyst was prepared from spent alkaline Zn-Mn batteries and applied for heterogeneous electro-activation of peroxydisulfate. Experiments were performed in a batch mode and at a constant current with Ti/IrO2-RuO2 anode and Ti cathode. Bisphenol A was used as a model pollutant and several process parameters were investigated: initial pH, current density, peroxydisulfate concentration, and catalyst dosage. Experiments were performed with/without catalyst, peroxydisulfate, or electric current. As expected, the most effective bisphenol A removal was achieved with the combination of all three. Analysis with electron paramagnetic resonance confirmed that both surface-bound sulfate radicals and hydroxyl radicals were responsible for bisphenol A degradation. However, degradation products were not studied because only HPLC-UV was used to determine bisphenol A concentration. In the work by Alikarami et  al. (2019), a combination of electrochemical and photocatalytic processes was applied to degrade bisphenol A as a model pollutant. The concentration of bisphenol A was monitored by HPLC-UV. Also, a bio-toxicity

152

B. N. Malinović et al.

assay was performed, consisting of measuring the specific oxygen uptake rate by the active biomass in the bio-solids sample. The actual degradation of bisphenol A occurred because of photocatalytic generation of OH. radicals on ZnO nanoparticles that were electrochemically generated from the sacrificial Zn anode. Separate processes, i.e. photolysis and electrochemical, were much less effective, while with the combined treatment, 84% removal of bisphenol A was achieved in 2 h. The removal efficiency was further enhanced by the addition of “solid” H2O2 (in fact, sodium percarbonate) to increase OH. production, or addition of peroxymonosulfate to produce SO4–. radicals. Under these conditions, 90.6% and 97.0% of bisphenol A was degraded in 2 h, respectively. The rate of degradation could be further increased by the application of ultrasound (100% in 1.5 h). Bio-toxicity assay also revealed that the least toxic effluent was produced by the hybrid treatment process. To summarize, some notable novelties in the field of electrochemical treatments for the removal of bisphenol A from aqueous systems have been identified in the reviewed papers. Promising technologies are used for electrochemical treatment of emerging contaminants in water, such as electrochemical filtration in multiwalled carbon nanotubes (Bakr and Rahaman 2017; Bakr and Rahaman 2019) and employment of new materials such as carbon aerogels as novel mesoporous carbon materials with a three-dimensional net structure of the only aerogel with electrical conductivity (Chen et  al. 2017) and graphene oxide-polyaniline/titanium dioxide composite (Cui et al. 2018). Also, the recycling of used materials from batteries was reported in the case of spent graphite from lithium-ion batteries where strong application potential was shown (Cao et al. 2018) and in the case of spent raw material from alkaline Zn-Mn batteries as well (Deng et al. 2018). Toxicity studies on transformation products formed during electrochemical treatments are scarce. Aravind et al. (2019) used biodegradability and in silico toxicity analysis for determination of bisphenol A electrooxidation products. Indirect electrochemical oxidation was employed under galvanostatic mode in a membrane-less electrochemical cell (see Table  2.2) at Ti-TiO2/Ir2O2/RuO2 anode. Bisphenol A removal was monitored by HPLC-UV analysis. Changes in biodegradability were determined by naphthalene-degraded consortia of microorganisms utilized to degrade aromatic intermediates in textile effluent. COD and bacterial growth (by optical density) were checked periodically for non-treated and treated solution of 100 ppm of bisphenol A. Experiments were conducted in a batch mode (100 mL); using the addition of microbe-active polyurethane foam (1 cm2, left 10 days to allow the biofilm to develop); and using a partially packed bed reactor with a capacity of 450 mL loaded by polyurethane foam. All biodegradation experiments were run at 37 °C. Toxtree-v.2.6.13 software was used by decision tree approach to determine aquatic toxicity as one of the possible parameters. Bisphenol A was toxic to bacteria, as no bacterial growth (as increase of optical density) was measured in 120 h. After electrochemical oxidation optical density of the sample increased (0.1/0.3) indicating increased biodegradability, probably as the result of reduced toxicity. Indirect electrooxidation led to high 99% of bisphenol A removal, but COD decrease was only 25%. Modelling of acute eco-toxicity of determined products after indirect electrooxidation, namely dichloro-2-hydroxy-acetophenone, 4-hydroxy benzoic

2  Electrochemical Treatments for the Removal of Emerging Contaminants

153

acid, phenyl isomers, dihydroxy benzoic acid, and trihydroxy benzene resulted as “not-known” aquatic toxicity. The toxicity of bisphenol A and other tested emerging contaminants was evaluated in the study by Sun et  al. (2019b), as well as of their degradation products formed during the electrochemical treatment. The applied assay was the inhibition of growth of selected bacterium Escherichia coli. Bacterial growth was followed by measuring the absorbance at 600  nm. For bisphenol A, interesting results were obtained. Although bisphenol A was not completely mineralized during the treatment in the system for electrocatalytic wet air oxidation by activated oxygen on the graphite felt-supported α-MnO2 catalyst, the residual transformation products actually promoted E. coli growth compared to the control, opposite to bisphenol A which inhibited bacterial growth. Li et al. (2017) monitored acute toxicity by Vibrio fischeri test during electrooxidation of bisphenol A at BDD anode in NaCl electrolyte. The test showed a slight increase in the bioluminiscence inhibition after 5 min of treatment from the initial 70.96% to 74.47%. This was due to the formation of more toxic chlorinated intermediates of bisphenol A, such as dichlorophenols and trichlorophenols. However, the inhibition of bioluminiscence dropped to 1.70% after 4 h of treatment, although transformation products were still present in the solution. However, the authors remarked that the applied toxicity test is the most indicative of estrogenic activity. Alikarami et al. (2019) used a bioassay to monitor the toxicity outcome of combined treatment of bisphenol A in Na2SO4 medium with electrochemical-­ photocatalytic-­ ultrasound process using hydroxyl or sulfate radical-generating chemicals. The specific oxygen uptake rate of the active biomass was followed by oxygen electrode. Initial inhibition of O2 uptake in bisphenol A-containing solution was 31.3%. Authors compared the inhibition of O2 uptake after 2 h of treatment with photolysis, electrochemical process, or hybrid process and found that it changed to 48.1%, 31.9%, and 18.8%, respectively. Therefore, it was concluded that the hybrid process converted bisphenol A to less toxic transformation products in 2 h after a peak in toxicity after approx. 30  min of treatment. In contrast, 2  h-photolysis resulted in increased toxicity, while electrochemical treatment alone did not change this parameter. The latter was in accordance with the determination of bisphenol A concentration which changed only slightly during the electrochemical treatment alone.

2.5  Phthalic Acid Esters 2.5.1  Properties, Toxicity and Environmental Occurrence Phthalic esters (Fig. 2.8) are a class of organic compounds mainly used as plasticizers for the manufacture of polyvinyl chloride plastics, and also for cellulose, polyvinyl acetate and polyurethane formulation. Additionally, phthalic esters are used in

154

B. N. Malinović et al.

Fig. 2.8  Structure of phthalic ester. R1 and R2 groups can be linear, branched, linear/branched or cyclic

personal care products, coatings (e.g., pharmaceuticals), dyes and insecticides (Radke et al. 2019). Phthalic esters differ in alkyl substituents bonded to an ester group (R1 and R2 on Fig. 2.8) and range in molecular mass from 194.18 Da for the shortest phthalic ester, dimethyl phthalate, to 530.82  Da for longest phthalic ester, ditridecyl phthalate. Di-(2-ethylhexyl) phthalate and diisononyl phthalate represent the majority of global production of phthalic esters, i.e. ≈50% and ≈25%, respectively, which added up to about 5.4 million tons in 2014 (Benjamin et al. 2017). The main source of phthalic esters in the environment are plastics. As phthalic esters are not chemically bonded to the polymer they can be relatively easily released into the surroundings even under slight changes of the environment, e.g. sunlight irradiation, a temperature or pH change, contact with solvents, etc. phthalic ester content blended in plastic can reach up to 70%, which makes phthalic esters omnipresent in the environment: e.g. in soil (Lü et al. 2018), water (Paluselli et al. 2018), air (Bergh et al. 2011), food (Fierens et al. 2012), and drinks (Rudel Ruthann et al. 2011). This wide occurrence of phthalic esters in the environment can lead to potential human exposure via ingestion, inhalation, and dermal absorption. Epidemiological and toxicological studies have demonstrated that some phthalic esters can adversely interact with the endocrine system. According to biomonitoring and safety levels set by regulatory and scientific bodies (such as the USA Environmental Protection Agency), it should be safe if a healthy man with 75 kg body weight is exposed to up to 6 mg of di-(2-ethylhexyl) phthalate/day or 263 mg dimethyl phthalate/day (Benjamin et  al. 2017). However, with the widespread occurrence of phthalic esters it is estimated that human exposure to them can go beyond 1.0 g/day. Some phthalic esters have been classified as priority pollutants by environmental protection agencies (e.g. European Union, US, Canada, and China). In 2005, the European Commission banned di-(2-ethylhexyl) phthalate, dibutyl phthalate, and butylbenzyl phthalate in all toys and childcare articles (The European Commission 2005). Additionally, diisononyl phthalate, diisodecyl phthalate, and dioctyl phthalate were banned from use in toys and childcare articles, if those can be put in the mouth by the children. The International Agency for Research on Cancer has classified di-(2-ethylhexyl) phthalate as possibly carcinogenic to humans and the World Health Organization has set a guideline value for maximum allowed

2  Electrochemical Treatments for the Removal of Emerging Contaminants

155

exposure limit in drinking water to 8 μg/L (WHO 2013) while the limit set by USA Environmental Protection Agency is 6 μg/L (US EPA 2012). Fate and environmental biodegradation of four phthalic esters with different alkyl chain was investigated by Sun et al. (2019a). Dimethylphthalate, diethylphthalate, diallyl phthalate, and dipropyl phthalate were selected to assess toxicity, uptake, and biodegradation in seawater in the presence of algea Karenia brevis detected in several harmful algae blooms. 96  h-effective concentration (EC50) of dimethyl phthalate, diethyl phthalate, diallyl phthalate, and dipropyl phthalate were 0.257, 0.178, 0.136, and 0.095 mmol/L, respectively. Bioconcentration factors were correlated to the length of the alkyl chain related to higher lipophylicity. Concentrations of dimethyl phthalate, diethyl phthalate, diallyl phthalate, and dipropyl phthalate in the water decreased by 93, 68, 57, and 47%, respectively, due to their biodegradation in the presence of K. brevis, resulting in lower biomass production of algae (87, 61, 46, and 40%). Some metabolites were detected in sea water after testing with diethyl phthalate, while some of them were also detected within the cells. Santangeli et  al. (2017) investigated the toxicity of diisononyl phthalate on zebrafish Danio rerio reproduction. Female fish were exposed to five different concentrations of high molecular weight phthalate esters used as plasticizers (0.42, 4.2. 42, 420, and 4200 μg/L). Fish fecundity, oocyte growth, autophagic, apoptotic processes, and changes in morphological and biochemical composition of oocytes were monitored. After 8–21 days of treatment, the mean fertilized egg number per female per day was determined expressing significant decrease in fecundity. The most evident difference was at the lowest (0.42 μg/L) and at the highest (4200 μg/L) concentration of diisononyl phthalate. Gonad-somatic index also decreased at both mentioned concentrations. Histological analyses of oocysts showed a decreased number of vitelogenic oocysts at 0.42, 4.2, and 42 μg/L. Results suggested the ability of diisononyl phthalate to inhibit the maturation/ovulation processes, which was also confirmed by the decreased number of ovulated eggs observed. Diisononyl phthalate exposure can disrupt normal reproduction in zebrafish, altering gene expression patterns, the chemical composition of follicles and ovarian morphology. However, authors concluded that further investigations are needed.

2.5.2  E  lectrochemical Treatment Studies and Degradation Pathways Almost all electrochemical treatments have been conducted in undivided cells with solution volumes ranging from 0.1 L to 1 L. Both batch and flow reactors have been used. Usually three to four parameters were optimized such as current density or voltage between working and counter electrode, depending on the mode of operation, pH, solution temperature, type and concentration of supporting electrolyte, and duration of treatment. Most studies used a synthetic sample solution with a single model phthalic ester in concentration range of 20 μg/mL to 162 μg/mL and 0.05 to

156

B. N. Malinović et al.

0.15 M of supporting electrolyte. It is difficult to directly compare degradation efficiencies by process because of different experimental set-ups, however, coupled processes seem to be more effective in general, i.e. photoelectrocatalysis, electro-­ Fenton and electro-peroxone process vs. direct electrochemical oxidation (Souza et al. 2014b, c, d; Aquino et al. 2015; Huang et al. 2015; Medina et al. 2015; Dong et al. 2016; Espinoza et al. 2016; Ren et al. 2018). Several studies used the simplest form of phthalic ester, i.e., dimethyl phthalate, as a model compound in the development of new electrochemical treatments. These include: electrocoagulation (Huang et  al. 2015), direct electrochemical oxidation with dimensionally stable anode (Souza et al. 2014d), fluoride-doped lead dioxide(β-­ PbO2F) film electrodeposited on a Ti substrate (Souza et al. 2014c), and conductive-­ diamond anodes (Souza et al. 2014a). Conductive-diamond anodes were also used for photo-assisted electrochemical removal or photoelectrocatalysis (Souza et  al. 2014b). Moreover, the electro-Fenton system was also utilized for removal of dimethyl phthalate in a recent study by Ren et al. (2018). Besides dimethyl phthalate, diethyl phthalate and di-(2-ethylhexyl) phthalate were also chosen as model compounds in electrochemical treatments of phthalic esters. Dong et al. (2016) and Medina et al. (2015) studied the removal of diethyl phthalate by direct electrooxidation in a batch reactor whereas Espinoza et al. (2016) used a one-compartment flow cell by direct electrooxidation for diethyl phthalate removal. The electro-peroxone process was studied by Hou et  al. (2016) for the removal of diethyl phthalate. Zolfaghari et  al. (2016) studied the treatment of di-(2-ethylhexyl) phthalate in a landfill leachate using a membrane bioreactor coupled with electrooxidation. Real samples were also used in two studies by the research group Yang et al. where they looked at the removal of two or more phthalic esters (Yang et al. 2014a; Yang et al. 2016). In both studies, a simultaneous electrocoagulation and electrofiltration process with different membranes was tested. Finally, Aquino et al. (2015) used phthalic acid, a common transformation product in electrochemical treatments of phthalic esters, as a model compound for treatment by photoelectrocatalysis. Works of all the aforementioned research are summarized in Table 2.3. Ren et  al. (2018) made the electro-Fenton system effective by using CeO2 nanoparticles as the catalysts for H2O2 generation (CeO2 electro-Fenton system). Experiments were done in a Plexiglas reactor with a three-electrode system: graphite electrodes as working electrode and counter electrode. Before electrocatalysis, oxygen was bubbled near the cathode with a flow rate of 0.3 L/min for 1 h. Authors compared the CeO2 electro-Fenton system to the traditional electro-Fenton system using four different electrolytes (all 0.1 M) with an initial pH value of 3.0: NaH2PO4, NaNO3, NaCl, and Na2SO4. Improvement of removal efficiency was observed when Na2SO4 was used, however, no significant change in removal efficiency was observed in NaNO3 and NaCl. Interestingly, nanoparticles CeO2 showed a negative effect in the NaH2PO4 electrolyte. By using Na2SO4, the removal efficiency of dimethyl phthalate improved from 86% to 95% using electro-Fenton and CeO2 electro-­Fenton, respectively, after 20 min of treatment. The mineralization rate (by TOC) was 94% using CeO2 electro-Fenton after 120 min. Transformation products were not investigated. Huang et  al. (2015) on the other hand investigated

100 μg/mL in 0.1 M NaCl

Sample characteristics Phthalic ester: DMP 29 μg/mL in 0.1 M Na2SO4 (NaH2PO4, NaNO3, NaCl)

UC

1

0.14

UC

Electro-Fenton-CeO2

Electrocoagulation

V [L]

Configuration

Process

Electrochemical reactor

Table 2.3  Review of the electrochemical treatments of phthalates

Mode: PM WE and CE: Graphite plate (82.4 cm2) E: – 0.7 V (vs. SCE) T: 25 °C pH: 3.0 Catalyst: CeO2 O2 (1 h): 0.3 L/ min Mode: GM Electrodes: Fe/Al (30 cm2) (also tested Fe/Fe, Al/ Al, Al/Fe) j: 20 mA/cm2 T: 15–45 °C (25 °C)

Electrodes/ operational parameters References

DMP: 92% DT: 1 h

(continued)

Huang et al. (2015)

Ren et al. Electro-Fenton-CeO2: (2018) 95% DT: 20 min TOC (DT: 120 min): 94% Electro-Fenton: 86% DT: 20 min TOC (DT: 120 min): 76%

Degradation efficiency

2  Electrochemical Treatments for the Removal of Emerging Contaminants 157

Configuration One compartment filter-press cell

One compartment filter-press cell

One-compartment flow cell

162 μg/mL in Na2SO4 (NaCl, Direct electrooxidation at 0.15 M ionic strength)

162 μg/mL in Na2SO4 (NaCl, Direct electrooxidation at 0.15 M ionic strength)

Electrochemical reactor

Sample characteristics Process 162 μg/mL in NaCl (Na2SO4, Direct electrooxidation at 0.15 M ionic strength)

Table 2.3 (continued)

0.4 (solution)

0.35 (solution)

V [L] 0.35 (solution)

Electrodes/ operational parameters Mode: GM Anode: DSA (Ti/ Ru0.3Ti0.7O2) (14 cm2) Cathode: Pt (14 cm2) j: 20 mA/cm2 T: 10–40 °C (30 °C) pH: 2.0–10.0 (2.0) Mode: GM Anode:Ti/β-PbO2F (14 cm2) Cathode: Pt (14 cm2) j: 10 mA/cm2 T: 10–50 °C (20 °C) pH: 3.0–10.0 (3.0) Mode: GM Anode and cathode: Si/BDD (78 cm2) j: 20 mA/cm2 T: 25 °C Souza et al. (2014c)

References Souza et al. (2014d)

Souza et al. DMP: ≈100% (2014a) TOC: ≈100% COD: ≈100% HPLC-UV: Some TPs are still present DT: 4.5 h

DMP: 57% TOC: 35% DT: 5 h

Degradation efficiency DMP: ≈40% TOC: ≈40% DT: 5 h

158 B. N. Malinović et al.

One-compartment flow cell

Photoelectro-catalysis

Sonoelectrolysis (direct electro-oxidation)

162 μg/mL in 0.07 M (also in 0.007 M and 0.035 M) NaCl (at constant ionic strength of 0.15 M – Addition of Na2SO4)

162 μg/mL (also 1.6 μg/mL and 16 μg/mL) in Na2SO4 (NaCl, at 0.15 M ionic strength) One-compartment flow cell

Configuration One-compartment flow cell

Sample characteristics Process 162 μg/mL in Na2SO4 (NaCl, Photoelectro-catalysis at 0.15 M ionic strength)

Electrochemical reactor

Electrodes/ operational V [L] parameters 0.8 Mode: GM (solution) Anode: Si/BDD (78 cm2) Cathode: SS (78 cm2) j: 20 mA/cm2 T: 25 °C LS: 15 W UVC (254 nm) 0.35 Mode: GM (soluti-on) Anode: DSA (Ti/ Ru0.3Ti0.7O2) (14 cm2) Cathode: Ti-mesh j: 20 mA/cm2 T: 25–45 °C (45 °C) pH: 2.0 LS: 250 W UVC (254 nm) 0.8 Mode: GM (solution) Anode/cathode: Si/BDD (78 cm2) j: 20 or 60 mA/cm2 T: 25 °C pH: 2.0 Ultrasound generator: 24 kHz (max. 200 W) Souza et al. (2014e)

(continued)

Souza et al. Sonoelectrolysis: DEP/ (2013) COD/TOC: ≈100%, (direct E-oxid.: DEP and TOC: ≈100% COD: ≈95%) DT: NA HPLC-UV: Some TPs are still present

DMP: ≈100% TOC: ≈90% DT: 5 h

Degradation efficiency References Souza et al. DMP: ≈100% (2014b) TOC: ≈100% COD: ≈100% HPLC-UV: Some TPs are still present DT: 4.5 h

2  Electrochemical Treatments for the Removal of Emerging Contaminants 159

Direct electrooxidation

Direct electrooxidation

50 μg/mL in 0.1 M Na2SO4 + WWTP water sample (spiked)

100 μg/mL in 0.05 M Na2SO4

Sample characteristics Process Phthalic ester: DEP 20 μg/mL in 0.05 M Na2SO4 Electro-peroxone (ozonation, direct electro-oxidation)

Table 2.3 (continued)

UC

0.25

0.1 (solution)

0.4 (solution)

UC (acrylic column reactor)

DC

V [L]

Configuration

Electrochemical reactor

Mode: GM Anode: Pt (12 cm2) Cathode: Carbon felt (RVC, carbon-PTFE) I: 400 mA O3/O2: 0.4 L/min Mode: GM Anode: BDD (1 cm2) (PbO2, Pt) Cathode: Ti plate (1 cm2) j: 200 mA/cm2 T: 25 °C Mode: GM Anode: Si/BDD gold modified (1 cm2) (Si/BDD) Cathode: 316 SS j: 30 mA/cm2

Electrodes/ operational parameters References

Dong et al. (2016)

Medina et al. (2015)

DEP (DT: 2 h): ≈100% TOC (DT: 3.5 h): ≈100%

DEP: ≈40% DT: 250 min

Hou et al. Electro-peroxone: DEP (2016) (DT: 5 min): 92%, TOC (DT: 1 h): 92% E-oxid.: DEP: 14%, TOC: 5% (DT: 1 h); Ozonation: DEP: 99%, TOC: 14% (DT: 1 h)

Degradation efficiency

160 B. N. Malinović et al.

One-compartment flow cell

Phthalic esters: Mixture Model solution: Hundreds of TGCCM coupled with electrocoagulation-­ ng/L to several μg/L; electro-­filtration WWTP water sample NA

1

1

One-compartment flow cell (PVC reservoir, with 2 anodes and cathodes)

One-compartment flow cell

V [L]

Configuration

Microbial bioreactor with electrooxidation

Municipal landfill leachate (≈50 μg/mL)

Sample characteristics Process Phthalic ester: DEHP 100 μg/mL in 7 mM Na2SO4 Direct electrooxidation

Electrochemical reactor

Anode: Al Cathode: 316 SS EFS: 40 V/cm CFV: 1.63 cm/s TMP: 294 kPa Effective filtration area: 62.8 cm2

Mode: GM Anode: DSA Nb/ BDD (45 cm2) (Ti/ IrO2, Ti/ IrO2-RuO2) Cathode: Ti (65 cm2) I: 200 mA T: 20 °C pH: 7 Mode: GM Anode: Nb/BDD (65 cm2) Cathode: Ti (65 cm2) j: 46 mA/cm2 T: 20 °C

Electrodes/ operational parameters

Model solution: DBP: 99%, DEHP: 99% WWTP water sample: DBP: 47–58%, DEHP: 80–83% DT: NA

(continued)

Yang et al. (2016)

Zolfaghari et al. (2016)

Espinoza et al. (2016)

DEHP: 81% DT: 2 h

DEHP: 80% TOC: 48% COD: 50% DT: 2 h Hydraulic retention time: 18 h

References

Degradation efficiency

2  Electrochemical Treatments for the Removal of Emerging Contaminants 161

Process TCCACM coupled with electrocoagulation-­ electro-­filtration

One compartment filter-press flow cell

Configuration One-compartment flow cell

Electrochemical reactor

0.5 (solution)

V [L] NA

Degradation efficiency DBP: e-coag./e-filtr.: 90%, (e-filtr.: 78%, CFF: 19%) DEHP: e-coag./e-filtr.: 89%, (e-filtr.: 73%, CFF: 20%) DNP: e-coag./e-filtr.: 81%, (e-filtr.: 61%, CFF: 27%) DT: 1 h

PhA: ≈100% Mode: GM TOC: 60% Anode: DSA; DT: 3 h Ru0.3Ti0.7O2 (15 cm2) Cathode: Ti-mesh (15 cm2) j: 20 mA/cm2 T: 30 °C pH: 3.0–11.0 (3.0) LS: 250 W UVC (254 nm)

Electrodes/ operational parameters Anode: Al (316 SS) Cathode: 316 SS EFS: 20 V/cm CFV: 6.91 cm/s TMP: 294 kPa Effective filtration area: 47.1 cm2

Aquino et al. (2015)

References Yang et al. (2014a)

Abbreviations: BBP butylbenzyl phthalate, carbon-PTFE carbon-polytetrafluorethylene, CE counter electrode, CFF conventional crossflow filtration, CFV crossflow velocity, DBP dibutyl phthalate, DC divided cell, DEHP di-(2-ethylhexyl) phthalate, DEP diethyl phthalate, DIBP diisobutyl phthalate, DINP diisononyl phthalate, DMP dimethyl phthalate, DNOP dinonyloctyl phthalate, DSA dimensionally stable anode, DT duration/time of experiment, EFS electric field strength, GM galvanostatic mode, LS light source, NA not available, PhA phthalic acid, PM potentiostatic mode, PVC polyvinyl chloride, RVC reticulated vitreous carbon, SCE saturated calomel electrode, SS stainless steel, TCCACM tubular carbon nanofiber/carbon/alumina composite membrane, TGCCM graphene-­containing ceramic composite tubular membrane, TMP transmembrane pressure, TP transformation product, UC undivided cell, WE working electrode, WWTP wastewater treatment plant

Phthalic ester: Phthalic acid 150 μg/mL in 0.05 M NaCl Photoelectro-catalysis (Na2SO4, Na2CO3, NaNO3)

Sample characteristics Tap water: 1–447 ng/L Drinking-fountain water: 1–316 ng/L (BBP, DBP, DIBP, DEHP, DEP, DMP, DINP, DNOP)

Table 2.3 (continued)

162 B. N. Malinović et al.

2  Electrochemical Treatments for the Removal of Emerging Contaminants

163

electrocoagulation as a possible treatment of dimethyl phthalate by electrochemically generated iron hydroxides. Tests with four different electrode pairs (Fe/Al, Fe/ Fe, Al/Al, and Al/Fe) showed best results using a Fe/Al electrode pair at 20 mA/ cm2. Removal efficiency of dimethyl phthalate at optimized conditions was 92% after 60 min of treatment. The Langmuir adsorption isotherm model best fitted the experimental data for the adsorption. TOC was not measured, as there were no degradation studies done. Souza’s group published a series of articles on the removal of dimethyl phthalate by direct and indirect electrochemical oxidation (Souza et al. 2014a, b, c, d). The focus of each article was on the removal efficiency of the specific anode and optimization of some key parameters, such as support electrolyte, pH, and temperature. Since electrolysis was done under galvanostatic conditions, different current densities were tested. Electrochemical degradation was monitored with HPLC-UV, COD, and TOC. In the first study, dimensionally stable anode (Ti/Ru0.3Ti0.7O2) was used as an anode and Ti as a cathode (Souza et al. 2014d). Testing two different electrolytes, Na2SO4 and NaCl (at 0.15  M ionic strength), revealed higher removal efficiency using NaCl probably due to the formation of Cl2 and HClO. The concentrations of recalcitrant organochlorine compounds that can form in such systems were minimized with low NaCl concentration (7 mM) and low current density at 20 mA/cm2. An acidic solution (pH 2) at high temperature (30 °C) yielded best dimethyl phthalate and TOC removal rates. However, these were low even after 5 h: approx. 40% removal of dimethyl phthalate and TOC, respectively. Better dimethyl phthalate removal efficiency was obtained by using fluoride-doped lead dioxide film electrodeposited on a Ti substrate (Ti/β-PbO2F) as an anode and Ti as a cathode (Souza et al. 2014c). Surprisingly, dimethyl phthalate removal was higher in the presence of Na2SO4 than in the presence of NaCl (at 0.15 M ionic strength); however, no difference was observed in the TOC removal. This could indicate that the dominant removal pathway was through hydroxyl radicals and that chlorine ions led to a parasitic reaction with formation of less oxidative chloro-oxidant species. For an effect of current density, higher dimethyl phthalate removal was observed with increasing levels of applied current density. The rate of change in the removal was smaller for TOC, indicating that with higher levels of current density more recalcitrant transformation products remain in the solution. Thus, the low current density of 10 mA/cm2 was chosen as optimal. On the other hand, the temperature (10–50 °C) and pH (3, 7, and 10) of the solution did not play a significant role in removal efficiency. The results of electrolysis at optimized conditions after 5 h showed that 57% of initial dimethyl phthalate was removed. Despite applying low current density, the TOC removal rate was only 35% under the same conditions, indicating significant formation of recalcitrant transformation products. Some of these were detected by HPLC-UV analysis and a transformation pathway was proposed. First, the formation of monomethyl phthalate from dimethyl phthalate, and then phthalic acid, which further oxidized into 2-hydroxybenzoic acid, benzene-1,3-diol (and benzene-­1,4-diol), and 1,4-benzoquinone. Additionally, maleic acid and oxalic acid were detected which can form from ring cleavage. However, the research group obtained best results using electrolysis with a conductive-diamond electrode

164

B. N. Malinović et al.

(p-Si-­boron-doped diamond, Si/BDD) as anode and cathode (Souza et al. 2014a). Comparing degradation results done with Na2SO4 or NaCl as a supporting electrolyte (at 0.15 M ionic strength), authors found that the concentration of transformation products was higher when using NaCl. Besides the higher concentration of monomethyl phthalate and phthalic acid, additional chlorinated phenolic derivatives were observed in NaCl which were not detected in Na2SO4: 3-chlorophenol, 2,4-dichlorophenol, 2,4,6-trichlorophenol, and 4-chlororesorcinol. Similar to other studies, authors found that removal efficiency was highest for low current density, thus the current density of 20 mA/cm2 was chosen. Dimethyl phthalate was completely removed after approx. 4.5 h in Na2SO4. Although TOC determination seems to indicate that no organic compounds remained in the solution after 4.5  h, HPLC-UV analysis revealed that some oxalic acid was still present in the solution (approx. 40  μM). Based on analysis of transformation products, the degradation pathway shown on Fig. 2.9 was proposed Besides developing methods that use anodic oxidation for removal of dimethyl phthalate, Souza et  al. (2014b) also developed a photo-assisted electrochemical approach, i.e. photoelectrocatalysis. They used a Si/BDD electrode that was previously tested separately (Souza et al. 2014a) and a 15 W UVC lamp (max. 254 nm) for UV irradiation. The electrochemical flow cell contained 0.8 L of solution. Other experimental conditions were similar to the previous studies. Regardless of current density (20, 60, or 120 mA/cm2), electrolysis and UV irradiation had a synergistic effect in removal efficiency with higher removal of TOC. This holds true for both electrolytes used: Na2SO4 and NaCl, although removal efficiency was worse in general for all three tested levels of current density when NaCl was used. Moreover, analysis of transformation products revealed that by using NaCl, chlorophenols remained in the solution after the treatment. Comparing concentrations of transformation products when using Na2SO4 showed similar concentrations for all tested current densities. An exception was found at the current density of 60  mA/cm2 where the concentration of phthalic acid and oxalic acid was 3–6 times greater after

Fig. 2.9  Currently known electrochemical degradation pathways of phthalic acid esters, namely dimethyl phthalate and diethyl phthalate, reported in literature. (Souza et  al. 2014a, b, c; Hou et al. 2016)

2  Electrochemical Treatments for the Removal of Emerging Contaminants

165

photo-electrolysis than after electrolysis. On the other hand, the energy consumption of individual parts (electrolysis and photolysis) is similar when operated at low current density, meaning that coupled processes are economically viable when one would like to avoid higher current densities. Additionally, the Souza research group coupled direct electrooxidation with ultrasound irradiation using the same Si/BDD electrodes (Souza et  al. 2013). Ultrasound irradiation improved COD and TOC removal rates compared to direct electrooxidation in both electrolytes (Na2SO4 and NaCl). However, significant synergistic effects are only seen when electrochemical processes are limited by mass transport and ultrasound irradiation promotes additional oxidation mechanism, i.e. at very large current densities. This means that such applications have relatively high operation costs. Characterization of transformation products showed similar results to transformation products found by direct electrooxidation (Souza et al. 2014a) and photoelectrocatalysis (Souza et al. 2014b): i.e. monomethyl phthalate, phthalic acid, 4-hydroxyphthalic acid, and small carboxylic acids (especially oxalic acid) were the main intermediates in both electrolytes, with chlorophenols also present in the NaCl electrolyte (see also Fig. 2.9). Apart from the boron-doped diamond electrode, dimensionally stable anode could also be a suitable electrode for electrocatalysis since lower generation of highly reactive oxidants can be overcome with irradiation with UV light as was demonstrated by Souza et al. (2014d). In this study, electrochemical degradation was studied at three different concentration of NaCl (0.007, 0.035, and 0.07 M), but at a constant ionic strength of 0.15 M (through addition of Na2SO4). Interestingly, dimethyl phthalate and TOC removal rates increased with higher concentrations of NaCl. This could be due to chloro-oxidant species (Cl2 and HOCl), which are more easily generated on the dimensionally stable anode in combination with UV irradiation. Additionally, chloro-oxidant species can (together with UV irradiation) further propagate the production of hydroxyl radicals. In order to avoid the formation of undesirable chlorooxidant species at higher oxidation states, the pH was set to 2. With the concentration of NaCl and current density set to 70 mM and 20 mA/cm2, respectively, complete degradation of dimethyl phthalate and approximately 90% TOC removal was achieved in 300 min. Unfortunately, transformation products were not elucidated in the study. As already mentioned, the formation of organochlorine compounds is especially important because they can be even more toxic than the parent compound. Conventional ozonation process is not an effective way of removing phthalic esters because they in general react very slowly with ozone. As a result, Hou et al. (2016) developed a method using the electro-peroxone process. Here, H2O2 is generated in situ from cathodic reduction of O2 and mixture of O2 and O3 is sparged into the reactor. A non-selective oxidant hydroxyl radical is then generated from the reaction of H2O2 with O3. Experiments were conducted in an undivided acrylic column reactor (400 mL) with 0.05 M Na2SO4 as a supporting electrolyte and ozone generator. A direct current power supply was used for electrolysis under galvanostatic conditions. The authors used diethyl phthalate as a model compound of phthalic esters and tested three different carbon-based cathodes: carbon-­ polytetrafluoroethylene, carbon-felt, and reticulated vitreous carbon. Research demonstrated that electrolysis alone is ineffective in the oxidation of short alkyl

166

B. N. Malinović et al.

chain phthalates such as diethyl phthalate with a Pt anode (less than 5% TOC removal and only 14% removal of diethyl phthalate after 60 min). Similarly, ozonation alone removed only 14% of the initial solution TOC, however removed almost all diethyl phthalate (99%) in 60 min. On the other hand, TOC removal yield using the electro-peroxone process was between 76% and 92% after 60 min and diethyl phthalate completely degraded after only 5 min. Electrocatalytic activity for converting O2 to H2O2 increased in order: carbon-felt, reticulated vitreous carbon, and carbon-polytetrafluoroethylene. By analyzing intermediates using HPLC-UV, hydroxyl radical was found to be the primary oxidant. Two aromatic intermediates (monoethyl phthalate and catechol) and three smaller carboxylic acids (oxalic, succinic, and pyruvic acid) were detected. The deethylation of diethyl phthalate and formation of monoethyl phthalate and phthalic acid as well can occur by hydrogen atom abstraction with hydroxyl radical from ethyl group. One the other hand, hydroxyl radical addition to the aromatic ring can yield diethyl 4-hydroxyphthalate. Phenols such as catechol, benzene-1,2,4-triol, and pyrocatechol can be products of hydroxyl radical-induced decarboxylation and hydroxylation. Lastly, smaller carboxylic acids form from ring cleavage through hydroxyl radical or ozone oxidation. Electrochemical degradation of diethyl phthalate was also studied by Dong et al. (2016) and Medina et al. (2015). Both research groups used direct electrooxidation with a BDD electrode under galvanostatic conditions. Medina et al. (2015) focused on degradation efficiency using BDD modified with gold particles with and without chloride ions. The degradation efficiency was improved by approx. 10% when using a modified BDD electrode. Interestingly, the presence of chloride ions in the solution did not affect degradation efficiency. However, the stability of the modified BDD electrode was an issue. Dong et al. (2016) performed experiments in 0.1 M Na2SO4 and also used municipal wastewater. Degradation of diethyl phthalate was found to be a pseudo-first-order reaction. Comparing three different electrodes, they found that removal efficiency increased in the following order: Pt   photoperoxi-coagulation process > photoelectro-Fenton-ZnFe2O4 nanoparticles > photoelectrocatalysis > photocatalysis > electro-Fenton-ZnFe2O4 nanoparticles ≈ direct photolysis > direct electrooxidation, depending on the chosen electrode. Ding et al. (2010) prepared a modified glassy carbon electrode by liquid phase deposition of TiO2 film on the glassy carbon electrode surface and used it as a working electrode in photoelectrocatalytic removal of 1H-benzotriazole. They compared the treatment to simpler processes: direct electrooxidation, direct photolysis, and photocatalysis. For photo-assisted experiments, the TiO2 film was irradiated with UV light (254 nm) to produce electron excitation and positively charge holes, which oxidized 1H-benzotriazole. Almost no 1H-benzotriazole degradation was observed during the direct electrooxidation process, whereas best results were observed with a combination of photocatalysis and electrochemical oxidation, which yielded synergetic effects with a degradation efficiency of 90% after 180 min. The mineralization rate was estimated with COD using the traditional dichromate method. Results

119 μg/mL in 0.2 M Na2SO4

18 μg/mL in 0.05 M Na2SO4

Heterogeneous photoelectro-FentonZFNPs (also adsorption by ZFNPs, direct electrooxidation, direct photolysis, photoelectrocatalysis, electro-Fenton-ZFNPs) Direct electrooxidation

Sample characteristics Process 24 μg/mL in 0.5 M Photoelectrocatalysis (also Na2SO4 direct electrooxidation, direct photolysis, photocatalysis)

UC

UC

0.4

NA

Electrochemical reactor V Configuration [L] UC 0.1

Table 2.4  Review of the electrochemical treatments of 1H-benzotriazole

Electrodes/operational parameters Mode: PM WE: TiO2 film on GC (1 cm2) CE: Pt E: +0.8 V (vs. SCE) T: 25 °C pH: 2.0–10.0 (optimal 6.2) LS: 15 W UVC (254 nm) Mode: PM WE: Graphite block (25 cm2) CE: Carbon rod (3.8 cm2) E: −0.6 V (vs. SCE) T: 25 °C pH: 3.0–10.0 (optimal 3.0) LS: 15 W UVC (254 nm) Catalyst: 0.067 g/L ZFNPs Mode: GM Anode: BDD (4 cm2) Cathode: SS plate (4 cm2) j: 20 mA/cm2 T: 20 °C adsorption by ZFNPs: 5% Dir. e-oxid.: < 5% Dir. photolys.: 34% Photoelectrocat.: 67% Electro-Fenton-ZFNPs: 32% Photoelectro-Fenton-­ZFNPs: 91% (COD: 85%) DT (for all experiments): 3 h Dir. e-oxid.: 72% COD: ≈48% DT: 5 h

Xing et al. (2012)

Wu et al. (2013)

Degradation efficiency References Ding et al. Dir. e-oxid.: < 5% (2010) Dir. photolys.: 65% Photocatal.: 70% Photoelectrocatal.: 90% (COD: 74%) DT (for all experiments): 3 h

172 B. N. Malinović et al.

Photoelectro-peroxone process /ZVI

50 μg/mL in 0.01 M Na2SO4 (NaCl, NaNO3, NaNO2) UC

UC

0.45 Mode: GM Anode: Fe (38 cm2) Cathode: ACF (99 cm2) I: 150 mA T: 26–28 °C pH: 3.0 LS: 12 W UVC 0.45 Mode: GM Anode: Pt (4 cm2) Cathode: Graphite felt (141 cm2) I: 300 mA T: 25–28 °C pH: 3.0 LS: 8 W UVC (254 nm) O3/O2: 0.25 L/min Catalyst: 100 mg/L ZVI

Ahmadi et al. (2016)

Ahmadi and Ghanbari (2018)

BTA: 96% COD: 76% TOC: 65% DT: 1 h

BTA: ≈100% TOC: 85% DT: 0.5 h

Abbreviations: ACF activated carbon fiber, BTA 1H-benzotriazole, CE counter electrode, DT duration/time of experiment, GC glassy carbon, GM galvanostatic mode, LS light source, NA not available, PM potentiostatic mode, SCE saturated calomel electrode, UC undivided cell, WE working electrode, ZFNPs ZnFe2O4 nanoparticles, ZVI zero valent ion (Fe0)

Photoperoxi-coagulation

40 μg/mL in 0.01 M Na2SO4

2  Electrochemical Treatments for the Removal of Emerging Contaminants 173

174

B. N. Malinović et al.

indicate that some transformation products remain in solution even after 180 min of treatment since COD removal efficiency is lower than the corresponding degradation efficiency measured by HPLC-UV. Wu et al. (2013) investigated a heterogeneous photoelectro-Fenton like process using ZnFe2O4 nanoparticles as a catalyst. Graphite electrodes were used as the working electrode and counter electrode and a 15 W UVC lamp (max. 254 nm) was used for UV irradiation. A comparison of the efficiency of 1H-benzotriazole degradation was made between adsorption by ZnFe2O4 nanoparticles, direct electrooxidation, direct photolysis, photoelectrocatalysis, heterogeneous electro-Fenton like process with ZnFe2O4 nanoparticles and heterogeneous photoelectro-Fenton like process with ZnFe2O4 nanoparticles. Very low removal efficiency was reported with direct electrooxidation and only 5% removal efficiency by adsorption by ZnFe2O4 nanoparticles. The removal efficiency was increased to 34% with direct photolysis and to 67% with photoelectrocatalysis. In the latter process, more hydroxyl radicals were produced since hydrogen peroxide, which was produced on the surface of the cathode, was decomposed by UV irradiation. For the heterogeneous electro-Fenton-­ like process, removal efficiency was 32%, which is significantly higher than without ZnFe2O4 nanoparticles. Adding a light source (photoelectro-Fenton like process) raised removal efficiency to 78%. With adjusting initial pH, dosage of ZnFe2O4 nanoparticles and applied potential for removal of 18 μg/mL 1H-benzotriazole to 3.0, 0.067 g/L, and − 0.6 V, respectively, final removal efficiency was 91%. Again, results for COD showed that intermediates are formed during the degradation of 1H-benzotriazole. Ahmadi et  al. (2016) tested photoperoxi-coagulation as a means of 1H-benzotriazole removal. Activated carbon fiber was used as a cathode and iron sheet as an anode. The supporting electrolyte was 0.01 M Na2SO4 with pH adjusted to 3.0. The applied current, time, initial concentration of 1H-benzotriazole, and UV light power were optimized using a Box-Behnken design. The optimal conditions for the removal of 40 μg/mL of 1H-benzotriazole were determined at 60 min electrolysis time, applied current of 150 mA and UV light power of 12 W, yielding 96% removal. A study of the mechanisms of photoperoxi-coagulation revealed that removal by adsorption represents about 12% and that the main mechanisms are coagulation and oxidation with the contribution of oxidation mechanism being greater in all tested conditions. However, the increase in electrolysis time enhanced the coagulation mechanism. Ahmadi and Ghanbari (2018) developed a degradation method by photoelectro-­ peroxone/zero valent ion (Fe0) process for removal of Metanil Yellow and tested the method on several emerging contaminants, including 1H-benzotriazole. Here, hydrogen peroxide was electrogenerated and a mixture of ozone and oxygen were sparged into the solution. Additionally, two UVC lamps were used to enhance the photo-activation of ozone and hydroxyl radical formation. Zero valent iron was used as a source of ferrous ions (oxidized by oxygen sparged into the solution and electrogenerated hydrogen peroxide) which in turn reacted with hydrogen peroxide

2  Electrochemical Treatments for the Removal of Emerging Contaminants

175

to form hydroxyl radical. By comparing this complex system with individual, binary, and ternary subsystems, researchers were able to show high synergistic effects for Metanil Yellow removal. Interestingly, the effect of supporting electrolyte was as follows: nitrite ions showed a high inhibitory effect while chloride ions had no significant effect, and highest efficiency was observed with sulfate ions. By applying optimal conditions found for Metanil Yellow degradation (pH  3.0, 100  mg/L Fe0, 33.2  mg/L ozone, 300  mA applied current) complete removal of 1H-benzotriazole was achieved in 30  min. Total organic carbon (TOC) for 1H-benzotriazole removal was 85%, indicating that some transformation products remained in solution. Degradation pathways after electrochemical treatment have not as yet been extensively studied. Currently known electrochemical degradation pathways of 1H-benzotriazole are presented in Fig. 2.11. Transformation products were in general identified by comparing UV-vis spectra and HPLC-UV chromatogram changes of 1H-benzotriazole solution at different reaction times (Ding et al. 2010). The degradation mechanism with electrochemical treatment of 1H-benzotriazole can be initiated with triazole ring opening and 2-aminobenzenediazonium formation and can be detected by comparing UV spectra (absorption peak at 400 nm). Loss of amino group yields benzenediazonium (m/z 105 detected by liquid chromatography coupled with mass spectrometry) and further aniline (confirmed by HPLC-UV analysis). A study by Xing et al. (2012) found that hydroxyl radicals were the dominant oxidants at BDD electrode using direct electrooxidation. By comparing theoretical calculations of delocalization energy and atom charge with results of GC-MS analysis they identified positions of hydroxylation together with three nitro-products: 2-nitroaniline, 1,2-dinitrobenzene, and 2-nitrophenol. In our own research (Prosen et al. 2020), we identified additional transformation products using LC-MS/MS. We tested three different supporting electrolytes (NaCl, Na2CO3, and H2SO4). Monochloro-benzotriazole and methyl-benzotriazole were positively confirmed and dichloro-benzotriazole was tentatively confirmed when using NaCl. Hydroxylated products were detected in H2SO4: aminophenol and monohydroxy-­benzotriazole positively; di- and tri-hydroxy-benzotriazole tentatively. These hydroxylated products were also detected in a study by Ma et al. (2019) where sulfate and hydroxyl radicals were generated by heat activation of persulfate.

2.7  Consequences of Electrochemical Treatments Electrochemical treatments represent a very complex challenge on multiple levels, i.e. from chemical to engineering point of view. Different aspects on consequences of electrochemical treatments are discussed in the following subsections. Table 2.5 summarizes these findings together with the suggestions for their improvement.

176

B. N. Malinović et al.

Fig. 2.11  Currently known electrochemical degradation pathways of 1H-benzotriazole from Ding et al. (2010), Xing et al. (2012), Wu et al. (2013) and from Prosen et al. (2020) (means of detection are specified in parenthesis for each product)

2.7.1  Transformation Products and Their Identification Complete mineralization of targeted organic contaminants in wastewater is usually not achieved due to substantial energy requirements needed in the actual water treatment. These can arise in part because of other coexisting chemical species (e.g. natural organic matter, anions, particulate matter) that may become potential competitors in consuming hydroxyl radicals. Therefore, degradation pathways need to be explored and toxicological assessment performed not only for target organic contaminants, but also for transformation products. Traditionally, degradation studies have usually been monitored by using HPLC coupled to UV, diode array detector, or low resolution mass spectrometry. Nowadays, high resolution mass spectrometry is predominantly used since it enables far better elucidation of results. The technique of choice for the identification of degradation products of bisphenol A in water solutions is LC-MS. Several papers describe their identification with

2  Electrochemical Treatments for the Removal of Emerging Contaminants

177

Table 2.5  Overview of some general aspects of electrochemical treatments and possibilities for their improvement Aspect of electrochemical treatment Improvement suggestion Effectiveness of electrochemical treatments Incomplete mineralization (coexisting Prolongation of the treatment time (high energy chemical species in samples) requirements) Combination of electrochemical methods with biological and other treatments methods (synergistic effect) Monitoring of pollutants and transformation products for degradation pathways elucidation Monitoring of mineralization based on total Non-target analysis of species using liquid or organic carbon and chemical oxygen demand gas chromatography coupled to high resolution mass spectrometer Monitoring of target species with liquid chromatography coupled to UV, diode array detector or low resolution mass spectrometer Toxicity of pollutants Transformation products can increase overall Battery of biotests with organisms from different trophic levels for the evaluation of treatment toxicity efficiency. Toxicity assessment based on various Ecotoxicological assessment bioassays Comprehensive biodegradability study based on non-target chemical analysis of transformation and degradation products Electrochemical treatment conditions Model system with different electrolytes Studies should be conducted on real systems (NO3−, SO42−, Cl−, ClO4−, CO32−) (wastewater or wastewater treatment plant effluents) which represent a very complex matrix Laboratory scale Large scale

this technique (Bakr and Rahaman 2017; Chen et al. 2017; Divyapriya et al. 2017; Zhang et al. 2018; Aravind et al. 2019; Bakr and Rahaman 2019; Pan et al. 2019). Also GC-MS was utilized, mainly for identification of degradation products with lower molecular masses (Li et  al. 2017; Burgos-Castillo et  al. 2018; Aravind et al. 2019). A large number of degradation products were identified in the reviewed papers (Fig. 2.7). Some authors reported the degradation of the primary pollutant bisphenol A and also further degradation of the degradation products (intermediates) (Aravind et al. 2019). Predominantly, these are compounds with the same skeleton as bisphenol A, but with different or additional functional groups (e.g., hydroxylated products). Furthermore, compounds with lower molecular masses including one benzene ring are also numerous, predominantly formed by the cleavage of the skeleton between the benzene rings. As the end products of degradation, relatively small molecules appear, which are also formed as degradation products of other organic compounds (e.g., maleic and oxalic acid). The final stage is the mineralization into inorganic CO2 and H2O.

178

B. N. Malinović et al.

In the case of application of electrolytes containing the chloride ion, chlorinated degradation products were observed (Li et al. 2017; Burgos-Castillo et al. 2018); mainly from the reaction with HClO or other chlorinated species formed from Cl− oxidation at anode (Li et al. 2017; Burgos-Castillo et al. 2018). In the electrolyte containing Cl−, the number of degradation products is higher. The toxicity of chlorinated organic compounds could be problematic, as chlorinated organic compounds have relatively high toxicity (Li et al. 2017). The degradation mechanism of bisphenol A in combination with other pollutants such as hexavalent chromium (Cr(VI)) was also described (Zhang et al. 2018). Most studies conducted for phthalic esters did not investigate transformation products but rather only monitored the concentration of selected phthalic esters and determined the degree of mineralization through TOC measurements. Possible transformation products were analyzed for dimethyl phthalate and diethyl phthalate using UV detection after chromatographic separation on HPLC reverse phase (at 274 nm) or ion exchange (at 190 nm) analytical column (Souza et al. 2013; Souza et al. 2014a, c; Hou et al. 2016). Such analysis gives only limited information and thus prevents detailed degradation pathway elucidations. Nevertheless, major degradation pathways can be deduced based on some key transformation products that were identified in these studies (Fig. 2.9). Hydroxyl radical was hypothesized as a common and one of the primary oxidants. First generation products are most likely produced by radical addition of hydroxyl radical to the aromatic ring (e.g. diethyl 4-hydroxyphthalate) and hydrogen atom abstraction with hydroxyl radical from the ethyl and methyl group of diethyl phthalate and dimethyl phthalate, respectively, yielding deethylation and demethylation (e.g., 2-(ethoxycarbonyl)benzoic acid and 2-(methoxycarbonyl)benzoic acid, respectively, and also phthalic acid). Further decarboxylation and hydroxylation produce common transformation products such as hydroquinone and pyrocatechol, and these transformation products can in turn be oxidized by ring cleavage to form several smaller carboxylic acids (e.g. oxalic acid and maleic acid). In chloride-containing matrix, significant concentrations of chlorophenols (i.e. 3-chlorophenol, 2,4-dichlorophenol, and 2,4,6-trichlorophenol) were measured, together with 4-chlorobenzene-1,3-diol (Souza et al. 2013; Souza et al. 2014c). Here, chlorine can be an important oxidant which can be electrochemically formed at anode. Moreover, hypochlorite, another very likely oxidant, can be produced from chlorine through disproportionation. These chloro-containing species are in general less oxidative, which makes them more difficult to remove by electrochemical treatments (Souza et al. 2014a). Additionally, such compounds can have higher toxicity than the parent compound. Therefore, absence of chloride in solution is preferred and in general better results were obtained with sulfate-containing matrix. Studies concerning transformation products after electrochemical treatment are scarce for 1H-benzotriazole. Similarly to phthalic esters, researchers predominantly looked at the removal efficiency of 1H-benzotriazole and assessed the mineralization rate from COD and TOC determinations. Comparing UV-vis spectra and HPLC-UV chromatograms, two main transformation products were identified in sulfate-containing matrix (Fig. 2.11): 2-aminobenzenediazonium and aniline (Ding et al. 2010). Additionally, benzenediazonium was detected using LC-MS at m/z 105

2  Electrochemical Treatments for the Removal of Emerging Contaminants

179

by the same research group. Three nitro-containing compounds (2-nitroaniline, 1,2-dinitrobenzene, and 2-nitrophenol) were also identified in Na2SO4 using GC-MS together with theoretical calculations (Xing et al. 2012). Our research group found additional transformation products in sulfate- and chloride-containing matrix, using LC-MS/MS (Prosen et al. 2020). In sulfate-containing matrix additional hydroxybenzotriazoles and aminophenol were identified. As for chloride-­containing matrix, chloro-benzotriazoles were found as one would expect.

2.7.2  Toxicity of Treated Wastewater Any determination of treatment efficiency of selected treatment procedure must include an evaluation of toxicity and biodegradability before and after as key factors describing the actual environmental impact of substances. To select the appropriate toxicity determination approach, knowledge on the fate of the treated effluent is important. If only changes in toxicity before and after investigated treatment are considered, simple automatized tests such as bioluminescence inhibition assay with Vibrio fischeri are feasible. If there is further biological treatment, biotests with a mixed culture of microorganisms (activated sludge) should be used, as these play an essential role in a biological wastewater treatment plant. Good choices are the test with inhibition of oxygen consumption measurement and the test measuring the inhibition of growth of the activated sludge (ISO 15522:1999; ISO 8192:2007). If the treated effluent undergoes an anaerobic biological treatment, then the test with the reduction of biogas production is often applied (ISO 13641-1:2003). To determine the actual reduction of toxicity due to the applied treatment procedure, toxicity assessment should involve a battery of biotests using species from different levels of the aquatic ecosystem: decomposers (bacteria), producers (algea Scenedesmus subspicatus or duckweed Lemna minor – Fig. 2.12), and consumers (brine shrimp

Fig. 2.12  Lemna minor toxicity test

180

B. N. Malinović et al.

Artemia salina; water flea Daphnia magna; fish Brachidanio rerio), as well as species belonging to the terrestrial ecosystem (white mustard Sinapis alba; white onion Allium cepa; earthworms Tubifex tubifex). In such a case, bioassays are a good complement to chemical analyses. For the toxicity assessment of selected emerging contaminants, various bioassays were employed. Some of the recent studies mainly focused on those presented in Table  2.6. However, at the time of this review, no application of a battery of biotests with organisms from different trophic levels for the evaluation of treatment efficiency was noticed in papers dealing with treatments of emerging contaminants. Mainly one or two biotests were applied (Borowska et al. 2016; Han et al. 2018; Aravind et al. 2019; Xiao et al. 2020) (Table 2.6), and most often, an ecotoxicological assessment of the treatment procedure was not even included in the study (Pan et al. 2019; Xu et al. 2019; Murrell and Dorman 2020). During AOPs and electrochemical advanced oxidation processes, the formation of toxic metabolites has already been noticed by several authors (Aravind et  al. 2019; MubarakAli et  al. 2019) and in spite of good removal rates of contaminants confirmed, sometimes even with substantial mineralization, toxicity has increased. All of the available papers also lack a comprehensive biodegradability study to address the actual hazard of treated effluents in a natural environmental compartment, related to possible bioavailability and (bio)accumulation. However, the problem of formation of more problematic transformation products is often emphasized and the main conclusion is, that further research is needed (Murrell and Dorman 2020; Wu et al. 2020). Murrell and Dorman (2020) investigated the applicability of conventional disinfection processes regarding their behavior and formation of problematic transformation products using comprehensive two-dimensional gas chromatography. A study based on the determination of six benzotriazoles (4-methyl-1H-benzotriazole, 5-methyl-1H-benzotriazole, and four chloromethyl isomers) in actual municipal wastewaters treated by means of primary treatment, then divided in two paths for the secondary treatment: (i) conventional activated sludge system, and (ii) trickling filter with nutrient removal. Afterwards, effluents were merged for subsequent chlorine disinfection and sprayed for irrigation. None of the six target benzotriazoles were removed completely during the wastewater treatment process. It was confirmed that their concentrations in the effluent remained in the range of 1.40–46.70 μg/L and even some chlorinated transformation products were identified. Wu et  al. (2020) used an innovative approach of coupling falling water film dielectric barrier discharge plasma with peroxymonosulfate or persulfate for the degradation of benzotriazoles. The degradation of investigated compounds was significantly improved (47%) after the addition of peroxymonosulfate, increasing the formation of active radicals that led to improved oxidation potential of plasma system. Toxicity, which was calculated by the USA Environmental Protection Agency Toxicity Estimation Software Tool, was based on Quantitative Structure-Activity Relationship methodology; a correlation between the structure of the compound and lethal oral rat dose (LD50) was determined.

2  Electrochemical Treatments for the Removal of Emerging Contaminants

181

Table 2.6  Organisms applied for the assessment of toxicity of selected emerging contaminants Organisms Compound Bisphenol A Decomposers Mixed culture of aromatic compounds-adapted microorganisms Escherichia coli 1H-benzotriazole Microtox® Producers

Plant Hydroponic lettuce, tomato, durum wheat, broad bean Plant Arabidopsis thaliana Duckweed Lemna minor Microalgae Pseudokircheriniella reticulatum Scenedesmus acutus Carteria cerisiformis Gonium pectoral Scenedesmus quadricauda Coleastrum reticulatum Cyanophora paradoxa Duckweed Lemna minor

Microalgae Cyclotella caspia

Diatom Novicula incerta

Type of the test/ ecosystem References Acute/wastewaters Aravind et al. (2019)

Acute/seed germination

Ye et al. (2018) Borowska et al. (2016) Xiao et al. (2020)

Bisphenol A

Acute/seed germination

Bahmani et al. (2020)

Bisphenol A

Acute/freshwater

Bisphenol A

Acute/freshwater

MubarakAli et al. (2019) Nakajima et al. (2007) Xiao et al. (2020)

1-H-benzotriazole 4-methyl-1H-­ benzotriazole 5-methyl-1H-­ benzotriazole Xylyltriazole 5-chlorobenzotriazole Bisphenol A

Acute/freshwater

Gatidou et al. (2017)

Acute/marine

Bisphenol A

Acute/marine

Li et al. (2008) Xiao et al. (2020) Liu et al. (2010) Xiao et al. (2020)

Benzotriazoles, benzothiazoles Bisphenol A

Stress resistant proteins Acute/marine

(continued)

182

B. N. Malinović et al.

Table 2.6 (continued) Organisms Consumers

Water flea Daphnia magna Rainbow trout cell lines (gill and liver) Onchorhynchus mykiss

Earthworm Eisenia fetida Gastropod Pomacea lineata

Compound Bisphenol A

Type of the test/ ecosystem Acute/freshwater

Benzotriazoles: 1H-benzotriazole 5-chlorobenzotriazole 1-hydroxybenzotriazole 5,6-dimethyl-1H-­ benzotriazole monohydrate Tolyltriazole 1H-benzotriazole

Cytotoxicity Oxidative stress Genotoxicity

Bisphenol A

Acute-chronic/ fresh water

Cu accumulation

References Hu et al. (2019) Zeng et al. (2016)

Xing et al. (2018) de Andrade et al. (2017)

The environmental impact of the effluents after the treatment procedures should be assessed by biodegradability studies to holistically identify environmentally relevant recalcitrant and toxic transformation products. The main tests for biodegradability are various non-standardized laboratory or pilot-scale long-term tests, using activated sludge as a source of active microorganisms. These tests are applied both for the testing of wastewaters and of pure chemicals. To assess biodegradability in common environmental conditions, determination of ready biodegradability is usually enough. Readily biodegradable substances or wastewater mineralize in common environmental conditions rapidly (less than 28 days) after a short lag phase and with non-hazardous transformation products. The substance represents the only source of food for non-adapted microorganisms. Laboratory tests are based on the measurement of sum parameters, e.g. chemical oxygen demand, dissolved organic carbon removal, oxygen consumption, etc. Tests are simple and cost-effective and can be accomplished almost in any environmental laboratory gathering more relevant information than Qualitative Structure Biodegradability Relationship computational modelling, which is also possible.

2.7.3  Cost-Effectiveness Analysis of Electrochemical Treatments Energy consumption is typically assessed as one of the large contributors to total costs when wastewater treatment is considered (de Haas and Dancey 2015). The reason for this is, that the energy assessment is easy to compare and allows to identify areas for improvement quite easily. In many studies operational costs are often only “roughly” calculated not including maintenance and repair costs. The industrial application of one of the treatments is the decision of the investor, where the

2  Electrochemical Treatments for the Removal of Emerging Contaminants

183

most important factors in making the decision are the efficiency and capital costs of the treatment plant. Therefore, it is important to include working capital, direct and indirect costs as part of the total capital investment. When operational costs are estimated, two main approaches can be seen: i) load-­ specific basis, where the mass of the pollutant in the influent is related to the catchment population served (“per capita”) or by “per equivalent population”, PE (kWh/ PE⋅year); and by ii) volume/flow-based load (kWh/m3⋅year). In the literature, operating costs are most commonly expressed as energy required to remove mass of certain pollutant in raw influent, kWh/kgpollutant; COD, kWh/kgCOD; TOC, kWh/ kgTOC; or energy required to treat specific volume of wastewater, kWh/m3. Therefore, it is very difficult to make comparative analyses of processes except for removal efficiency of the same pollutant/wastewater type. Wastewater treatment plant comparison of energy efficiency on a load-specific basis is preferable, because wastewater composition is taken into the account. Unfortunately, there is no uniform approach of assessing energy efficiency for all types of the wastewater treatment plants, therefore, the final decision has to be done by the user. A lot has been done lately in the area of conventional municipal wastewater treatment plants in terms of energy efficiency comparison, but there is a lack of studies in the field of other advanced treatment processes. The ratio between removal of emerging contaminants and operational costs has yet to be established. In selected 11 studies on electrooxidation of the real wastewater from the textile industry, consumption of electricity was from 0.26 to 100  kW/h (Garcia-Segura et al. 2018). In three studies of the process of electrodialysis of different wastewaters, electricity consumption ranged from 0.16 to 21.3 kWh/m3, four different studies show 1.7 to 17.5 kWh/m3 of energy consumption for electrocoagulation, while three studies confirmed 0.105 to 3.05 kWh/m3 of energy consumption for capacitive deionization (Simon et al. 2018). The above data show a huge scale of electricity requirements for the same processes, depending on the type of the wastewater and process parameters. Not many papers comparing the costs of conventional and electrochemical treatments can be found and due to their scarcity they are very valuable. For example, industrial plants are developed for electrocoagulation in which costs for chemicals are reduced ten-fold in comparison to chemical precipitation technique, 3.2-fold relative to chemical coagulation of textile wastewater or nearly two-fold compared to chemical coagulation for the treatment of metallurgical wastewater (Moussa et al. 2017). Although classical electrochemical treatments are known and applied for a longer period of time, a successful scale-up from a laboratory to full scale is still a great challenge. An example of a successful scale-up is given in Fig. 2.13. There are many suggestions for reducing operational costs such as the development of electrodes with a large active surface (Moussa et al. 2017; Rondinini et al. 2018), as well as modification of the distance between electrodes having a linear influence on the electrolyte resistance and thus on the electricity consumption (Santiago Santos et al. 2018). Most of the research is focused on efficacy of the removal of pollutants, not on the design optimization of the reactor. Classical electrochemical treatments can

184

B. N. Malinović et al.

Fig. 2.13  Diagram of an electrocoagulation treatment plant

be used with other treatments approaches and techniques as pre-treatment or post-­ treatment (Santiago Santos et  al. 2018). By studying the characteristics of each wastewater, the best combination can be chosen. The most interesting studies are those that show the cost-effectiveness of a certain treatment approach in terms of reactor design development and the possibility of a scale-up process. The treatment of pharmaceutical pollutants in water by the electro-Fenton process has been studied with focus on cost, efficiency and in combination with biological treatment (Monteil et al. 2019). They presented an overview of advanced reactor configuration that enabled higher treatment efficiency at reduced costs. The geometric parameters, out-streaming, the way of supplying oxygen, energy consumption, and microfluidic reactors as new types of reactors are described in detail. It was confirmed that reactor configuration is a key point for the electro-Fenton efficiency. Although electro-Fenton is effective in the treatment of pharmaceutical pollutants, the process costs are high (at least 0.39  Wh/gTOC). A combination of electro-Fenton as pretreatment and biological treatment has synergistic effects on increasing efficiency and reducing operating costs. All experiments were performed on a model and real wastewater at a laboratory scale and at batch conditions. One of the conclusions is that experiments need to be carried out at a pilot scale to achieve effective electro-Fenton treatment at the industrial level. In terms of engineering, the need to research hydrodynamic behavior of the wastewater and electrochemical reactions to predict the efficiency of reactors is emphasized. Another interesting study was performed by Martínez-Huitle et al. (2015). They did a review of single and coupled processes and reactors for the treatment of different organic pollutants. In order to make a successful electrolytic treatment of wastewater in terms of efficiency and economy, they point out that cell design is a key factor that involves considering mass and heat transport, reactions kinetics, and current efficiency. This factor is especially significant when performing a scale-up process at an industrial scale. The principles of dimensional analysis applicable to scale-up of chemical reactors cannot be applied in the design of an electrochemical reactor because not all the

2  Electrochemical Treatments for the Removal of Emerging Contaminants

185

similarity criteria match. The scale-up principle of chemical processes requires the values corresponding to dimensionless groups of the two units similar to be obtained. Commonly used criteria are geometric, kinematic, and thermal similarities between reactors. In the case of an electrochemical reactor, an additional criterion for scale­up is the current/potential similarity (Goodridge and Scott 1995). The geometric similarity criterion for electrochemical cells cannot be fulfilled, as increasing distances and electrodes would result in a large voltage drop and an increase in energy consumption. Thermal similarity is difficult to apply because of the effect of Joule heating inside the reactor. Current/potential similarity requires a constant ratio of potential to current density, where the current and potential distribution is very important. It is well known in electrochemical engineering that the current distribution can be classified as primary, secondary, and tertiary current distribution. A tertiary current distribution occurs when the current is controlled by mass transfer (Goodridge and Scott 1995). In addition, an important parameter is the mass transfer at the electrode surface where the reactions take place. Most often, mass transfer is the bottleneck of all electrochemical processes, including wastewater treatments. If there is a requirement for complete mineralization of the treated effluent, process optimization is necessary due to high operating costs. In this case as well as in the case of complex wastewater, electrochemical treatments can be successfully combined with other treatments which leads to greater process efficiency and increased process economy. Process optimization of wastewater treatment plants is an important task to achieve a goal of cost efficient wastewater treatment process and to extend the lifetime of the existing infrastructure. Optimizing the process in terms of treatment efficacy and acceptable costs can be as simple as maintaining optimal values of technological parameters (concentrations of reagents, pH, temperature, retention time, wastewater flows and loads, etc.) while in the case of different combination of techniques more effort is needed to identify the potential for such optimizations. This is especially true when new treatment methods like electrochemical processes in all possible varieties are implemented at large industrial scale and combined with well established conventional processes. Valuable contribution of operators that have been working on those plants helps engineers to identify opportunities for process optimization.

2.8  O  ther Advanced Oxidation Processes for the Removal of Pollutants Electrochemical treatments are arguably the least applied of the available treatments for the removal of pollutants from wastewater and other sources of pollution. Other available treatments were already mentioned in the Sect. 2.2.3. Here, we shortly present the most frequently applied treatments for the removal of emerging contaminants.

186

B. N. Malinović et al.

In the case of confirmed bio-treatability in terms of low toxicity and significant biodegradability, various biological processes could be applied for removal of emerging contaminants. Biological treatment is simple, reliable, and highly cost-­ effective with negligible environmental impacts. It can be performed in different types of reactors, both at aerobic and anaerobic conditions, and applying various technical solutions in batch, semi-continuous, or continuous mode. Conventional systems with activated sludge, sequencing batch reactors, bio-filters, membrane bioreactors, and moving bed membrane bioreactors, additionally also up-flow anaerobic sludge blanket processes and fluidized bed reactors are well established and through optimization could be effective enough (Tang et al. 2019). Combinations of conventional activated sludge processes with membrane separation increase the performance of treatment processes in terms of lowering the legislative effluent parameter as well as reducing micro-pollutant environmental release. Slowly growing nitrifying bacteria requires high sludge retention time which leads to better removal of biodegradable micro-pollutants. It was confirmed that membrane bioreactor systems are more efficient than conventional activated sludge systems, especially for compounds that are not readily biodegradable (Park et  al. 2017). Sorption, which was estimated as an important accompanying mechanism, removes mainly lipophilic compounds, and it actually means only the transport of pollutants from aqueous to solid phase (microorganisms) in the treatment process. Adsorbed pollutants can seriously affect the further management of biosolids (composting and biogas production via waste sludge anaerobic digestion) and ways of its final disposal (Boševski et al. 2019). However, typical characteristics of emerging contaminants reduce their susceptibility to biological treatment (Tiwari et  al. 2017). Danner et  al. (2019) clearly showed that antibiotics, as one of the most daunting groups of emerging contaminants, are detected in various environmental compartments world-wide. However, to relate the extent of treatment of industrial and municipal wastewater in particular country or region to the measured concentrations of emerging contaminants in the environment is not reliable or even not possible. Simultaneously, antibiotic resistant genes resulting from the release of non-­ adequately treated effluents are labeled as emerging contaminants because their entrance into the environment promotes their establishment, resulting in increased risks to be transmitted to human pathogens (Corno et al. 2019). Wastewater treatment plants can be labeled as hot spots for emerging contaminants and their upgrade for removal of those pollutants should be a priority (Sabri et al. 2018: Eramo et al. 2019). They pose risks not only to the receiving compartments but through possible use for cattle, irrigation, or drinking water production (Sabri et al. 2018). Chemical oxidation processes are effective options for the treatment of wastewaters containing emerging contaminants to prevent their local or global impacts. Their main drawback is that they are relatively expensive for complete mineralization of pollutants, the principal reason being that the oxidation intermediates into which the emerging contaminants are transformed during treatment tend to be progressively more resistant to complete (bio)chemical degradation (Sharma et  al. 2018). However, the combination of oxidative processes and biological treatment

2  Electrochemical Treatments for the Removal of Emerging Contaminants

187

has been investigated as one of the attractive potential alternatives. Fenton oxidation or ozonation modified in numerous possible ways can be applied as a pre-treatment process to achieve the conversion of the persistent and non-biodegradable organics into more biodegradable intermediates, which would be consequently successfully treated by the biological treatment process. On the other hand, oxidation processes can also be very effective in terms of polishing the final effluent to remove remaining emerging contaminants and in different treatment stages formed by-products. Advanced oxidation processes (AOPs) are technologies involving in situ generation of hydroxyl radicals (OH.) and proved to be effective for the degradation of various emerging contaminants, such as antibiotics, personal care products, pharmaceuticals, plasticizers, additives, etc. They combine high effectiveness of pollutant removal through various oxidation reactions and highly efficient generation of non-selective hydroxyl radicals with a high oxidation potential (2.8  V vs. SHE), often enhanced by catalytic (TiO2, ZnO, WO3, Bi2O3, MoO3, etc.) and photocatalytic processes (UV or visible light). Reactions of oxidation of pollutants could be single- or multistep, depending primarily on their physical and chemical properties, such as chemical complexity, volatility and/or hydrophilicity. Hydroxyl radicals can be produced in situ using a variety of chemical processes. Generally, there are two main approaches that are commonly used: (i) generation of hydroxyl radicals through chemical reactions; and (ii) photochemical generation of hydroxyl radicals. Chemically, hydroxyl radicals can be generated using strong oxidants such as hydrogen peroxide and ozone or by their combination, so called peroxone. Ozone can also be combined with the hydroxide anion, persulfate and monopersulfate. Susceptibility of micropollutants to ozone depends on their chemical characteristics and structure. Based on their ozone rate constants, they can be classified in three groups (Lee et al. 2014; Ur Rehman et al. 2019): (i) ozone reactive: kO3 >  104 M/s (ii) moderately ozone reactive: kO3 50–104 M/s (iii) ozone refractory: kO3 ≤ 50 M/s Other reagents used for the chemical generation of hydroxyl radicals include the Fenton and Fenton-like processes, based on the combination of oxidant (H2O2) and Fe in different forms as catalyst (Fe2+/Fe3+, zero-valent Fe0, Fe-TiO2, α-FeOOH, RhFeSi, etc.). The process can be conducted both homogeneously and heterogeneously also with the presence of radiation. However, a major operational issue for the hydroxyl radical-based AOP is high instability of radicals requiring on site production limiting their technological feasibility (Ribeiro et al. 2015). In photochemical processes (Fig. 2.14), mainly UV light is used along with H2O2, ozone, and other catalysts to generate hydroxyl radicals in the cases where the conventional H2O2/O3 combination is not efficient enough. UV energy can be helpful in the destruction of organic compounds which are not transformed in its absence (Sharma et al. 2018). Main technologically feasible approaches are a combination of (i) ozone/UV, (ii) hydrogen peroxide/UV, (iii) photo-Fenton and Fenton-like reactions, and (iv) photocatalytic oxidation (UV/TiO2). Ozone has a molar absorptivity of 3300 1/M·cm and strongly absorbs UV light at 245  nm. It produces

188

B. N. Malinović et al.

Fig. 2.14  UV treatment of industrial wastewater for reuse in a closed loop system. (Source: Matej Čehovin)

hydrogen peroxide which further decomposes into OH., subsequently producing more radicals. H2O2/UV combination also enhances the formation of OH. radicals which propagate a chain reaction and finally lead to the formation of O2 and H2O. The efficiency of the H2O2 photolysis depends on the absorption of UV energy based on H2O2 molar absorptivity (19.6 1/M·cm), which is usually much lower than for organic compounds present in water. If H2O2/UV is combined with ozone, the generation of OH. is even greater. The combination of Fenton (Fe2+/Fe3+ with H2O2) with UV/visible radiation is sometimes used to remove organic pollutants from drinking water. Fe3+ is converted to Fe2+ by photo-reduction, which than again reacts with peroxide and increases the rate and efficiency of the process. In photocatalytic oxidation, TiO2 as a metal oxide semiconductor has been found as the most effective catalyst, which can be applied in the slurry phase or immobilized on some support. It is excited after absorption of UV light and produces conduction band electrons and valence band holes. Holes are able to oxidize a variety of organic compounds due to their extremely positive oxidation potential. In some cases, UV can also be combined with chlorine to generate OH., Cl., Cl2− ., and ClO.. Hydrodynamic cavitation was found to have a synergistic effect on AOPs, utilizing hydrogen peroxide and derivatives (Braeutigam et al. 2012; Zupanc et al. 2013; Patil et al. 2014). The application of hydrodynamic cavitation in effluent treatment ensures a substantial increase in the efficiency of oxidation of contaminants, as

2  Electrochemical Treatments for the Removal of Emerging Contaminants

189

demonstrated by a number of literature reports based on recent research in treatment of industrial effluents involving the use of hybrid processes; conventional AOPs combined with cavitation (Raut-Jadhav et al. 2013; Gogate et al. 2014; Choi et al. 2019). Cavitation is the formation, cyclic growth and rarefaction leading to the terminal implosive collapse of vapor bubbles in the liquid phase. It can be generated by different triggers: turbulent flows, fast moving particles, boiling, electrical discharge, and laser or ultrasonic irradiation. When the collapse of the cavitation bubbles happens, extreme temperatures (up to 5000 °C inside cavitation bubble and up to 2000 °C at the bubble-liquid interface) and pressures (approx. 500–1000 bar) can be locally generated for a very brief time of micro- to milliseconds (Gogate et al. 2014). Therefore, the effects of cavitation can be used for degradation of pollutants due to both mechanical (extreme shear forces in the bulk media) and chemical effects in the fluid (Raut-Jadhav et al. 2013; Čehovin et al. 2017). The usual way to generate the hydrodynamic cavitation is by the contractions in the flow (e.g. Venturi) or by the constrictions in the form of nozzles, orifice plates, and rotation elements. Chemical reactions could involve pyrolysis reactions and radical reactions that can lead to the formation of hydroxyl (OH.) and hydroperoxyl (HOO.) radicals and hydrogen peroxide (H2O2). Hydrodynamic cavitation can be efficiently coupled with other AOPs (H2O2/UV; O3/UV; H2O2/O3; etc.) to obtain improved oxidative performance and removal efficiency. This can be explained by more intensive turbulence, especially radial, leading to dispersion with better distribution of oxidative species (Čehovin et al. 2017).

2.9  Conclusion Conventional methods of wastewater treatment are less effective for the removal of emerging contaminants in contrast to electrochemical treatments, which are the leading eco-friendly technique within the advanced oxidation processes as well. Electrons are the main reactive species in the electrochemical treatments and cause the further chemical transformation of contaminants. Therefore, electrochemical treatments do not require the additional use of chemicals (e.g., metal salts and polyelectrolytes as coagulants, or hazardous oxidizing agents). The proper choice of technique and electrode material can in some cases completely avoid the formation of by-products (secondary pollution) and allow for the complete removal and even mineralization of a wide range of different pollutants. The equipment is relatively simple, easily amenable to automation and modular adaptation. The investment costs of electrochemical treatments can be high if advanced electrode materials are used (e.g. diamond electrodes), and for cheaper materials, the life span is generally short, therefore, replacement is required. The highest operating costs relate to the consumption of electrical energy and the replacement of sacrificial electrodes, however, renewable energy from wind and solar sources can be used for electrochemical treatments. Moreover, there is ongoing research on the utilization of recycled material from e.g. spent batteries for the fabrication of

190

B. N. Malinović et al.

electrodes. Some examples are given in the present review as well. Low wastewater conductivity requires the addition of a supporting electrolyte (non-toxic substances), but most wastewaters already contain salts and have a satisfactory conductivity. Nowadays, electrochemical treatments have found application in many pilot plants, but still in a small number of large-scale applications, except for larger applications for the automatic disinfection of swimming pool water, both private and public (Radha and Sirisha 2018). Several parameters need to be improved for electrochemical treatments to achieve a greater industrial application, such as: electrocatalytic activity of the electrodes, reactor design and hydrodynamics, electrode arrangement, and current efficiency. Different pollutants and different types of wastewater require different optimal process parameters, making it difficult to standardize the process, i.e., the widespread use of standard devices and equipment. Future research should cover the economic feasibility of electrochemical advanced oxidation processes, including investment costs (e.g., for the reactor, UV lamps, and photo-reactors) and operating costs, and in particular electrical energy, reagents, and maintenance costs. The coupling with other treatments (e.g. conventional) as a pre- or post-­treatment unit is the most promising state of the art for the current application of electrochemical treatments. Also, coupling with other advanced oxidation processes can improve the degradation efficiency of pollutants because of the synergetic effect. Moreira et al. (2017) in their review paper covering over 100 different studies of electrochemical advanced oxidation processes for the treatment of dyes, pesticides, pharmaceuticals, and synthetic and real wastewaters conclude that electrochemical advanced oxidation processes are more effective than their chemical analogues, showing higher removal efficiency and lower energy consumption. In this chapter, we have shown the advancements of electrochemical treatments for wastewater and similar matrices containing selected emerging contaminants: bisphenol A, phthalic acid esters, and benzotriazoles, which arise from intensive plastics production. The common conclusion is that a complete mineralization of target contaminants is usually not achieved due to substantial energy requirements needed for the actual wastewater treatment. This hindrance can arise in part due to the other coexisting chemical species in wastewater (e.g. natural organic matter, anions, particulate matter) that may compete with the target contaminants in consuming the hydroxyl radicals and other reactive species produced in electrochemical treatments. Because of the potential formation of unwanted and possibly more toxic and refractory transformation products during the electrochemical treatment, degradation pathways need to be elucidated and toxicological assessment need to be performed, not only for the target contaminants, but for their transformation products. Most of the reviewed studies using electrochemical treatments were found to be lacking in the above aspects, using only total organic carbon or chemical oxygen demand monitoring or, at the most, HPLC-UV (or diode array detector) or low resolution mass spectrometry. Today, tandem mass spectrometry or high resolution mass spectrometry are the techniques of choice for the monitoring in degradation studies, because they enable far better identification of the transformation products and elucidation of results. Therefore, additional studies with tandem mass spectrometry or

2  Electrochemical Treatments for the Removal of Emerging Contaminants

191

high resolution mass spectrometry need to be done in the field of electrochemical treatments, at least for the contaminants covered in the present review. Acknowledgements and Funding  The authors wish to acknowledge the funding by research grants P1-0153 and P2-0191 (Slovenian Research Agency, Slovenia), as well as by project grants 19.032/961-49/19 and 19/6-020/964-14-2/18 (Ministry of Scientific and Technological Development, Higher Education and Information Society Government of Republic of Srpska, Bosnia and Herzegovina). Photos provided by Matej Čehovin are gratefully acknowledged.

References Abo R, Kummer N-A, Merkel BJ (2016) Optimized photodegradation of Bisphenol A in water using ZnO, TiO2 and SnO2 photocatalysts under UV radiation as a decontamination procedure. Drink Water Eng Sci 9:27–35. https://doi.org/10.5194/dwes-­9-­27-­2016 Adyel TM, Rahman SH, Zaman MM, Sayem HM, Khan M, Abdul Gafur M, Islam SMN (2013) Reuse feasibility of electrocoagulated metal hydroxide sludge of textile industry in the manufacturing of building blocks. J Waste Manag 686981:1–9. https://doi.org/10.1155/2013/686981 Ahmadi M, Ghanbari F (2018) Degradation of organic pollutants by photoelectro-peroxone/ZVI process: synergistic, kinetic and feasibility studies. J Environ Manag 228:32–39. https://doi. org/10.1016/j.jenvman.2018.08.102 Ahmadi M, Ghanbari F, Madihi-Bidgoli S (2016) Photoperoxi-coagulation using activated carbon fiber cathode as an efficient method for benzotriazole removal from aqueous solutions: modeling, optimization and mechanism. J Photochem Photobiol A Chem 322–323:85–94. https://doi. org/10.1016/j.jphotochem.2016.02.025 Al-Amshawee S, Yunus MYBM, Azoddein AAM, Hassell DG, Dakhil IH, Hasan HA (2020) Electrodialysis desalination for water and wastewater: a review. Chem Eng J 380:122231. https://doi.org/10.1016/j.cej.2019.122231 Alikarami M, Darvishi Cheshmeh Soltani R, Khataee A (2019) An innovative combination of electrochemical and photocatalytic processes for decontamination of bisphenol A endocrine disruptor form aquatic phase: insight into mechanism, enhancers and bio-toxicity assay. Sep Purif Technol 220:42–51. https://doi.org/10.1016/j.seppur.2019.03.056 Ameta RK, Chohadia A, Jain A, Punjabi PB (2018) Chapter 3: Fenton and photo-fenton processes. In: Ameta SC, Ameta R (eds) Advanced oxidation processes for waste water treatment. Academic Press, pp 49–87. https://doi.org/10.1016/B978-­0-­12-­810499-­6.00003-­6 Antolini E (2019) Photoelectrocatalytic fuel cells and photoelectrode microbial fuel cells for wastewater treatment and power generation. J Environ Chem Eng 7:103241. https://doi. org/10.1016/j.jece.2019.103241 Aquino JM, Parra KN, Miwa DW, Motheo AJ (2015) Removal of phthalic acid from aqueous solution using a photo-assisted electrochemical method. J Environ Chem Eng 3:429–435. https:// doi.org/10.1016/j.jece.2015.01.009 Aravind P, Devarajan A, Solaiappan A, Selvaraj H, Sundaram M (2019) Removal of BPA from thermal cash receipts via electro oxidation cum biodegradation: evaluating its degradation mechanism and in silico toxicity analysis. J Water Process Eng 31:100849. https://doi. org/10.1016/j.jwpe.2019.100849 Arola K, Ward A, Mänttäri M, Kallioinen M, Batstone D (2019) Transport of pharmaceuticals during electrodialysis treatment of wastewater. Water Res 161:496–504. https://doi.org/10.1016/j. watres.2019.06.031 Asimakopoulos AG, Ajibola A, Kannan K, Thomaidis NS (2013) Occurrence and removal efficiencies of benzotriazoles and benzothiazoles in a wastewater treatment plant in Greece. Sci Total Environ 452-453:163–171. https://doi.org/10.1016/j.scitotenv.2013.02.041

192

B. N. Malinović et al.

Babuponnusami A, Muthukumar K (2012) Advanced oxidation of phenol: a comparison between Fenton, electro-Fenton, sono-electro-Fenton and photo-electro-Fenton processes. Chem Eng J 183:1–9. https://doi.org/10.1016/j.cej.2011.12.010 Bahmani R, Kim D, Modareszadeh M, Thompson AJ, Park JH, Yoo HH, Hwang S (2020) The mechanism of root growth inhibition by the endocrine disruptor bisphenol A (BPA). Environ Pollut 257:113516. https://doi.org/10.1016/j.envpol.2019.113516 Bakr AR, Rahaman MS (2017) Removal of bisphenol A by electrochemical carbon-nanotube filter: Influential factors and degradation pathway. Chemosphere 185:879–887. https://doi. org/10.1016/j.chemosphere.2017.07.082 Bakr AR, Rahaman MS (2019) Crossflow electrochemical filtration for elimination of ibuprofen and bisphenol a from pure and competing electrolytic solution conditions. J Hazard Mater 365:615–621. https://doi.org/10.1016/j.jhazmat.2018.11.015 Barrera-Diaz CE, Balderas-Hernandez P, Bilyeu B (2018) Chapter 3: Electrocoagulation: fundamentals and prospectives. In: Martínez-Huitle AC, Rodrigo MA, Scialdone O (eds) Electrochemical water and wastewater treatment. Butterworth-Heinemann, pp 61–76. https:// doi.org/10.1016/B978-­0-­12-­813160-­2.00003-­1 Barroso PJ, Santos JL, Martín J, Aparicio I, Alonso E (2019) Emerging contaminants in the atmosphere: analysis, occurrence and future challenges. Crit Rev Env Sci Tec 49:104–171. https:// doi.org/10.1080/10643389.2018.1540761 Bazrafshan E, Mohammadi L, Ansari-Moghaddam A, Mahvi AH (2015) Heavy metals removal from aqueous environments by electrocoagulation process – a systematic review. J Environ Health Sci Eng 13:74. https://doi.org/10.1186/s40201-­015-­0233-­8 Benjamin S, Masai E, Kamimura N, Takahashi K, Anderson RC, Faisal PA (2017) Phthalates impact human health: epidemiological evidences and plausible mechanism of action. J Hazard Mater 340:360–383. https://doi.org/10.1016/j.jhazmat.2017.06.036 Bergh C, Torgrip R, Emenius G, Östman C (2011) Organophosphate and phthalate esters in air and settled dust – a multi-location indoor study. Indoor Air 21:67–76. https://doi. org/10.1111/j.1600-­0668.2010.00684.x Bieber S, Snyder SA, Dagnino S, Rauch-Williams T, Drewes JE (2018) Management strategies for trace organic chemicals in water – a review of international approaches. Chemosphere 195:410–426. https://doi.org/10.1016/j.chemosphere.2017.12.100 Bilal M, Adeel M, Rasheed T, Zhao Y, Iqbal HMN (2019) Emerging contaminants of high concern and their enzyme-assisted biodegradation – a review. Environ Int 124:336–353. https://doi. org/10.1016/j.envint.2019.01.011 Bletsou AA, Jeon J, Hollender J, Archontaki E, Thomaidis NS (2015) Targeted and non-­targeted liquid chromatography-mass spectrometric workflows for identification of transformation products of emerging pollutants in the aquatic environment. Trends Anal Chem 66:32–44. https://doi.org/10.1016/j.trac.2014.11.009 Bokare AD, Choi W (2014) Review of iron-free Fenton-like systems for activating H2O2 in advanced oxidation processes. J Hazard Mater 275:121–135. https://doi.org/10.1016/j. jhazmat.2014.04.054 Borowska E, Felis E, Kalka J (2016) Oxidation of benzotriazole and benzothiazole in photochemical processes: kinetics and formation of transformation products. Chem Eng J 304:852–863. https://doi.org/10.1016/j.cej.2016.06.123 Boševski I, Kalčikova G, Cerkovnik J, Gotvajn AŽ (2019) Ozone as a pretreatment method for antibiotic contaminated wastewater and sludge. Ozone-Sci Eng 42:128–135. https://doi.org/1 0.1080/01919512.2019.1624149 Braeutigam P, Franke M, Schneider RJ, Lehmann A, Stolle A, Ondruschka B (2012) Degradation of carbamazepine in environmentally relevant concentrations in water by hydrodynamic-­acoustic-­ cavitation (HAC). Water Res 46:2469–2477. https://doi.org/10.1016/j.watres.2012.02.013 Brillas E, Sirés I, Oturan MA (2009) Electro-fenton process and related electrochemical technologies based on fenton’s reaction chemistry. Chem Rev 109:6570–6631. https://doi.org/10.1021/ cr900136g

2  Electrochemical Treatments for the Removal of Emerging Contaminants

193

Bunke D, Moritz S, Brack W, Herráez DL, Posthuma L, Nuss M (2019) Developments in society and implications for emerging pollutants in the aquatic environment. Environ Sci Eur 31:32. https://doi.org/10.1186/s12302-­019-­0213-­1 Burgos-Castillo RC, Sirés I, Sillanpää M, Brillas E (2018) Application of electrochemical advanced oxidation to bisphenol A degradation in water. Effect of sulfate and chloride ions. Chemosphere 194:812–820. https://doi.org/10.1016/j.chemosphere.2017.12.014 Cantwell MG, Sullivan JC, Katz DR, Burgess RM, Hubeny JB, King J (2015) Source determination of benzotriazoles in sediment cores from two urban estuaries on the Atlantic Coast of the United States. Mar Pollut Bull 101:208–218. https://doi.org/10.1016/j.marpolbul.2015.10.075 Cao Z, Zheng X, Cao H, Zhao H, Sun Z, Guo Z, Wang K, Zhou B (2018) Efficient reuse of anode scrap from lithium-ion batteries as cathode for pollutant degradation in electro-Fenton process: role of different recovery processes. Chem Eng J 337:256–264. https://doi.org/10.1016/j. cej.2017.12.104 Cao P, Quan X, Zhao K, Chen S, Yu H, Niu J (2020) Selective electrochemical H2O2 generation and activation on a bifunctional catalyst for heterogeneous electro-Fenton catalysis. J Hazard Mater 382:121102. https://doi.org/10.1016/j.jhazmat.2019.121102 Careghini A, Mastorgio AF, Saponaro S, Sezenna E (2015) Bisphenol A, nonylphenols, benzophenones, and benzotriazoles in soils, groundwater, surface water, sediments, and food: a review. Environ Sci Pollut Res 22:5711–5741. https://doi.org/10.1007/s11356-­014-­3974-­5 Carmona E, Picó Y (2018) The use of chromatographic methods coupled to mass spectrometry for the study of emerging pollutants in the environment. Crit Rev Anal Chem 48:305–316. https:// doi.org/10.1080/10408347.2018.1430555 Carson R (2002) Silent Spring, 40th Anniversary edition. Houghton Mifflin Harcourt, 400 p. ISBN: 0618249060. 978-0618249060 Casado J (2019) Towards industrial implementation of Electro-Fenton and derived technologies for wastewater treatment: a review. J Environ Chem Eng 7:102823. https://doi.org/10.1016/j. jece.2018.102823 Čehovin M, Medic A, Scheideler J, Mielcke J, Ried A, Kompare B, Žgajnar Gotvajn A (2017) Hydrodynamic cavitation in combination with the ozone, hydrogen peroxide and the UV-based advanced oxidation processes for the removal of natural organic matter from drinking water. Ultrason Sonochem 37:394–404. https://doi.org/10.1016/j.ultsonch.2017.01.036 Chen GH (2004) Electrochemical technologies in wastewater treatment. Sep Purif Technol 38:11–41. https://doi.org/10.1016/j.seppur.2003.10.006 Chen X, Chen G, Yue PL (2002) Novel electrode system for electroflotation of wastewater. Environ Sci Technol 36:778–783. https://doi.org/10.1021/es011003u Chen Z, Zhang Y, Zhou L, Zhu H, Wan F, Wang Y, Zhang D (2017) Performance of nitrogen-doped graphene aerogel particle electrodes for electro-catalytic oxidation of simulated Bisphenol A wastewaters. J Hazard Mater 332:70–78. https://doi.org/10.1016/j.jhazmat.2017.02.048 Chmayssem A, Taha S, Hauchard D (2017) Scaled-up electrochemical reactor with a fixed bed three-dimensional cathode for electro-Fenton process: application to the treatment of bisphenol A. Electrochim Acta 225:435–442. https://doi.org/10.1016/j.electacta.2016.12.183 Choi J, Cui M, Lee Y, Ma J, Kim J, Son Y, Khim J (2019) Hybrid reactor based on hydrodynamic cavitation, ozonation, and persulfate oxidation for oxalic acid decomposition during rare-earth extraction processes. Ultrason Sonochem 52:326–335. https://doi.org/10.1016/j. ultsonch.2018.12.004 Christou A, Papadavid G, Dalias P, Fotopoulos V, Michael C, Bayona JM, Piña B, Fatta-­Kassinos D (2019) Ranking of crop plants according to their potential to uptake and accumulate contaminants of emerging concern. Environ Res 170:422–432. https://doi.org/10.1016/j. envres.2018.12.048 Collivignarelli MC, Abbà A, Carnevale Miino M, Arab H, Bestetti M, Franz S (2019) Decolorization and biodegradability of a real pharmaceutical wastewater treated by H2O2-assisted photoelectrocatalysis on TiO2 meshes. J Hazard Mater 121668. https://doi.org/10.1016/j. jhazmat.2019.121668

194

B. N. Malinović et al.

Comninellis C (1994) Electrocatalysis in the electrochemical conversion/combustion of organic pollutants for waste water treatment. Electrochim Acta 39:1857–1862. https://doi.org/10.101 6/0013-­4686(94)85175-­1 Corno G, Yang Y, Eckert EM, Fontaneto D, Fiorentino A, Galafassi S, Zhang T, Di Cesare A (2019) Effluents of wastewater treatment plants promote the rapid stabilization of the antibiotic resistome in receiving freshwater bodies. Water Res 158:72–81. https://doi.org/10.1016/j. watres.2019.04.031 Corrales J, Kristofco LA, Steele WB, Yates BS, Breed CS, Spencer Williams E, Brooks BW (2015) Global assessment of Bisphenol A in the environment: review and analysis of its occurrence and bioaccumulation. Dose Response 13:1–29. https://doi.org/10.1177/1559325815598308 Cui W, He J, Wang H, Hu J, Liu L, Liang Y (2018) Polyaniline hybridization promotes photo-­ electro-­catalytic removal of organic contaminants over 3D network structure of rGH-PANI/TiO2 hydrogel. Appl Catal B Environ 232:232–245. https://doi.org/10.1016/j.apcatb.2018.03.069 Danner M-C, Robertson A, Behrends V, Reiss J (2019) Antibiotic pollution in surface fresh waters: occurrence and effects. Sci Total Environ 664:793–804. https://doi.org/10.1016/j. scitotenv.2019.01.406 de Andrade ALC, Soares PRL, da Silva SCBL, da Silva MCG, Santos TP, Cadena MRS, Soares PC, Cadena PG (2017) Evaluation of the toxic effect of endocrine disruptor Bisphenol A (BPA) in the acute and chronic toxicity tests with Pomacea lineata gastropod. Comp Biochem Phys C 197:1–7. https://doi.org/10.1016/j.cbpc.2017.04.002 De Battisti A, Martínez-Huitle CA (2018) Chapter 5: Electrocatalysis in wastewater treatment. In: Martínez-Huitle CA, Rodrigo MA, Scialdone O (eds) Electrochemical water and wastewater treatment. Butterworth-Heinemann, pp 119–131. https://doi.org/10.1016/B978-­0-­12-­813160-­ 2.00005-­5 de Haas D, Dancey M (2015) Wastewater treatment energy efficiency: a review with current Australian perspectives. Water: J AWA 42:53–58. https://doi.org/10.13140/RG.2.1.2493.7363 de Moura DC, de Araújo CKC, Zanta CLPS, Salazar R, Martínez-Huitle CA (2014) Active chlorine species electrogenerated on Ti/Ru0.3Ti0.7O2 surface: electrochemical behavior, concentration determination and their application. J Electroanal Chem 731:145–152. https://doi. org/10.1016/j.jelechem.2014.08.008 Deng B, Li Y, Tan W, Wang Z, Yu Z, Xing S, Lin H, Zhang H (2018) Degradation of bisphenol A by electro-enhanced heterogeneous activation of peroxydisulfate using Mn-Zn ferrite from spent alkaline Zn-Mn batteries. Chemosphere 204:178–185. https://doi.org/10.1016/j. chemosphere.2018.03.194 Ding Y, Yang C, Zhu L, Zhang J (2010) Photoelectrochemical activity of liquid phase deposited TiO2 film for degradation of benzotriazole. J Hazard Mater 175:96–103. https://doi. org/10.1016/j.jhazmat.2009.09.037 Divyapriya G, Nambi IM, Senthilnathan J (2017) An innate quinone functionalized electrochemically exfoliated graphene/Fe3O4 composite electrode for the continuous generation of reactive oxygen species. Chem Eng J 316:964–977. https://doi.org/10.1016/j.cej.2017.01.074 Dong C-D, Chen T-S, Chen C-W, Huang K-L (2016) Electrochemical degradation of diethyl phthalate under different operating conditions. Int J Electrochem Sci 11:5009–5020. https:// doi.org/10.20964/2016.06.77 dos Santos EV, Scialdone O (2018) Chapter 10: Photo-electrochemical technologies for removing organic compounds in wastewater. In: Martínez-Huitle CA, Rodrigo MA, Scialdone O (eds) Electrochemical water and wastewater treatment. Butterworth-Heinemann, pp 239–266. https://doi.org/10.1016/B978-­0-­12-­813160-­2.00010-­9 Dummer NM (2014) 4(5)-Methylbenzotriazole: a review of the life-cycle of an emerging contaminant. Rev Environ Sci Biotechnol 13:53–61. https://doi.org/10.1007/s11157-­013-­9318-­y ECHA Candidate List of substances of very high concern for Authorisation. https://echa.europa. eu/candidate-­list-­table Accessed 4 Feb 2020

2  Electrochemical Treatments for the Removal of Emerging Contaminants

195

Ensano BMB, Borea L, Naddeo V, Belgiorno V, de Luna MDG, Ballesteros FC (2017) Removal of pharmaceuticals from wastewater by intermittent electrocoagulation. Water 9:85. https://doi. org/10.3390/w9020085 Ensano BMB, Borea L, Naddeo V, de Luna MDG, Belgiorno V (2019) Control of emerging contaminants by the combination of electrochemical processes and membrane bioreactors. Environ Sci Pollut Res 26:1103–1112. https://doi.org/10.1007/s11356-­017-­9097-­z Eramo A, Morales Medina WR, Fahrenfeld NL (2019) Viability-based quantification of antibiotic resistance genes and human fecal markers in wastewater effluent and receiving waters. Sci Total Environ 656:495–502. https://doi.org/10.1016/j.scitotenv.2018.11.325 Espinoza JDG, Drogui P, Zolfaghari M, Dirany A, Ledesma MTO, Gortáres-Moroyoqui P, Buelna G (2016) Performance of electrochemical oxidation process for removal of di (2-ethylhexyl) phthalate. Environ Sci Pollut Res 23:12164–12173. https://doi.org/10.1007/s11356-­016-­6304-­2 Fatima R, Afridi MN, Kumar V, Lee J, Ali I, Kim KH, Kim J-O (2019) Photocatalytic degradation performance of various types of modified TiO2 against nitrophenols in aqueous systems. J Clean Prod 231:899–912. https://doi.org/10.1016/j.jclepro.2019.05.292 Feng Y, Yang L, Liu J, Logan BE (2016) Electrochemical technologies for wastewater treatment and resource reclamation. Environ Sci Water Res Technol 2:800–831. https://doi.org/10.1039/ C5EW00289C Fernández-Domene RM, Sánchez-Tovar R, Lucas-granados B, Muñoz-Portero MJ, García-Antón J (2018) Elimination of pesticide atrazine by photoelectrocatalysis using a photoanode based on WO3 nanosheets. Chem Eng J 350:1114–1124. https://doi.org/10.1016/j.cej.2018.06.015 Fernández-Sáez N, Villela-Martinez DE, Carrasco-Marín F, Pérez-Cadenas AF, Pastrana-­Martínez LM (2019) Heteroatom-doped graphene aerogels and carbon-magnetite catalysts for the heterogeneous electro-Fenton degradation of acetaminophen in aqueous solution. J Catal 378:68–79. https://doi.org/10.1016/j.jcat.2019.08.020 Ferrag-Siagh F, Fourcade F, Soutrel I, Aït-Amar H, Djelal H, Amrane A (2014) Electro-­Fenton pretreatment for the improvement of tylosin biodegradability. Environ Sci Pollut Res 21:8534–8542. https://doi.org/10.1007/s11356-­014-­2771-­5 Fierens T, Servaes K, Van Holderbeke M, Geerts L, De Henauw S, Sioen I, Vanermen G (2012) Analysis of phthalates in food products and packaging materials sold on the Belgian market. Food Chem Toxicol 50:2575–2583. https://doi.org/10.1016/j.fct.2012.04.029 Fischer A, van Wezel AP, Hollender J, Cornelissen E, Hofman R, van der Hoek JP (2019) Development and application of relevance and reliability criteria for water treatment removal efficiencies of chemicals of emerging concern. Water Res 161:274–287. https://doi. org/10.1016/j.watres.2019.05.088 Ganiyu SO, Martínez-Huitle CA (2019) Nature, mechanisms and reactivity of electrogenerated reactive species at thin-film boron-doped diamond (BDD) electrodes during electrochemical wastewater treatment. ChemElectroChem 6:2379–2392. https://doi.org/10.1002/ celc.201900159 Ganzenko O, Huguenot D, van Hullebusch ED, Esposito G, Oturan MA (2014) Electrochemical advanced oxidation and biological processes for wastewater treatment: a review of the combined approaches. Environ Sci Pollut Res 21:8493–8524. https://doi.org/10.1007/s11356-­014-­2770-­6 García-Córcoles MT, Rodríguez-Gómez R, de Alarcón-Gómez B, Çipa M, Martín-Pozo L, Kauffmann J-M, Zafra-Gómez A (2019) Chromatographic methods for the determination of emerging contaminants in natural water and wastewater samples: a review. Crit Rev Anal Chem 49:160–186. https://doi.org/10.1080/10408347.2018.1496010 Garcia-Segura S, Brillas E (2017) Applied photoelectrocatalysis on the degradation of organic pollutants in wastewaters. J Photochem Photobiol C 31:1–35. https://doi.org/10.1016/j. jphotochemrev.2017.01.005 Garcia-Segura S, Keller J, Brillas E, Radjenovic J (2015) Removal of organic contaminants from secondary effluent by anodic oxidation with a boron-doped diamond anode as tertiary treatment. J Hazard Mater 283:551–557. https://doi.org/10.1016/j.jhazmat.2014.10.003

196

B. N. Malinović et al.

Garcia-Segura S, Ocon JD, Chong MN (2018) Electrochemical oxidation remediation of real wastewater effluents – a review. Process Saf Environ Protect 113:48–67. https://doi.org/10.1016/j. psep.2017.09.014 Gaston L, Lapworth DJ, Stuart M, Arnscheidt J (2019) Prioritization approaches for substances of emerging concern in groundwater: a critical review. Environ Sci Technol 53:6107–6122. https://doi.org/10.1021/acs.est.8b04490 Gatidou G, Oursouzidou M, Stefanatou A, Stasinakis AS (2017) Removal mechanisms of benzotriazoles in duckweed Lemna minor wastewater treatment systems. Sci Total Environ 596-597:12–17. https://doi.org/10.1016/j.scitotenv.2017.04.051 Geissen V, Mol H, Klumpp E, Umlauf G, Nadal M, van der Ploeg M, van de Zee SEATM, Ritsema CJ (2015) Emerging pollutants in the environment: a challenge for water resource management. Int Soil Water Conserv Res 3:57–65. https://doi.org/10.1016/j.iswcr.2015.03.002 Genc A, Goc S (2018) Electroflotation of oily wastewater using stainless steel sponge electrodes. Water Sci Technol 78:1481–1488. https://doi.org/10.2166/wst.2018.422 GilPavas E, Dobrosz-Gómez I, Gómez-García MÁ (2018) Optimization of solar-driven photo-­ electro-­Fenton process for the treatment of textile industrial wastewater. J Water Process Eng 24:49–55. https://doi.org/10.1016/j.jwpe.2018.05.007 Gogate PR, Mededovic-Thagard S, McGuire D, Chapas G, Blackmon J, Cathey R (2014) Hybrid reactor based on combined cavitation and ozonation: from concept to practical reality. Ultrason Sonochem 21:590–598. https://doi.org/10.1016/j.ultsonch.2013.08.016 Gogoi A, Mazumder P, Tyagi VK, Tushara Chaminda GG, An AK, Kumar M (2018) Occurrence and fate of emerging contaminants in water environment: a review. Groundwater Sustain Develop 6:169–180. https://doi.org/10.1016/j.gsd.2017.12.009 Goodridge F, Scott K (1995) Electrochemical process engineering: a guide to the design of electrolytic plant. Springer US, 244 p. https://doi.org/10.1007/978-­1-­4899-­0224-­5 Goulart LA, Alves SA, Mascaro LH (2019) Photoelectrochemical degradation of bisphenol A using Cu doped WO3 electrodes. J Electroanal Chem 839:123–133. https://doi.org/10.1016/j. jelechem.2019.03.027 Groenen Serrano K (2018) Chapter 6: Indirect electrochemical oxidation using hydroxyl radical, active chlorine, and peroxodisulfate. In: Martínez-Huitle CA, Rodrigo MA, Scialdone O (eds) Electrochemical water and wastewater treatment. Butterworth-Heinemann, pp 133–164. https://doi.org/10.1016/B978-­0-­12-­813160-­2.00006-­7 Gupta SS, Islam MR, Pradeep T (2019) Chapter 7: Capacitive deionization (CDI): an alternative cost-efficient desalination technique. In: Ahuja S (ed) Advances in water purification techniques. Elsevier, pp 165–202. https://doi.org/10.1016/B978-­0-­12-­814790-­0.00007-­7 Han Q, Dong W, Wang H, Liu T, Tian Y, Song X (2018) Degradation of tetrabromobisphenol A by ferrate(VI) oxidation: performance, inorganic and organic products, pathway and toxicity control. Chemosphere 198:92–102. https://doi.org/10.1016/j.chemosphere.2018.01.117 Herrero P, Borrull F, Pocurull E, Marcé RM (2014) An overview of analytical methods and occurrence of benzotriazoles, benzothiazoles and benzenesulfonamides in the environment. Trends Anal Chem 62:46–55. https://doi.org/10.1016/j.trac.2014.06.017 Hou M, Chu Y, Li X, Wang H, Yao W, Yu G, Murayama S, Wang Y (2016) Electro-peroxone degradation of diethyl phthalate: cathode selection, operational parameters, and degradation mechanisms. J Hazard Mater 319:61–68. https://doi.org/10.1016/j.jhazmat.2015.12.054 Hu Y, Zhu Q, Yan X, Liao C, Jiang G (2019) Occurrence, fate and risk assessment of BPA and its substituents in wastewater treatment plant: a review. Environ Res 178:108732. https://doi. org/10.1016/j.envres.2019.108732 Huang KY, Chou WL, Wang CT, Chang YC, Shu CM (2015) Electrochemically assisted coagulation for the adsorptive removal of dimethyl phthalate from aqueous solutions using iron hydroxides. J Taiwan Inst Chem E 50:236–241. https://doi.org/10.1016/j.jtice.2014.12.028 Huang KL, Chu YC, Chen HR, Kuo YM (2017) Electrochemical degradation of Bisphenol A in water with/without Ce(IV) addition. Int J Electrochem Sci:12098–12111. https://doi. org/10.20964/2017.12.600

2  Electrochemical Treatments for the Removal of Emerging Contaminants

197

Huntscha S, Hofstetter TB, Schymanski EL, Spahr S, Hollender J (2014) Biotransformation of benzotriazoles: insights from transformation product identification and compound-specific isotope analysis. Environ Sci Technol 48:4435–4443. https://doi.org/10.1021/es405694z Huong Le TX, Drobek M, Bechelany M, Motuzas J, Julbe A, Cretin M (2019) Application of Fe-MFI zeolite catalyst in heterogeneous electro-Fenton process for water pollutants abatement. Microporous Mesoporous Mater 278:64–69. https://doi.org/10.1016/j.micromeso.2018.11.021 Islam SMDU (2019) Electrocoagulation (EC) technology for wastewater treatment and pollutants removal. Sustain Water Resour Manag 5:359–380. https://doi.org/10.1007/s40899-­017-­0152-­1 ISO 15522:1999 Water quality – Determination of the inhibitory effect of water constituents on the growth of activated sludge microorganisms. https://www.iso.org/standard/27554.html Accessed 4 Feb 2020 ISO 13641-1:2003 Water quality – Determination of inhibition of gas production of anaerobic bacteria – Part 1: General test. https://www.iso.org/standard/36232.html Accessed 4 Feb 2020 ISO 8192:2007 Water quality – Test for inhibition of oxygen consumption by activated sludge for carbonaceous and ammonium oxidation. https://www.iso.org/standard/37369.html Accessed 4 Feb 2020 Jia L, Pei X, Yang F (2019) Electrolysis-assisted Mn(II)/sulfite process for organic contaminant degradation at near-neutral pH. Water 11:1608. https://doi.org/10.3390/w11081608 Jo EY, Lee TK, Kim Y, Park CG (2016) Effect of anions on the removal of bisphenol A in wastewater by electro-oxidation process. Desalin Water Treat 57:29500–29508. https://doi.org/10.108 0/19443994.2016.1197156 Jurado A, Walther M, Díaz-Cruz MS (2019) Occurrence, fate and environmental risk assessment of the organic microcontaminants included in the Watch Lists set by EU Decisions 2015/495 and 2018/840  in the groundwater of Spain. Sci Total Environ 663:285–296. https://doi. org/10.1016/j.scitotenv.2019.01.270 Kameda Y, Kimura K, Miyazaki M (2011) Occurrence and profiles of organic sun-blocking agents in surface waters and sediments in Japanese rivers and lakes. Environ Pollut 159:1570–1576. https://doi.org/10.1016/j.envpol.2011.02.055 Khosravanipour Mostafazadeh A, Zolfaghari M, Drogui P (2016) Electrofiltration technique for water and wastewater treatment and bio-products management: a review. J Water Process Eng 14:28–40. https://doi.org/10.1016/j.jwpe.2016.10.003 Kolesnikov VA, Il’in VI, Kolesnikov AV (2019) Electroflotation in wastewater treatment from oil products, dyes, surfactants, ligands, and biological pollutants: a review. Theor Found Chem Eng 53:251–273. https://doi.org/10.1134/S0040579519010093 Kotthoff L, Keller J, Loerchner D, Mekonnen TF, Koch M (2019) Transformation products of organic contaminants and residues: overview of current simulation methods. Molecules 24:753. https://doi.org/10.3390/molecules24040753 Kraševec I, Prosen H (2018) Solid-phase extraction of polar benzotriazoles as environmental pollutants: a review. Molecules 23:2501. https://doi.org/10.3390/molecules23102501 Kyzas GZ, Matis KA (2016) Electroflotation process: a review. J Mol Liq 220:657–664. https:// doi.org/10.1016/j.molliq.2016.04.128 Labiadh L, Ammar S, Kamali AR (2019) Oxidation/mineralization of AO7 by electro-Fenton process using chalcopyrite as the heterogeneous source of iron and copper catalysts with enhanced degradation activity and reusability. J Electroanal Chem 853:113532. https://doi.org/10.1016/j. jelechem.2019.113532 Lauw SJL, Lee JHQ, Tessensohn ME, Leong WQ, Webster RD (2017) The electrochemical reduction of di-(2-ethylhexyl) phthalate (DEHP) in acetonitrile. J Electroanal Chem 794:103–111. https://doi.org/10.1016/j.jelechem.2017.04.012 Lee Y, Kovalova L, McArdell CS, von Gunten U (2014) Prediction of micropollutant elimination during ozonation of a hospital wastewater effluent. Water Res 64:134–148. https://doi. org/10.1016/j.watres.2014.06.027 Lei M, Zhang L, Lei J, Zong L, Li J, Wu Z, Wang Z (2015) Overview of emerging contaminants and associated human health effects. Biomed Res Int 404796. https://doi.org/10.1155/2015/404796

198

B. N. Malinović et al.

Lei C, Liang F, Li J, Chen W, Huang B (2019) Electrochemical reductive dechlorination of chlorinated volatile organic compounds (Cl-VOCs): effects of molecular structure on the dehalogenation reactivity and mechanisms. Chem Eng J 358:1054–1064. https://doi.org/10.1016/j. cej.2018.10.105 Li R, Liu Y, Chen G, Tam NFY, Shin PKS, Cheung SG, Luan T (2008) Physiological responses of the alga Cyclotella caspia to bisphenol A exposure. Bot Mar 51:360–369. https://doi. org/10.1515/BOT.2008.050 Li H, Long Y, Zhu X, Tian Y, Ye J (2017) Influencing factors and chlorinated byproducts in electrochemical oxidation of bisphenol A with boron-doped diamond anodes. Electrochim Acta 246:1121–1130. https://doi.org/10.1016/j.electacta.2017.06.163 Li X, Chen S, Angelidaki I, Zhang Y (2018a) Bio-electro-Fenton processes for wastewater treatment: advances and prospects. Chem Eng J 354:492–506. https://doi.org/10.1016/j. cej.2018.08.052 Li X, Wang L, Asimakopoulos AG, Sun H, Zhao Z, Zhang J, Zhang L, Wang Q (2018b) Benzotriazoles and benzothiazoles in paired maternal urine and amniotic fluid samples from Tianjin, China. Chemosphere 199:524–530. https://doi.org/10.1016/j.chemosphere.2018.02.076 Li J, Wang Y, Zhao S, Jan AK, Zhang X, Zhao X (2019) Electrospun nanostructured Co3O4/BiVO4 composite films for photoelectrochemical applications. J Colloid Interface Sci 539:442–447. https://doi.org/10.1016/j.jcis.2018.12.081 Lin H, Zhang H, Wang X, Wang L, Wu J (2014) Electro-Fenton removal of Orange II in a divided cell: reaction mechanism, degradation pathway and toxicity evolution. Sep Purif Technol 122:533–540. https://doi.org/10.1016/j.seppur.2013.12.010 Liu Y, Guan Y, Gao Q, Tam NFY, Zhu W (2010) Cellular responses, biodegradation and bioaccumulation of endocrine disrupting chemicals in marine diatom Navicula incerta. Chemosphere 80:592–599. https://doi.org/10.1016/j.chemosphere.2010.03.042 Liu Y-S, Ying G-G, Shareef A, Kookana RS (2011) Biodegradation of three selected benzotriazoles under aerobic and anaerobic conditions. Water Res 45:5005–5014. https://doi. org/10.1016/j.watres.2011.07.001 Liu Y-S, Ying G-G, Shareef A, Kookana RS (2012) Occurrence and removal of benzotriazoles and ultraviolet filters in a municipal wastewater treatment plant. Environ Pollut 165:225–232. https://doi.org/10.1016/j.envpol.2011.10.009 Liu X, Zhou Y, Zhang J, Luo L, Yang Y, Huang H, Peng H, Tang L, Mu Y (2018) Insight into electro-­Fenton and photo-Fenton for the degradation of antibiotics: mechanism study and research gaps. Chem Eng J 347:379–397. https://doi.org/10.1016/j.cej.2018.04.142 Lopez de Alda MJ, Dı́az-Cruz S, Petrovic M, Barceló D (2003) Liquid chromatography-(tandem) mass spectrometry of selected emerging pollutants (steroid sex hormones, drugs and alkylphenolic surfactants) in the aquatic environment. J Chromatogr A 1000:503–526. https://doi. org/10.1016/S0021-­9673(03)00509-­0 López-Pacheco IY, Silva-Núñez A, Salinas-Salazar C, Arévalo-Gallegos A, Lizarazo-Holguin LA, Barceló D, Iqbal HMN, Parra-Saldívar R (2019) Anthropogenic contaminants of high ­concern: existence in water resources and their adverse effects. Sci Total Environ 690:1068–1088. https://doi.org/10.1016/j.scitotenv.2019.07.052 Ltaïef AH, Sabatino S, Proietto F, Ammar S, Gadri A, Galia A, Scialdone O (2018) Electrochemical treatment of aqueous solutions of organic pollutants by electro-Fenton with natural heterogeneous catalysts under pressure using Ti/IrO2-Ta2O5 or BDD anodes. Chemosphere 202:111–118. https://doi.org/10.1016/j.chemosphere.2018.03.061 Lu Z, De Silva AO, Peart TE, Cook CJ, Tetreault GR, Servos MR, Muir DCG (2016) Distribution, partitioning and bioaccumulation of substituted diphenylamine antioxidants and benzotriazole UV stabilizers in an urban creek in Canada. Environ Sci Technol 50:9089–9097. https://doi. org/10.1021/acs.est.6b01796 Lü H, Mo CH, Zhao HM, Xiang L, Katsoyiannis A, Li YW, Cai QY, Wong MH (2018) Soil contamination and sources of phthalates and its health risk in China: a review. Environ Res 164:417–429. https://doi.org/10.1016/j.envres.2018.03.013

2  Electrochemical Treatments for the Removal of Emerging Contaminants

199

Luo T, Feng H, Tang L, Lu Y, Tang W, Chen S, Yu J, Xie Q, Ouyang X, Chen Z (2019) Efficient degradation of tetracycline by heterogeneous electro-Fenton process using Cu-doped Fe@ Fe2O3: mechanism and degradation pathway. Chem Eng J 122970. https://doi.org/10.1016/j. cej.2019.122970 Ma J, Ding Y, Chi L, Yang X, Zhong Y, Wang Z, Shi Q (2019) Degradation of benzotriazole by sulfate radical-based advanced oxidation process. Environ Technol:1–10. https://doi.org/10.108 0/09593330.2019.1625959 Malinovic BN, Pavlovic MG (2016) Decolorization of reactive violet 5 dye in textile wastewater by electrocoagulation. J Electrochem Sci Eng 6:67–75. https://doi.org/10.5599/jese.223 Malinovic BN, Pavlovic MG, Halilovic N (2015) Electrochemical removal of nitrate from wastewater using copper cathode. J Environ Prot Ecol 16:1273–1281 Malinovic BN, Pavlovic MG, Djuricic T (2017) Electrocoagulation of textile wastewater containing a mixture of organic dyes by iron electrode. J Electrochem Sci Eng 7:103–110. https://doi. org/10.5599/jese.366 Prosen H, Malinović BN, Korenčič E, Markelj J, Vranješ S, Djuričić T (2020) Electrooxidation of polar benzotriazole - The impact of supporting electrolyte. J Chem Technol Environmental 1:28-33. https://doi.org/10.7251/JCTE2001028P Martínez-Huitle CA, Panizza M (2018) Electrochemical oxidation of organic pollutants for wastewater treatment. Curr Opin Electrochem 11:62–71. https://doi.org/10.1016/j. coelec.2018.07.010 Martínez-Huitle CA, Rodrigo MA, Sirés I, Scialdone O (2015) Single and coupled electrochemical processes and reactors for the abatement of organic water pollutants: a critical review. Chem Rev 115:13362–13407. https://doi.org/10.1021/acs.chemrev.5b00361 Matich EK, Chavez Soria NG, Aga DS, Atilla-Gokcumen GE (2019) Applications of metabolomics in assessing ecological effects of emerging contaminants and pollutants on plants. J Hazard Mater 373:527–535. https://doi.org/10.1016/j.jhazmat.2019.02.084 Medina F, Duran BG, Armijo F, del Rio R (2015) Electrochemical oxidation of diethyl phthalate using boron doped diamond, modified with gold particles. Int J Electrochem Sci 10:8973–8981 Mizukawa A, Molins-Delgado D, de Azevedo JCR, Fernandes CVS, Díaz-Cruz S, Barceló D (2017) Sediments as a sink for UV filters and benzotriazoles: the case study of Upper Iguaçu watershed, Curitiba (Brazil). Environ Sci Pollut Res 24:18284–18294. https://doi.org/10.1007/ s11356-­017-­9472-­9 Molins-Delgado D, Távora J, Silvia Díaz-Cruz M, Barceló D (2017) UV filters and benzotriazoles in urban aquatic ecosystems: the footprint of daily use products. Sci Total Environ 601-602:975–986. https://doi.org/10.1016/j.scitotenv.2017.05.176 Mollah MYA, Schennach R, Parga JR, Cocke DL (2001) Electrocoagulation (EC) – science and applications. J Hazard Mater 84:29–41. https://doi.org/10.1016/S0304-­3894(01)00176-­5 Mollah MYA, Morkovsky P, Gomes JAG, Kesmez M, Parga J, Cocke DL (2004) Fundamentals, present and future perspectives of electrocoagulation. J Hazard Mater 114:199–210. https://doi. org/10.1016/j.jhazmat.2004.08.009 Montagner CC, Vidal C, Acayaba RD (2017) Emerging contaminants in aquatic matrices from Brazil: current scenario and analytical, ecotoxicological and legislational aspects. Quim Nova 40:1094–1110. https://doi.org/10.21577/0100-­4042.20170091 Monteil H, Péchaud Y, Oturan N, Oturan MA (2019) A review on efficiency and cost effectiveness of electro- and bio-electro-Fenton processes: application to the treatment of pharmaceutical pollutants in water. Chem Eng J 376:119577. https://doi.org/10.1016/j.cej.2018.07.179 Montesdeoca-Esponda S, Vega-Morales T, Sosa-Ferrera Z, Santana-Rodríguez JJ (2013) Extraction and determination methodologies for benzotriazole UV stabilizers in personal-care products in environmental and biological samples. Trends Anal Chem 51:23–32. https://doi.org/10.1016/j. trac.2013.05.012 Moreira FC, Boaventura RAR, Brillas E, Vilar VJP (2017) Electrochemical advanced oxidation processes: a review on their application to synthetic and real wastewaters. Appl Catal B Environ 202:217–261. https://doi.org/10.1016/j.apcatb.2016.08.037

200

B. N. Malinović et al.

Moussa DT, El-Naas MH, Nasser M, Al-Marri MJ (2017) A comprehensive review of electrocoagulation for water treatment: potentials and challenges. J Environ Manag 186:24–41. https:// doi.org/10.1016/j.jenvman.2016.10.032 MubarakAli D, Park J, Han T, Srinivasan H, Lee SY, Kim JW (2019) Solution plasma process: an option to degrade bisphenol A in liquid-phase to non-toxic products. J Mol Liq 276:605–610. https://doi.org/10.1016/j.molliq.2018.12.041 Murillo-Sierra JC, Ruiz-Ruiz E, Hinojosa-Reyes L, Guzmán-Mar JL, Machuca-Martínez F, Hernández-Ramírez A (2018) Sulfamethoxazole mineralization by solar photo electro-Fenton process in a pilot plant. Catal Today 313:175–181. https://doi.org/10.1016/j.cattod.2017.11.003 Murrell KA, Dorman FL (2020) Characterization and quantification of methyl-benzotriazoles and chloromethyl-benzotriazoles produced from disinfection processes in wastewater treatment. Sci Total Environ 699:134310. https://doi.org/10.1016/j.scitotenv.2019.134310 Nakajima N, Teramoto T, Kasai F, Sano T, Tamaoki M, Aono M, Kubo A, Kamada H, Azumi Y, Saji H (2007) Glycosylation of bisphenol A by freshwater microalgae. Chemosphere 69:934–941. https://doi.org/10.1016/j.chemosphere.2007.05.088 Nakata H, Shinohara R, Murata S, Watanabe M (2010) Detection of benzotriazole UV stabilizers in the blubber of marine mammals by gas chromatography-high resolution mass spectrometry (GC-HRMS). J Environ Monit 12:2088–2092. https://doi.org/10.1039/C0EM00170H Nidheesh PV, Gandhimathi R (2012) Trends in electro-Fenton process for water and wastewater treatment: an overview. Des 299:1–15. https://doi.org/10.1016/j.desal.2012.05.011 Nidheesh PV, Divyapriya G, Oturan N, Trellu C, Oturan MA (2019) Environmental applications of boron-doped diamond electrodes: 1. Applications in water and wastewater treatment ChemElectroChem 6:2124–2142. https://doi.org/10.1002/celc.201801876 Nilsen E, Smalling KL, Ahrens L, Gros M, Miglioranza KSB, Picó Y, Schoenfuss HL (2019) Critical review: grand challenges in assessing the adverse effects of contaminants of emerging concern on aquatic food webs. Environ Toxicol Chem 38:46–60. https://doi.org/10.1002/ etc.4290 Nishiki Y, Kitaori N, Nakamuro K (2011) Performances of small-sized generator of ozone-­ dissolved water using boron-doped diamond electrodes. Ozone-Sci Eng 33:114–120. https:// doi.org/10.1080/01919512.2011.549411 Noguera-Oviedo K, Aga DS (2016) Lessons learned from more than two decades of research on emerging contaminants in the environment. J Hazard Mater 316:242–251. https://doi. org/10.1016/j.jhazmat.2016.04.058 NORMAN Network of reference laboratories, research centres and related organisations for monitoring of emerging environmental substances. https://www.norman-­network. net/?q=NORMAN%20Network. Accessed 4 Feb 2020 Oh W-D, Veksha A, Chen X, Adnan R, Lim JW, Leong KH, Lim TT (2019) Catalytically active nitrogen-doped porous carbon derived from biowastes for organics removal via peroxymonosulfate activation. Chem Eng J 374:947–957. https://doi.org/10.1016/j.cej.2019.06.001 Oturan N, Oturan MA (2018) Chapter 8: Electro-Fenton process: background, new developments, and applications. In: Martínez-Huitle CA, Rodrigo MA, Scialdone O (eds) Electrochemical water and wastewater treatment. Butterworth-Heinemann, pp 193–221. https://doi.org/10.1016/ B978-­0-­12-­813160-­2.00008-­0 Ouiriemmi I, Karrab A, Oturan N, Pazos M, Rozales E, Gadri A, Sanromán MÁ, Ammar S, Oturan MA (2017) Heterogeneous electro-Fenton using natural pyrite as solid catalyst for oxidative degradation of vanillic acid. J Electroanal Chem 797:69–77. https://doi.org/10.1016/j. jelechem.2017.05.028 Paluselli A, Aminot Y, Galgani F, Net S, Sempéré R (2018) Occurrence of phthalate acid esters (PAEs) in the northwestern Mediterranean Sea and the Rhone River. Prog Oceanogr 163:221–231. https://doi.org/10.1016/j.pocean.2017.06.002 Pan Z, Yu F, Li L, Song C, Yang J, Wang C, Pan Y, Wang T (2019) Electrochemical microfiltration treatment of bisphenol A wastewater using coal-based carbon membrane. Sep Purif Technol 227:115695. https://doi.org/10.1016/j.seppur.2019.115695

2  Electrochemical Treatments for the Removal of Emerging Contaminants

201

Panizza M, Cerisola G (2009) Direct and mediated anodic oxidation of organic pollutants. Chem Rev 109:6541–6569. https://doi.org/10.1021/cr9001319 Park J, Yamashita N, Park C, Shimono T, Takeuchi DM, Tanaka H (2017) Removal characteristics of pharmaceuticals and personal care products: comparison between membrane bioreactor and various biological treatment processes. Chemosphere 179:347–358. https://doi.org/10.1016/j. chemosphere.2017.03.135 Pastushok O, Zhao F, Ramasamy DL, Sillanpää M (2019) Nitrate removal and recovery by capacitive deionization (CDI). Chem Eng J 375:121943. https://doi.org/10.1016/j.cej.2019.121943 Patil AL, Patil PN, Gogate PR (2014) Degradation of imidacloprid containing wastewaters using ultrasound based treatment strategies. Ultrason Sonochem 21:1778–1786. https://doi. org/10.1016/j.ultsonch.2014.02.029 Pervova MG, Kirichenko VE, Saloutin VI (2010) Determination of 1,2,3-benzotriazole in aqueous solutions and air by reaction-gas-liquid chromatography. J Anal Chem 65:276–279. https://doi. org/10.1134/S1061934810030111 Petrie B, Barden R, Kasprzyk-Hordern B (2015) A review on emerging contaminants in wastewaters and the environment: current knowledge, understudied areas and recommendations for future monitoring. Water Res 72:3–27. https://doi.org/10.1016/j.watres.2014.08.053 Philip JM, Aravind UK, Aravindakumar CT (2018) Emerging contaminants in Indian environmental matrices  – a review. Chemosphere 190:307–326. https://doi.org/10.1016/j. chemosphere.2017.09.120 Pillard DA, Cornell JS, DuFresne DL, Hernandez MT (2001) Toxicity of benzotriazole and benzotriazole derivatives to three aquatic species. Water Res 35:557–560. https://doi.org/10.1016/ S0043-­1354(00)00268-­2 Poza-Nogueiras V, Rosales E, Pazos M, Sanromán MÁ (2018) Current advances and trends in electro-Fenton process using heterogeneous catalysts – a review. Chemosphere 201:399–416. https://doi.org/10.1016/j.chemosphere.2018.03.002 Priac A, Morin-Crini N, Druart C, Gavoille S, Bradu C, Lagarrigue C, Torri G, Winterton P, Crini G (2017) Alkylphenol and alkylphenol polyethoxylates in water and wastewater: a review of options for their elimination. Arab J Chem 10:S3749–S3773. https://doi.org/10.1016/j. arabjc.2014.05.011 Prosen H, Malinovic BN, Korencic E, Markelj J, Vranjes S, Djuricic T (2020) Electrooxidation of polar benzotriazole - The impact of supporting electrolyte. J Chem Technol Environment 1:28-33. https://doi.org/10.7251/JCTE2001028P Radha KV, Sirisha K (2018) Chapter 11: Electrochemical oxidation processes. In: Ameta SC, Ameta R (eds) Advanced oxidation processes for waste water treatment. Academic Press, pp 359–373. https://doi.org/10.1016/B978-­0-­12-­810499-­6.00011-­5 Radke EG, Galizia A, Thayer KA, Cooper GS (2019) Phthalate exposure and metabolic effects: a systematic review of the human epidemiological evidence. Environ Int 132:104768. https://doi. org/10.1016/j.envint.2019.04.040 Rasheed T, Bilal M, Nabeel F, Adeel M, Iqbal HMN (2019) Environmentally-related contaminants of high concern: potential sources and analytical modalities for detection, quantification, and treatment. Environ Int 122:52–66. https://doi.org/10.1016/j.envint.2018.11.038 Raut-Jadhav S, Saharan VK, Pinjari D, Sonawane S, Saini D, Pandit A (2013) Synergetic effect of combination of AOP’s (hydrodynamic cavitation and H2O2) on the degradation of neonicotinoid class of insecticide. J Hazard Mater 261:139–147. https://doi.org/10.1016/j. jhazmat.2013.07.012 Ren W, Tang D, Huang M, Sun J, Lv K (2018) Remarkable improved electro-Fenton efficiency by electric-field-induced catalysis of CeO2. J Hazard Mater 350:88–97. https://doi.org/10.1016/j. jhazmat.2018.02.018 Research and Markets, The global bisphenol A market report 2020: analysis 2014–2018 & forecasts 2019–2024. https://www.researchandmarkets.com/research/hl86rz/global_bisphenol?w=5. Accessed 7 Jan 2020

202

B. N. Malinović et al.

Ribeiro AR, Nunes OC, Pereira MFR, Silva AMT (2015) An overview on the advanced oxidation processes applied for the treatment of water pollutants defined in the recently launched directive 2013/39/EU. Environ Int 75:33–51. https://doi.org/10.1016/j.envint.2014.10.027 Richardson SD (2002) Environmental mass spectrometry: emerging contaminants and current issues. Anal Chem 74:2719–2742. https://doi.org/10.1021/ac020211h Richardson SD (2003) Water analysis: emerging contaminants and current issues. Anal Chem 75:2831–2857.D. https://doi.org/10.1021/ac0301301 Richardson SD (2004) Environmental mass spectrometry: emerging contaminants and current issues. Anal Chem 76:3337–3364. https://doi.org/10.1021/ac040060d Richardson SD (2006) Environmental mass spectrometry: emerging contaminants and current issues. Anal Chem 78:4021–4046. https://doi.org/10.1021/ac060682u Richardson SD (2007) Water analysis: emerging contaminants and current issues. Anal Chem 79:4295–4324. https://doi.org/10.1021/ac070719q Richardson SD (2008) Environmental mass spectrometry: emerging contaminants and current issues. Anal Chem 80:4373–4402. https://doi.org/10.1021/ac800660d Richardson SD (2009) Water analysis: emerging contaminants and current issues. Anal Chem 81:4645–4677. https://doi.org/10.1021/ac9008012 Richardson SD (2010) Environmental mass spectrometry: emerging contaminants and current issues. Anal Chem 82:4742–4774. https://doi.org/10.1021/ac101102d Richardson SD (2012) Environmental mass spectrometry: emerging contaminants and current issues. Anal Chem 84:747–778. https://doi.org/10.1021/ac202903d Richardson SD, Kimura SY (2016) Water analysis: emerging contaminants and current issues. Anal Chem 88:546–582. https://doi.org/10.1021/acs.analchem.5b04493 Richardson SD, Kimura SY (2017) Emerging environmental contaminants: challenges facing our next generation and potential engineering solutions. Environ Technol Inno 8:40–56. https://doi. org/10.1016/j.eti.2017.04.002 Richardson SD, Ternes TA (2005) Water analysis: emerging contaminants and current issues. Anal Chem 77:3807–3838. https://doi.org/10.1021/ac058022x Richardson SD, Ternes TA (2011) Water analysis: emerging contaminants and current issues. Anal Chem 83:4614–4648. https://doi.org/10.1021/ac200915r Richardson SD, Ternes TA (2014) Water analysis: emerging contaminants and current issues. Anal Chem 86:2813–2848. https://doi.org/10.1021/ac500508t Richardson SD, Ternes TA (2018) Water analysis: emerging contaminants and current issues. Anal Chem 90:398–428. https://doi.org/10.1021/acs.analchem.7b04577 Rizzo L, Malato S, Antakyali D, Beretsou VG, Đolić MB, Gernjak W, Heath E, Ivancev-Tumbas I, Karaolia P, Lado Ribeiro AR, Mascolo G, McArdell CS, Schaar H, Silva AMT, Fatta-Kassinos D (2019) Consolidated vs new advanced treatment methods for the removal of contaminants of emerging concern from urban wastewater. Sci Total Environ 655:986–1008. https://doi. org/10.1016/j.scitotenv.2018.11.265 Rodriguez-Narvaez OM, Peralta-Hernandez JM, Goonetilleke A, Bandala ER (2017) Treatment technologies for emerging contaminants in water: a review. Chem Eng J 323:361–380. https:// doi.org/10.1016/j.cej.2017.04.106 Rondinini S, Locatelli C, Minguzzi A, Vertova A (2018) Chapter 1: Electroreduction. In: Martínez-­ Huitle CA, Rodrigo MA, Scialdone O (eds) Electrochemical water and wastewater treatment. Butterworth-Heinemann, pp 3–28. https://doi.org/10.1016/B978-­0-­12-­813160-­2.00001-­8 Rudel Ruthann A, Gray JM, Engel CL, Rawsthorne TW, Dodson RE, Ackerman JM, Rizzo J, Nudelman JL, Brody JG (2011) Food packaging and bisphenol A and bis(2-ethyhexyl) phthalate exposure: findings from a dietary intervention. Environ Health Perspect 119:914–920. https://doi.org/10.1289/ehp.1003170 Sabri NA, Schmitt H, Van der Zaan B, Gerritsen HW, Zuidema T, Rijnaarts HHM, Langenhoff AAM (2018) Prevalence of antibiotics and antibiotic resistance genes in a wastewater effluent-­ receiving river in the Netherlands. J Environ Chem Eng 102245. https://doi.org/10.1016/j. jece.2018.03.004

2  Electrochemical Treatments for the Removal of Emerging Contaminants

203

Safwat SM (2019) Coupling microbial fuel cells with electrocoagulation cells to form an integrated system for wastewater treatment. Pol J Environ Stud 28:1909–1915. https://doi.org/10.15244/ pjoes/89543 Safwat SM, Hamed A, Rozaik E (2019) Electrocoagulation/electroflotation of real printing wastewater using copper electrodes: a comparative study with aluminum electrodes. Sep Sci Technol 54:183–194. https://doi.org/10.1080/01496395.2018.1494744 Salgueiro-González N, López de Alda MJ, Muniategui-Lorenzo S, Prada-Rodríguez D, Barceló D (2015) Analysis and occurrence of endocrine-disrupting chemicals in airborne particles. Trends Anal Chem 66:45–52. https://doi.org/10.1016/j.trac.2014.11.006 Salimi M, Esrafili A, Gholami M, Jonidi Jafari A, Rezaei Kalantary R, Farzadkia M, Kermani M, Sobhi HR (2017) Contaminants of emerging concern: a review of new approach in AOP technologies. Environ Monit Assess 189:414. https://doi.org/10.1007/s10661-­017-­6097-­x Santangeli S, Maradonna F, Zanardini M, Notarstefano V, Gioacchini G, Forner-Piquer I, Habibi H, Carnevali O (2017) Effects of diisononyl phthalate on Danio rerio reproduction. Environ Pollut 231:1051–1062. https://doi.org/10.1016/j.envpol.2017.08.060 Santiago Santos G de O, de Salles Pupo MM, Vasconcelos VM, Barrios Eguiluz KI, Salazar Banda GR (2018) Chapter 4: Electroflotation. In: Martínez-Huitle CA, Rodrigo MA, Scialdone O (eds) Electrochemical water and wastewater treatment. Butterworth-Heinemann, pp 77–118. https://doi.org/10.1016/B978-­0-­12-­813160-­2.00004-­3 Sauvé S, Desrosiers M (2014) A review of what is an emerging contaminant. Chem Cent J 8:15. https://doi.org/10.1186/1752-­153X-­8-­15 Senoussi H, Bouhidel K-E (2018) Feasibility and optimisation of a batch mode capacitive deionization (BM CDI) process for textile cationic dyes (TCD) removal and recovery from industrial wastewaters. J Clean Prod 205:721–727. https://doi.org/10.1016/j.jclepro.2018.09.026 Sharma A, Ahmad J, Flora SJS (2018) Application of advanced oxidation processes and toxicity assessment of transformation products. Environ Res 167:223–233. https://doi.org/10.1016/j. envres.2018.07.010 Shestakova M, Vinatoru M, Mason TJ, Iakovleva E, Sillanpää M (2016) Sonoelectrochemical degradation of formic acid using Ti/Ta2O5-SnO2 electrodes. J Mol Liq 223:388–394. https://doi. org/10.1016/j.molliq.2016.08.054 Shi ZQ, Liu YS, Xiong Q, Cai WW, Ying GG (2019) Occurrence, toxicity and transformation of six typical benzotriazoles in the environment: a review. Sci Total Environ 661:407–421. https:// doi.org/10.1016/j.scitotenv.2019.01.138 Simon RG, Stöckl M, Becker D, Steinkamp A-D, Abt C, Jungfer C, Weidlich C, Track T, Mangold K-M (2018) Current to clean water – electrochemical solutions for groundwater, water, and wastewater treatment. Chemie Ingenieur Technik 90:1832–1854. https://doi.org/10.1002/ cite.201800081 Sirés I, Brillas E, Oturan MA, Rodrigo MA, Panizza M (2014) Electrochemical advanced oxidation processes: today and tomorrow. A review. Environ Sci Pollut Res 21:8336–8367. https:// doi.org/10.1007/s11356-­014-­2783-­1 Sorensen JPR, Lapworth DJ, Nkhuwa DCW, Stuart ME, Gooddy DC, Bell RA, Chirwa M, Kabika J, Liemisa M, Chibesa M, Pedley S (2015) Emerging contaminants in urban groundwater sources in Africa. Water Res 72:51–63. https://doi.org/10.1016/j.watres.2014.08.002 Sousa JCG, Ribeiro AR, Barbosa MO, Pereira MFR, Silva AMT (2018) A review on environmental monitoring of water organic pollutants identified by EU guidelines. J Hazard Mater 344:146–162. https://doi.org/10.1016/j.jhazmat.2017.09.058 Souza FL, Sáez C, Cañizares P, Motheo AJ, Rodrigo MA (2013) Sonoelectrolysis of wastewaters polluted with dimethyl phthalate. Ind Eng Chem Res 52:9674–9682. https://doi.org/10.1021/ ie400552z Souza FL, Sáez C, Cañizares P, Motheo AJ, Rodrigo MA (2014a) Electrochemical removal of dimethyl phthalate with diamond anodes. J Chem Technol Biot 89:282–289. https://doi. org/10.1002/jctb.4118

204

B. N. Malinović et al.

Souza FL, Sáez C, Cañizares P, Motheo AJ, Rodrigo MA (2014b) Using a new photo-reactor to promote conductive-diamond electrochemical oxidation of dimethyl phthalate. J Chem Technol Biol 89:1251–1258. https://doi.org/10.1002/jctb.4340 Souza FL, Aquino JM, Irikura K, Miwa DW, Rodrigo MA, Motheo AJ (2014c) Electrochemical degradation of the dimethyl phthalate ester on a fluoride-doped Ti/β-PbO2 anode. Chemosphere 109:187–194. https://doi.org/10.1016/j.chemosphere.2014.02.018 Souza FL, Aquino JM, Miwa DW, Rodrigo MA, Motheo AJ (2014d) Electrochemical Degradation of dimethyl phthalate ester on a DSA electrode. J Braz Chem 25:492–501. https://doi. org/10.5935/0103-­5053.20140007 Souza FL, Aquino JM, Miwa DW, Rodrigo MA, Motheo AJ (2014e) Photo-assisted electrochemical degradation of the dimethyl phthalate ester on DSA® electrode. J Environ Chem Eng 2:811–818. https://doi.org/10.1016/j.jece.2014.02.003 Speltini A, Sturini M, Maraschi F, Porta A, Profumo A (2016) Fast low-pressurized microwave-­ assisted extraction of benzotriazole, benzothiazole and benezenesulfonamide compounds from soil samples. Talanta 147:322–327. https://doi.org/10.1016/j.talanta.2015.09.074 Su CC, Cada CA, Dalida MLP, Lu MC (2013) Effect of UV light on acetaminophen degradation in the electro-Fenton process. Sep Purif Technol 120:43–51. https://doi.org/10.1016/j. seppur.2013.09.034 Sun C, Zhang G, Zheng H, Liu N, Shi M, Luo X, Chen L, Li F, Hu S (2019a) Fate of four phthalate esters with presence of Karenia brevis: uptake and biodegradation. Aquat Toxicol 206:81–90. https://doi.org/10.1016/j.aquatox.2018.11.010 Sun M, Fang LM, Liu JQ, Zhang F, Zhai LF (2019b) Electro-activation of O2 on MnO2/graphite felt for efficient oxidation of water contaminants under room condition. Chemosphere 234:269–276. https://doi.org/10.1016/j.chemosphere.2019.06.083 Taheran M, Naghdi M, Brar SK, Verma M, Surampalli RY (2018) Emerging contaminants: Here today, there tomorrow! Environ Nano Monit Manage 10:122–126. https://doi.org/10.1016/j. enmm.2018.05.010 Tang K, Spiliotopoulou A, Chhetri RK, Ooi GTH, Kaarsholm KMS, Sundmark K, Florian B, Kragelund C, Bester K, Andersen HR (2019) Removal of pharmaceuticals, toxicity and natural fluorescence through the ozonation of biologically-treated hospital wastewater, with further polishing via a suspended biofilm. Chem Eng J 359:321–330. https://doi.org/10.1016/j. cej.2018.11.112 The European Commission (2005) Directive 2005/84/EC. https://eur-­lexeuropaeu/eli/dir/2005/84/ oj. Accessed 4 Feb 2020 Thiam A, Brillas E, Centellas F, Cabot PL, Sirés I (2015) Electrochemical reactivity of Ponceau 4R (food additive E124) in different electrolytes and batch cells. Electrochim Acta 173:523–533. https://doi.org/10.1016/j.electacta.2015.05.085 Tiwari B, Sellamuthu B, Ouarda Y, Drogui P, Tyagi RD, Buelna G (2017) Review on fate and mechanism of removal of pharmaceutical pollutants from wastewater using biological approach. Bioresour Technol 224:1–12. https://doi.org/10.1016/j.biortech.2016.11.042 Tran NH, Reinhard M, Gin KY-H (2018) Occurrence and fate of emerging contaminants in municipal wastewater treatment plants from different geographical regions-a review. Water Res 133:182–207. https://doi.org/10.1016/j.watres.2017.12.029 Turkay O, Ersoy ZG, Barışçı S (2017) Review-the application of an electro-peroxone process in water and wastewater treatment. J Electrochem Soc 164:E94–E102. https://doi. org/10.1149/2.0321706jes UNESCO Emerging pollutants in water and wastewater. https://en.unesco.org/emergingpollutants. Accessed 7 Jan 2020 Ur Rehman SW, Wang H, Yao W, Deantes-Espinosa VM, Wang B, Huang J, Deng S, Yu G, Wang Y (2019) Ozonation of the algaecide irgarol: kinetics, transformation products, and toxicity. Chemosphere 236:124374. https://doi.org/10.1016/j.chemosphere.2019.124374 US EPA (2012) Phthalates action plan. https://wwwepagov/sites/production/files/2015-­09/documents/phthalates_actionplan_revised_2012-­03-­14pdf. Accessed 4 Feb 2020

2  Electrochemical Treatments for the Removal of Emerging Contaminants

205

Vodyanitskii YN, Yakovlev AS (2016) Contamination of soils and groundwater with new organic micropollutants: a review. Eurasian Soil Sc 49:560–569. https://doi.org/10.1134/ S1064229316050148 Wang L, Asimakopoulos AG, Moon H-B, Nakata H, Kannan K (2013) Benzotriazole, benzothiazole, and benzophenone compounds in indoor dust from the United States and east asian countries. Environ Sci Technol 47:4752–4759. https://doi.org/10.1021/es305000d Wang L, Asimakopoulos AG, Kannan K (2015) Accumulation of 19 environmental phenolic and xenobiotic heterocyclic aromatic compounds in human adipose tissue. Environ Int 78:45–50. https://doi.org/10.1016/j.envint.2015.02.015 Wang L, Zhang J, Sun H, Zhou Q (2016) Widespread occurrence of benzotriazoles and benzothiazoles in tap water: influencing factors and contribution to human exposure. Environ Sci Technol 50:2709–2717. https://doi.org/10.1021/acs.est.5b06093 WHO (2013) Di(2-ethylhexyl)phthalate in drinking water. https://wwwwhoint/water_sanitation_ health/dwq/chemicals/di2ethylhexyphthalatepdf Accessed 4 Feb 2020 Wilkinson J, Hooda P, Barker J, Barton S, Swinden J (2017) Occurrence, fate and transformation of emerging contaminants in water: an overarching review of the field. Environ Pollut 231:954–970. https://doi.org/10.1016/j.envpol.2017.08.032 Wu J, Pu W, Yang C, Zhang M, Zhang J (2013) Removal of benzotriazole by heterogeneous photoelectro-Fenton like process using ZnFe2O4 nanoparticles as catalyst. J Environ Sci 25:801–807. https://doi.org/10.1016/S1001-­0742(12)60117-­X Wu D, Chen GQ, Hu B, Deng H (2019a) Feasibility and energy consumption analysis of phenol removal from salty wastewater by electro-electrodialysis. Sep Purif Technol 215:44–50. https:// doi.org/10.1016/j.seppur.2019.01.001 Wu P, Zhang Y, Chen Z, Duan Y, Lai Y, Fang Q, Wang F, Li S (2019b) Performance of boron-­ doped graphene aerogel modified gas diffusion electrode for in-situ metal-free electrochemical advanced oxidation of bisphenol A. Appl Catal B Environ 255:117784. https://doi. org/10.1016/j.apcatb.2019.117784 Wu J, Xiong Q, Liang J, He Q, Yang D, Deng R, Chen Y (2020) Degradation of benzotriazole by DBD plasma and peroxymonosulfate: mechanism, degradation pathway and potential toxicity. Chem Eng J 384:123300. https://doi.org/10.1016/j.cej.2019.123300 Xiao K, Liang H, Chen S, Yang B, Zhang J, Li J (2019) Enhanced photoelectrocatalytic degradation of bisphenol A and simultaneous production of hydrogen peroxide in saline wastewater treatment. Chemosphere 222:141–148. https://doi.org/10.1016/j.chemosphere.2019.01.109 Xiao C, Wang L, Zhou Q, Huang X (2020) Hazards of bisphenol A (BPA) exposure: a systematic review of plant toxicology studies. J Hazard Mater 384:121488. https://doi.org/10.1016/j. jhazmat.2019.121488 Xing X, Zhu X, Li H, Jiang Y, Ni J (2012) Electrochemical oxidation of nitrogen-­heterocyclic compounds at boron-doped diamond electrode. Chemosphere 86:368–375. https://doi. org/10.1016/j.chemosphere.2011.10.020 Xing Y, Meng X, Wang L, Zhang J, Wu Z, Gong X, Wang C, Sun H (2018) Effects of benzotriazole on copper accumulation and toxicity in earthworm (Eisenia fetida). J Hazard Mater 351:330–336. https://doi.org/10.1016/j.jhazmat.2018.03.019 Xu A, Brillas E, Han W, Wang L, Sirés I (2019) On the positive effect of UVC light during the removal of benzothiazoles by photoelectro-Fenton with UVA light. Appl Catal B Environ 259:118127. https://doi.org/10.1016/j.apcatb.2019.118127 Xue J, Wan Y, Kannan K (2017) Occurrence of benzotriazoles (BTRs) in indoor air from Albany, New York, USA, and its implications for inhalation exposure. Toxicol Environ Chem 99:402–414. https://doi.org/10.1080/02772248.2016.1196208 Yan S, Shi Y, Tao Y, Zhang H (2019) Enhanced persulfate-mediated photocatalytic oxidation of bisphenol A using bioelectricity and a g-C3N4/Fe2O3 heterojunction. Chem Eng J 359:933–943. https://doi.org/10.1016/j.cej.2018.11.093

206

B. N. Malinović et al.

Yang GCC, Yen CH, Wang CL (2014a) Monitoring and removal of residual phthalate esters and pharmaceuticals in the drinking water of Kaohsiung City, Taiwan. J Hazard Mater 277:53–61. https://doi.org/10.1016/j.jhazmat.2014.03.005 Yang HG, Chun HY, Pak D (2014b) Improvement of sludge anaerobic degradability by combined electro-flotation and electro-oxidation treatment. Biochem Eng J 90:44–48. https://doi. org/10.1016/j.bej.2014.05.010 Yang GCC, Chen YC, Yang HX, Yen CH (2016) Performance and mechanisms for the removal of phthalates and pharmaceuticals from aqueous solution by graphene-containing ceramic composite tubular membrane coupled with the simultaneous electrocoagulation and electrofiltration process. Chemosphere 155:274–282. https://doi.org/10.1016/j.chemosphere.2016.04.060 Ye J, Zhou P, Chen Y, Ou H, Liu J, Li C, Li Q (2018) Degradation of 1H-benzotriazole using ultraviolet activating persulfate: mechanisms, products and toxicological analysis. Chem Eng J 334:1493–1501. https://doi.org/10.1016/j.cej.2017.11.101 Zedda M, Zwiener C (2012) Is nontarget screening of emerging contaminants by LC-HRMS successful? A plea for compound libraries and computer tools. Anal Bioanal Chem 403:2493–2502. https://doi.org/10.1007/s00216-­012-­5893-­y Zeng F, Sherry JP, Bols NC (2016) Use of the rainbow trout cell lines, RTgill-W1 and RTL-W1 to evaluate the toxic potential of benzotriazoles. Ecotoxicol Environ Saf 124:315–323. https://doi. org/10.1016/j.ecoenv.2015.11.003 Zhang Y, Zhang D, Zhou L, Zhao Y, Chen J, Chen Z, Wang F (2018) Polypyrrole/reduced graphene oxide aerogel particle electrodes for high-efficiency electro-catalytic synergistic removal of Cr(VI) and bisphenol A. Chem Eng J 336:690–700. https://doi.org/10.1016/j.cej.2017.11.109 Zhang J, Tang L, Tang W, Zhong Y, Luo K, Duan M, Xing W, Liang J (2019) Removal and recovery of phosphorus from low-strength wastewaters by flow-electrode capacitive deionization. Sep Purif Technol 116322. https://doi.org/10.1016/j.seppur.2019.116322 Zhao H, Qian L, Chen Y, Wang Q, Zhao G (2018) Selective catalytic two-electron O2 reduction for onsite efficient oxidation reaction in heterogeneous electro-Fenton process. Chem Eng J 332:486–498. https://doi.org/10.1016/j.cej.2017.09.093 Zhu G, Yu X, Xie F, Feng W (2019) Ultraviolet light assisted heterogeneous Fenton degradation of tetracycline based on polyhedral Fe3O4 nanoparticles with exposed high-energy {110} facets. Appl Surf Sci 485:496–505. https://doi.org/10.1016/j.apsusc.2019.04.239 Zolfaghari M, Jardak K, Drogui P, Brar SK, Buelna G, Dubé R (2016) Landfill leachate treatment by sequential membrane bioreactor and electro-oxidation processes. J Environ Manag 184:318–326. https://doi.org/10.1016/j.jenvman.2016.10.010 Zupanc M, Kosjek T, Petkovšek M, Dular M, Kompare B, Širok B, Blažeka Ž, Heath E (2013) Removal of pharmaceuticals from wastewater by biological processes, hydrodynamic ­cavitation and UV treatment. Ultrason Sonochem 20:1104–1112. https://doi.org/10.1016/j. ultsonch.2012.12.003

Chapter 3

Technologies to Remove Selenium from Water and Wastewater Eric Lichtfouse , Nadia Morin-Crini , Corina Bradu, Youssef-­ Amine Boussouga, Mehran Aliaskari, Andrea Iris Schäfer, Soumya Das, Lee D. Wilson, Michihiko Ike, Daisuke Inoue, Masashi Kuroda, Sébastien Déon , Patrick Fievet , and Grégorio Crini

Abstract  After major pollution by nitrates and pesticides, soils and groundwater in some parts of the world are now facing the emergence of a third major issue of selenium (Se) contamination. Selenium occurrence in ecosystems results naturally from weathering of Se-containing rocks, and is further aggravated by human activities. Selenium is ubiquitous in the environment, and the two main sources of human exposure by Se are food and water. Se, a metalloid, is an important micronutrient due to Se antioxidant, anti-inflammatory and chemo-preventative properties. At normal dietary doses, selenium is an essential diet element that has nutritional properE. Lichtfouse (*) Aix Marseille University, CNRS, IRD, INRA, Coll France, CEREGE, Aix-en-Provence, France N. Morin-Crini · G. Crini (*) Laboratoire Chrono-environnement, UMR 6249, UFR Sciences et Techniques, Université Bourgogne Franche-Comté, Besançon, France e-mail: [email protected]; [email protected] C. Bradu Department of Systems Ecology and Sustainability, PROTMED Research Centre, University of Bucharest, Bucharest, Romania e-mail: [email protected] Y.-A. Boussouga · M. Aliaskari · A. I. Schäfer Institute for Advanced Membrane Technologies (IAMT), Karlsruhe Institute of Technology (KIT), Karlsruhe, Eggenstein-Leopoldshafen, Germany e-mail: [email protected]; [email protected]; andrea.iris.schaefer@ kit.edu S. Das Department of Geological Sciences, University of Saskatchewan, Saskatoon, SK, Canada e-mail: [email protected] L. D. Wilson Department of Chemistry, University of Saskatchewan, Saskatoon, SK, Canada e-mail: [email protected] © The Author(s), under exclusive license to Springer Nature Switzerland AG 2021 N. Morin-Crini et al. (eds.), Emerging Contaminants Vol. 2, Environmental Chemistry for a Sustainable World 66, https://doi.org/10.1007/978-3-030-69090-8_3

207

208

E. Lichtfouse et al.

ties and is necessary to maintain good health in humans and animals. Nonetheless, exposure to high concentrations of selenium is harmful to living beings. In terms of contamination, selenium as an emerging hazardous substance is receiving particular attention in developing countries, where research is focussing on water treatment. Actual remediation techniques are limited because removing Se from complex mixtures of substances is very challenging. Yet, techniques of water decontamination are developing rapidly. Here, we review selenium occurrence, pollution, properties and remediation. Advanced remediation include technologies based on zero-valent iron, iron-oxy-hydroxides, supported materials, nanofiltration, reverse osmosis, enhanced ultrafiltration, electrodialysis, and activated granular sludge. Keywords  Selenium · Water · Wastewater · Removal technologies · Coprecipitation · Reduction techniques · Zero-valent iron · Metal oxides · Activated alumina · Activated carbon · Ion-exchange · Supported materials · Membrane filtration · Electrodialysis · Activated granular sludge

3.1  Introduction Sola dosis facit venenum – All things are poison, and nothing is without poison; only the dose determines what is not a poison. Paracelsus (1493–1541)

Metals and metalloids, and their organometallic derivatives are among the substances that present a recognized danger to humans and the environment. Metals and metalloids are part of our daily lives because they are used in a wide variety of products and applications. Due to their particular physical, chemical and biological properties, they have many industrial applications, e.g. in anti-corrosion coatings, stainless steel composition, electrical industry, pigment manufacturing, fruit fungicide manufacturing, construction, catalytic converters, alloy and semiconductor manufacturing, wood processing, glass manufacturing, catalysis, and cosmetology. In the field of nutrition, certain metals and metalloids are also useful or even essential at trace levels for our health, contributing to the proper functioning of living organisms (Amiard 2011; Crini 2017; Hejna et  al. 2018). However, these M. Ike · D. Inoue Division of Sustainable Energy and Environmental Engineering, Graduate School of Engineering, Osaka University, Suita, Osaka, Japan e-mail: [email protected]; [email protected] M. Kuroda Graduate School of Environmental and Disaster Research, Tokoha University, Suruga-ku, Shizuoka, Japan e-mail: [email protected] S. Déon · P. Fievet Institut UTINAM, UMR CNRS 6213, UFR Sciences et Techniques, Université Bourgogne Franche-Comté, Besançon, France e-mail: [email protected]; [email protected]

3  Technologies to Remove Selenium from Water and Wastewater

209

substances can also have negative impacts at individuals, population and ecosystem levels, resulting from their transport, transfer, bioaccumulation and biomagnification along the food chain. For example, metals such as iron, cobalt and zinc, are either essential for human health because they have a physiological role, or toxic in larger amounts or certain forms. Other substances have no biological function and can become toxic if they are absorbed in excess. Other substances such as cadmium, mercury, lead and arsenic are poisonous in their ionic form. This is the paradox of these contaminants, they are both useful and potentially dangerous. For these reasons, many metals and their derivatives are among the most monitored hazardous substances. Toxicity depends not only on concentration but also on speciation. For instance, many elements such as arsenic, chromium, nickel, manganese and vanadium, exist under several different oxidation states. Other elements such as mercury, tin, manganese and arsenic occur under different chemical forms. From a chemical and toxicological point of view, these substances are now well-known (Amiard 2011). The intensive use of metals and metalloids by humans began with the Industrial Revolution more than 250 years ago (Crini and Lichtfouse 2018). From the beginning, this use began to affect the environment, and therefore our ecosystems. Their presence in different environmental compartments, e.g. water, soil and air, is mainly due to contamination generated by human activities, e.g. soil degradation. Yet there are also natural sources such as earth’s crust elements, volcanic activities and natural biological activity. Anthropogenic sources of Se are numerous and varied, for instance: agricultural activities, e.g. fertilizing; industrial activities, e.g. mining and metallurgy; industrial discharges in water, soil and air; combustion by incineration for energy production, and transport activities (Rosenfeld and Beath 1964). In Europe, in order to improve the quality of the environment and guarantee the health of populations, regulations have been gradually set up since the 1970s, particularly in the field of industrial waste (Crini and Badot 2007, 2010; Fordyce 2013; Crini 2017; Hejna et al. 2018). The objective was to reduce or eliminate emissions of targeted substances according to criteria of toxicity, persistence and bioaccumulation. Metals are now one of the most closely controlled classes of chemicals, classified as hazardous or priority hazardous substances. In Europe, currently, out of forty or so metals in the periodic table, the following twelve metals are subject to special monitoring, particularly for discharges and industrial wastes: aluminum, arsenic, cadmium, chromium, chromium, nickel, copper, tin, iron, manganese, mercury, lead and zinc. Considerable environmental efforts have been made by the industry over the past 30 years, yet metals and metalloids continue to be the subject of constant debate and concern. Selenium is a subject of particular attention. Se is known as the “double-edged sword element” for its dual beneficial and toxic character to health (Fernández-­ Martínez and Charlet 2009). Indeed, despite Se nutritional benefits, it is one of the most toxic natural elements. Selenium is ubiquitous in the environment and the two main sources of human exposure are food and water. This metalloid is an important micronutrient for living beings and an essential trace element of fundamental importance to health due to its nutritional and biological properties. At normal dietary doses, selenium is necessary to maintain good health in humans and animals.

210

E. Lichtfouse et al.

Indeed, Se deficiency is a major problem worldwide, with several cases of deficiency reported. However, exposure to high concentrations of selenium is harmful to living beings. Overexposure and selenium deficiency have been associated with adverse health effects. In 1999, Peter M. Chapman, a world-renowned environmental toxicologist, asked the following question: is selenium a potential time bomb or just another contaminant? This issue highlights “the valuable but risky nature of this chemical element” (Chapman 1999). Over the past two decades, selenium has become a new substance of concern, not only for nutrition and medicine but also for water pollution (Hatfield 2001; Wu 2004; Fernández-Martínez and Charlet 2009; Santos et al. 2015; Vinceti et al. 2017, 2018a, b; Hejna et al. 2018; Ullah et al. 2018, 2019). As a consequence, research on selenium is developing in environmental science to study its presence, behavior, transfer, and bioaccumulation; in toxicology to assess toxicity and impact; and in water engineering to clean waters. Selenium is naturally present and widespread in the environment. It is released by natural processes, including geological and geothermal activities, and human activities, e.g. mining, industry, and agriculture. Various parts of the world, especially in North America and in Europe, are experiencing issues of selenium contamination, especially in soils, aquifer sediments and groundwater (Wu 2004; Conley et al. 2009; DeForest et al. 2012; Fordyce 2013; Santos et al. 2015; Crini 2017; Di Marzio et al. 2019; Paul and Saha 2019). Living organisms can be exposed to selenium through Se presence in food and drinking water. Exposure to selenium also occurs when living beings come into contact with soil or air that contains selenium. The two main sources of selenium exposure are food, e.g. bread, cereals, nuts, fish, eggs and milk, and tap water (Rosenfeld and Beath 1964; Combs and Combs 1986; Combs 1988; Mayland 1994; Hatfield 2001; Reilly 2002; Fernández-Martínez and Charlet 2009; ANSES 2012; Bañuelos et al. 2014; Health Canada 2014; Santos et al. 2015; Crini 2017; Kieliszek 2019). After pollution by nitrates and pesticides in some parts of the world like France face the issue of selenium pollution. Unlike most pesticides, selenium is mainly of  natural origin (INERIS 2011; INRS 2011; OMS 2011; ANSES 2012; Health Canada 2014; Paul and Saha 2019). There are also anthropogenic sources of selenium in many industries including mining, petroleum refining and metallurgical activities. Other human activities such as the release of sedimentary construction waste rocks, irrigation and fossil fuel combustion also contribute to selenium contamination (Crini 2017). For drinking water, the World Health Organization has set a regulation limit of 40 μg Se/L, most European countries have set the regulation limit at 10  μg/L, and the upper limit set by the United States Environmental Protection Agency is 50 μg/L. However, the value of 10 μg/L is often exceeded in groundwater (Fordyce 2013; Santos et al. 2015; Crini 2017). Today, we have reached the Pollutant Removal Age  of the anthropocene, and there has been considerable efforts to develop technologies to reduce contaminant emissions (Crini and Lichtfouse 2018). The removal of selenium from water and wastewaters is of great interest in the field of water pollution. However, the problem is not simple because it is difficult to remove trace selenium from complex mixtures

3  Technologies to Remove Selenium from Water and Wastewater

211

of substances. In its natural form as an element, selenium cannot be destroyed but it does have the ability to change of form. From a chemical point of view, due to their structure and stability, selenium forms are difficult to treat and often compete with other substances (BRGM 2011; Crini 2017). The removal of selenium from water and wastewater is controlled by Se speciation and the chemical composition of the water from the supply source. Treatment is also costly due to the characteristics of the aqueous solution and due to the strict discharge limits of selenium and its oxyanions such as Se(IV) and Se(VI) (Koren et  al. 1992; Kapoor et  al. 1995; Crini 2017; Stefaniak et al. 2018; Rene et al. 2019). Speciation of selenium in groundwater or in a raw effluents plays an essential role in the effectiveness of treatment methods used for Se removal, especially because selenium is often present at low concentrations, less than 1 mg/L (Fernández-Martínez and Charlet 2009; BRGM 2011; Santos et al. 2015). Most of the current research focuses on Se(IV) and Se(VI) removal, and an interesting challenge is the removal of the organic form of selenium. Like other many metalloids, selenium is difficult to remove, especially the oxyanion of Se(VI) present, for example, in mining effluents (Rene et al. 2019). Two types of contaminated waters should be distinguished: waters to be treated for drinking, which usually contain less than 0.1 mg Se/L, and industrial waters that contain more than 1  mg/L.  In both cases effluents containing selenium are often associated with other substances and high salinity (Koren et al. 1992; Kapoor et al. 1995). As a consequence, the choice and effectiveness of a treatment process is influenced not only by the oxidation state of selenium, its concentration and the presence of other contaminants, but also by several other factors such as pre-­existing treatment facilities and processes, treatment objectives, as well as waste treatment concerns, and costs. Technologies available for the removal of selenium can be classified in chemical methods, e.g. coprecipitation, reduction-adsorption, oxidation-reduction, electrocoagulation, catalyzed cementation; physical treatments, e.g. adsorption, membrane filtration, ion-exchange; and biological methods such as microbial reduction, bacterial treatment, enzymatic reduction, fluidized bed reactor, algal assimilation and constructed wetlands (Koren et al. 1992; Kapoor et al. 1995; Sandy and DiSante 2010; Santos et al. 2015; Crini 2017; Rene et al. 2019). For drinking water, local disposal solutions such as sand filtration coupled with ion exchange resins and membrane treatments, e.g. microfiltration and nanofiltration, show removal performances above 95% (Crini 2017). However, such solutions are often poorly adapted, poorly selective and costly. Innovations are therefore needed to find treatment methods that are efficient, inexpensive, technologically feasible and environmentally friendly. Generally, in France, the solution consists either in asking the competent authorities for operating exemptions or, more often, in seeking another water resource. For industrial-scale selenium removal, the first possible method is iron coprecipitation and adsorption, coupled, if necessary, with a coagulation-flocculation. Other treatments include reduction techniques, adsorption, e.g. using metal oxides, activated alumina or activated carbon; ion exchange; reverse osmosis and nanofiltration; showing 75–99% removal efficiencies depending on selenium form. For

212

E. Lichtfouse et al.

drinking water production, lime softening (decarbonation) and reverse electrodialysis show lower removal, below 70%. For the treatment of industrial effluents, a combination of physicochemical processes such as chemical reduction, coprecipitation, coagulation, adsorption, filtration is generally used (Rene et  al. 2019). Biological techniques such as microbial reduction, aerobic wetlands and biochemical reactors can also be used. Adsorption on non-conventional materials, e.g. biosorbents, and innovative biological techniques such as microalgal-bacterial treatment, bioremediation, and phytoremediation are being explored (Crini 2017; He et  al. 2018b). Among innovative treatments to removing selenium, one of the most promising approaches is fixed-bed biological treatment in terms of efficiency and cost. The next section present selenium chemistry, occurrence and decontamination methods from water and wastewater.

3.2  Selenium Chemistry and Applications For more general information on selenium, the reader can consult a very interesting technical document, published in March 2014 by Health Canada of the Canadian Federal Department (Health Canada 2014; www.santecanada.gc.ca). Other comprehensive reports are given in OEHHA (2010), BRGM (2011) and ANSES (2012). Book references include Rosenfeld and Beath (1964), Combs and Combs (1986), Frankenberger and Benson (1994), Frankenberger and Engberg (1998), Hatfield (2001), Lemly (2002), Surai and Taylor-Pickard (2008), Reilly (2006), Woollins and Laitinen (2011), Preedy (2015), Crini et al. (2017), and van Hullebusch (2017).

3.2.1  Selenium, a Metalloid Selenium (Se) belongs to the non-metallic family but is considered as a metalloid because selenium has properties of both metals and non-metals. This chemical element was identified as a new substance in 1817 by the Swedish chemist Berzelius in leaden chamber mud during sulfuric acid production. It was given the name selenium in resemblance after the Greek goddess of the moon “Selene”, in homology to the chemically similar tellurium. Selenium belongs to the elements of group 16 (chalcogens) of the periodic table, together with oxygen, sulphur, tellurium and polonium. Se has thus similar chemical and physical properties with these elements. Particularly, selenium displays a chemical behavior similar to sulfur, and as a result, Se is found associated with naturel sulfides, e.g. pyrite and chalcopyrite, mainly in trace concentrations. Selenium can conduct electricity or act as an insulator and display unipolar conductance of electricity. The most outstanding physical property of crystalline selenium is its photoconductivity. The gray, metallic form of Se is the most stable under ordinary conditions. This form has the unusual property of greatly increasing in

3  Technologies to Remove Selenium from Water and Wastewater

213

electrical conductivity when exposed to light (Combs and Combs 1986; Simonoff and Simonoff 1991; White and Dubrovsky 1994; Frankenberger and Engberg 1998; Plant et al. 2003; Fernández-Martínez and Charlet 2009; Lide 2009; INERIS 2011; INRS 2011; Eklund and Persson 2014; Health Canada 2014). Selenium has five redox states: -II, -I, 0, +IV and +VI (Table 3.1). There are also six stable selenium isotopes. Selenium speciation is complex because it exists in different oxidation states in nature, in both inorganic and organic forms, in solid, liquid and gas phase (Haygarth 1994). Biological transformations of selenium are manifold. In selenium-contaminated environments, a cocktail of different Se species may be present due to the variety of transformations, which poses a major challenge for the analysis of selenium speciation, and therefore the difficulty of choosing an appropriate method for selenium removal. Elemental selenium Se(0) is present in nature as a non-soluble form. Se(0) occurs in different allotropic forms and may be amorphous or crystalline; seven different crystalline forms have been described. Se chemistry is complex, e.g. selenium can be mixed with sulfur in any ratio. Elemental selenium is precipitated after microbial reduction processes and also by inorganic processes. The oxidation states Se(-II) and Se(-I) are stable under strongly reducing conditions in metallic selenides and organic compounds (Table 3.1). The -II state is also present in nature as a product of microbial processes, e.g. H2Se, and in metastable anions such as selenosulfate SO3Se2-, which is involved in the formation of metallic selenides. Volatile H2Se is an analogue to H2S under strongly reducing conditions. Organic species include: volatile compounds formed upon bacterial methylation, e.g. dimethylselenide and dimethyldiselenide; products of microbial methylation processes, e.g. dimethylseleniumsufilde and dimethyseleniumdisulfide; selenocysteine present in organic tissues; selenomethionine in

Table 3.1  The different forms of selenium compounds Oxidation state -II, -I 0

Solubility Form (g/L) Inorganic Insoluble Inorganic Insoluble

+IV

Inorganic >850 (20°C)

Selenatec

+VI

Inorganic >840 (20°C)

Selenium dioxide Methylated species Selenomethionine Selenocysteine

+IV -II

Gas Organic

-II -II

Organic Organic

Compound Selenides Elemental selenium Seleniteb

H2Se Na2SeO3 c Na2SeO4

a

b

pKa1 and pKa2 3.8 and 14a

Example(s) H2Se, SO3Se2crystalline trigonal Se 2.7 and 8.54 SeO32-, HSeO3-, H2SeO3 −2.01 and SeO42-, HSeO41.8 SeO2 (CH3)2Se

1.56 and 9.5 C5H11NO2Se 2, 5.2 C3H7NO2Se and 10

214

E. Lichtfouse et al.

plants; trimethylselenonium, an urinary metabolite; selenoproteins such as proteins and enzymes; and selenocyanate present in industrial wastewaters. Table 3.2 compares the different forms of selenium from a chemical and geochemical point of view. Inorganic species of selenium +IV and +VI are soluble Table 3.2  Main differences between the different forms of selenium from a chemical and geochemical point of view Elemental selenium – insoluble – garlic smell in water – mildly reducing – colloidal: 1 nm to 1 μm – red-color – associated with natural organic matter Selenate – soluble – not odiferous in water – weakly adsorbed (outer-sphere) – adsorption: surface hydroxyl group, oxyhydroxide charge – net positive at lower pH: for anion adsorption – net negative at higher pH: for cation adsorption – two adsorption types: outer-sphere (weak ionic charge attraction - ionic strength, pH), inner-sphere (strong covalent bonds - pH) – low precipitation and adsorption capacities – aqueous chemistry similar to that of sulfates – low reduction kinetics – microbial reduction to selenite and elemental selenium Selenides – soluble unless metals are present – garlic smell in water – precipitation – strongly reducing – may adsorb weakly – H2Se toxicity Selenite – soluble – not odiferous in water – more strongly adsorbed (inner-sphere) – reduction kinetics faster than selenate – microbial reduction to elemental selenium and selenide – reduced by organic acids – more toxic than selenate Organic selenium – formed from biological activity, e.g. in plants, microorganisms – some compounds are volatile

3  Technologies to Remove Selenium from Water and Wastewater

215

Scheme 3.1  Structures of selenite and selenate oxyanions, showing distances between Se and O atoms

oxyanions known as selenite and selenate, respectively (Scheme 3.1). Oxyanion selenite SeO32- has a C3v symmetry with pyramidal shape which is reduced to C1 upon protonation. The oxidation state +IV also exists as gaseous selenium dioxide SeO2, which is present in volcanic eruptions and combustions processes. Selenite is a weak acid occurring as H2SeO3, HSeO3- and SeO32-, depending upon pH. Selenite is present in oil refinery wastewaters, for example. In the moderate redox potential range, selenite is the major species, and selenite mobility is governed by adsorption/ desorption processes on various solid surfaces including organic matter. Oxyanion selenate SeO42- has a Td tetrahedral symmetry, reduced to C3v upon single protonation and to Cs upon double protonation. This fully oxidized form exists in solution as biselenate HSeO4- or selenate SeO42- with a pKa2 of 1.8. The doubly protonated species with pKa1 of −2.01 do not exist under natural conditions. Selenate species are predominant in waters, sediments and soils. Noteworthy, selenate has high solubility,  low precipitation and adsorption capacities, and the aqueous chemistry of selenate and sulfate are quite similar (Fordyce 2007). The inorganic forms selenate and selenite are the two most common oxyanions in water, due in particular to their high solubility (White and Dubrovsky 1994). For example, these two most oxidized species are frequently encountered in surface waters and are transported mostly in particulate-associated form. Both oxyanions display a high bioavailability and bioaccumulation potential (Sharma et al. 2015).

3.2.2  Industrial Applications of Selenium In the industry, selenium is considered as a rare metal with many and varied applications. The most important applications are electronics, optoelectronics, glass, metallurgy, chemistry, pigments and nutrition (Surai and Taylor-Pickard 2008; Bañuelos et al. 2014; Gu et al. 2019; Zhu et al. 2019). These industries require only small quantities of Se. In addition, despite numerous applications, selenium currently competes in industry with other elements such as silicon, germanium, sulfur and tellurium. For example, many photocell applications using selenium are replaced by

216

E. Lichtfouse et al.

other devices using elements that are more sensitive and more readily available than selenium. Due to its photovoltaic, photoconductive and photosensitive properties and its ability to transform light into electricity, selenium is used extensively in electronic and electrical industries, mainly in rectifiers for electroplating, and in photoelectric cells (Zhu et al. 2019). Below its melting point, Se is a semiconductor, a property highly sought-after in the electronics sector. Other uses include metal alloys such as lead plates used in storage batteries, light meters, solar cells, detectors, colorimeters, photographic photosensitizers, photocopiers, and semiconductors (Gu et  al. 2019; Zhu et al. 2019). Rechargeable metal batteries using selenium or selenide as cathodes have attracted considerable attention during the past few years because selenium/selenide possess a high volumetric energy density that is comparable to that of sulfur, and a moderate output voltage (Gu et al. 2019). Various other devices such as alarm devices, mechanical opening and closing devices, safety systems, and television rely on properties of selenium. Selenium is widely used in photocopying, xerography to copy documents, e.g. photographic toner, toning of photographs, and in sound films. For example, Se is used artistically to intensify and extend the tonal range of black and white photographic images. The second major application of selenium is in the glass industry. Here, Se is used to decolorize glass by neutralizing the glass greenish tinge caused by ferrous impurities. Selenium is also added to glass to reduce the transmission of solar heat. As a red pigment, Se is used to make ruby-red colored glasses and enamels for ceramics and steel ware. For example, selenium imparts to glass a clear red color that is useful in signal lights. Se is used in the paint, plastic and ceramic industries to produce stains and dyes. Selenium is also used as an additive to stainless steel, e.g. to control porosity in stainless-steel castings. The third main use is sodium selenite for animal feeds and food supplements. Selenium and derivatives are also used in the manufacture of many chemicals such as pigments, reducing agents, parasiticides, bactericides, insecticides, fertilizers, fungicides, herbicides, lubricating oils and solvents; they are also used in metallurgical applications, and in the military field. Selenium pigments are used to color many products such as plastics, paints, inks, glass, enamels and rubber. Selenious acid is used in the steel industry as an etchant. Selenium find applications as lubricants for metal polishing and is replacing lead in brass alloy plumbing fittings. Selenium serves for the vulcanization of rubber to increase resistance to abrasion; here selenium diethyldithiocarbonate acts as an accelerator and vulcanizing agent. In the rubber industry, Se also promotes resistance to heat and oxidation, and increases the resilience of rubber. Selenium is used as a paint and varnish remover, and as a solvent for rubber resins and other organic substances. Selenium catalysts, due to their high efficiency, moderate reaction conditions, good functional compatibility and excellent selectivity, have attracted a lot of interest over the past two decades as recently discussed by Shao et al. (2019). Finally, selenium find applications in pharmaceuticals where selenium sulphide is used as a catalyst, and in cosmetology and agriculture. Selenium has been applied in producing cortisone, and radioactive selenium has been utilized in radiography.

3  Technologies to Remove Selenium from Water and Wastewater

217

Nanoparticles containing selenium were proposed for the diagnosis and therapy of Alzheimer’s disease (Gupta et al. 2019), and as biosensors for detection of biological targets (Gandin et al. 2018). Selenium sulfide and other selenium-based compounds are added to anti-dandruff shampoos. As fungicide, Se is used in the manufacture of deodorant. Selenium is also used as a protective agent against pests. Mechora (2019) recently summarized pest control by selenium. Selenium can repel pests, reduce their growth, or cause toxic effects while having a positive effect on plant growth. Accumulated selenium in plants protects plants against aphids, weevils, cabbage loopers, cabbage root flies, beetles, caterpillars, and crickets due to both deterrence and toxicity. Mechora (2019) concluded that the use of selenium can be an alternative pest management method to conventional plant protection products that pose environmental and health problems. Further developments are expected in the near future in the following domains: cancer prevention and therapy (Gandin et al. 2018; Sayehmiri et al. 2018; Tan et al. 2019b), biomedical, imaging and detection (Gandin et al. 2018; Gupta et al. 2019), designing selenium functional foods and beverages (Adadi et al. 2019), and agrochemistry (Garduño-Zepeda and Márquez-Quiroz 2018; Mechora 2019). Selenium has most probably a protective role against the development of prostate cancer, and therefore selenium supplementation is suggested for prevention of prostate cancer (Sayehmiri et al. 2018). Selenium supplementation also yielded promising results concerning radioprotection in tumor patients and should be considered as a promising adjuvant treatment option in subjects with a relative selenium deficit (Muecke et al. 2018). The agronomic and genetic biofortification of crops with selenium are novel strategies to improve the nutraceutical quality of staple crops. Biofortification with selenium in agricultural crops is increasingly becoming a solution to solve trace element deficiency in the human population, as well as to increase the content of bioactive compounds (Garduño-Zepeda and Márquez-Quiroz 2018).

3.2.3  Selenium in the Environment Selenium has been found in all environmental compartments of the Earth, including the atmosphere, geosphere, hydrosphere and biosphere, due to the presence of both natural, e.g. alteration and leaching of the earth’s crust, volcanism, and anthropogenic processes such as fossil fuel combustion and mining (Sharma et al. 2015). Se is found in the environment in both inorganic and organic forms, it is generally present in selenide, selenite or selenate forms, and more rarely in the elemental state (Shrift 1964; Rosenfeld and Beath 1964; Combs and Combs 1986; Combs 1988; Ihnat 1989; Frankenberger and Benson 1994; Frankenberger and Engberg 1998; Lemly 2002, 2004; Plant et al. 2003; Fernández-Martínez and Charlet 2009; Chapman et al. 2010; Bañuelos et  al. 2014; Pettine et  al. 2015; Sharma et  al. 2015; Wu and Sun 2016; Donner et al. 2018; Kumkrong et al. 2018; LeBlanc et al. 2018; Etteieb et al. 2020). Selenium concentrations in natural waters is a subject of intense interest (Sharma et al. 2015). Selenium in groundwater is of both natural, e.g. from inputs such as soil

218

E. Lichtfouse et al.

leaching, or anthropogenic origin, e.g. industrial emissions, metal refining and coal combustion. Inorganic forms of selenate and selenite and organic species such as mono- and dimethylated derivatives have been reported in aquatic systems, the two common forms being oxyanions (White and Dubrovsky 1994; Lemly 2002; Fernández-Martínez and Charlet 2009; Chapman et al. 2010; Sharma et al. 2015). However, the valence states of the selenium in water are still poorly known and few studies have been published on transformations of selenium species in aqueous systems (Sharma et al. 2015; Pettine et al. 2015; Wu and Sun 2016; Donner et al. 2018). This depends on Se origin, e.g. natural leaching of soils or industrial discharges. Due to their high solubility, oxyanions are mobile in soils, which explains their presence in some catchment areas. Due to rock erosion, selenium also enters inland waters and oceans. Selenate and selenite can be found in fresh and salt waters. Therefore, not only drinking water play an important role in human exposure to selenium but also the oceans via seafood (Fernández-Martínez and Charlet 2009; Santos et al. 2015). The global average concentration of selenium in freshwater is 0.02 μg/L and less than 0.08 μg/L in seawater. Groundwater generally contains higher selenium levels, from a few μg/L to more than 50 μg/L, than surface water due to contact with rocks (Fernández-Martínez and Charlet 2009; Santos et  al. 2015). In industrial wastewater, concentrations are much higher (Crini 2017). The bioavailability, mobility and reactivity of selenium in waters are also determined by Se speciation. Selenium speciation is controlled by physical processes, e.g. adsorption effects of soil and sediments; chemical processes, e.g. pH, redox conditions, organic matter, and  presence of competitive ions; and biological processes, e.g. bacterial transformations. In natural waters of pH 6–9, under oxidizing conditions, Se(VI) is predominant in a divalent ionic form, i.e. the selenate anion SeO42-. Se(IV) is the most frequent form encountered under reducing conditions; the hydrogenoselenite ion HSeO3- being the dominant form below pH 8.15. At a pH above 8.15, the divalent selenite anion SeO32- is the dominant form. Insoluble reduced species such as elemental selenium and selenides are generally released as colloidal suspensions into surface waters. The organic forms of selenium found in natural waters are produced by microbiological assimilation and degradation (Ivanenko et al. 2018). A comprehensive discussion on the chemistry and biogeochemistry of selenium in terms of variation of pH and redox conditions can be found in the reviews by Fernández-Martínez and Charlet (2009) and by Sharma et al. (2015).

3.2.4  Selenium and Industrial Emissions Selenium in water also come from anthropogenic sources such as mining, oil refining, agricultural irrigation and discharges from industries producing and using selenium compounds (Crini 2017; Dinh et al. 2018; Donner et al. 2018; Tabelin et al. 2018). Mining and metal refining (copper), coal mining and fossil fuel combustion in coal-fired power plants are industrial sectors particularly affected by selenium releases (Plant et  al. 2003; Wen and Carignan 2007; Fernández-Martínez and Charlet 2009; BRGM 2011; Health Canada 2014; Santos et  al. 2015). These Se

3  Technologies to Remove Selenium from Water and Wastewater

219

releases are mainly responsible for the migration of selenium to different compartments of the environment. In mining wastewaters, selenium is found at concentrations from 3 μg/L to more than 12 mg/L.  Wastewater resulting from flue-gas-desulfurization contains selenium at a typical concentrations of 1–10 mg/L (Santos et  al. 2015). Other contaminated wastewaters such as coal mining pond water, uranium mine discharges, gold mine effluent, petroleum industry wastewater and lead smelter wash water also contain significant levels of selenium. Other human activities such as agricultural irrigation can also promote corrosion of selenium-­bearing iron rocks, thus leaching selenium into aquifers as soluble oxyanions of Se(IV) and Se(VI). Selenium is also present in sewage treatment plants, mainly in sludge (Crini 2017).

3.3  Selenium and Water: A Substance of Concern? Selenium is ubiquitous in the environment. Life is exposed to selenium through Se presence in soil, air, water and food, the two latter being the major sources of human exposure (Rosenfeld and Beath 1964; Combs and Combs 1986; Combs 1988; Reilly 2002; Fernández-Martínez and Charlet 2009; ANSES 2012; Fordyce 2013; Health Canada 2014; Santos et al. 2015; Donner et al. 2018). In Canada, food is recognized as the main source of selenium, while in France this is debated (Crini 2017). Nowadays, it is widely recognized that selenium is both essential to human health and toxic in high quantities or in certain forms (Mayland 1994; Amiard 2011; Chauhan et al. 2019; Ibrahim et al. 2019; Kieliszek 2019; Varlamova and Maltseva 2019). Humans need to absorb small amounts of selenium daily in order to maintain good health and to prevent diseases, and food usually contains enough selenium because Se is naturally present in cereals, breads, nuts, fish, eggs, milk, meat, crab and tuna (Combs 1988; Mayland 1994; Hatfield 2001; Surai and Taylor-Pickard 2008; Amiard 2011). Selenium is essential to human health and should be present in the diet of all age groups to provide an adequate intake because Se is a key component of amino acids, e.g. selenocysteine and selenomethionine in selenoproteins found in all forms of life (Cai et  al. 2019; Ibrahim et  al. 2019; Varlamova and Maltseva 2019). In adults and teenagers, Se daily needs are estimated at 50–200 μg/ day while in children they range from 30 to 120 μg/day. The water consumed also provides selenium (Reilly 2002; Crini 2017; Kieliszek 2019). However, despite Se nutritional benefits, it is one of the most toxic natural elements, and therefore particularly followed from a regulatory point of view. In the field of water pollution, a worldwide problem is the presence of selenium in drinking water, groundwater and wastewater. Although benefits and toxicity of selenium are known, the levels considered to represent a threat to humans and environment are not yet well established (Fordyce 2013; Santos et al. 2015). For drinking water, according to European standards (Directive 98/83, European Commission, EU 1998) and Canadian water quality guidelines (Kwon et al. 2015), the selenium threshold should not exceed 10 μg/L, while the upper limit set by the United State Environmental Protection Agency USEPA (2004, 2016) is 50  μg/L.  The EU has

220

E. Lichtfouse et al.

revised the drinking water directive on 16 December 2020 for total selenium which is now 20 µg/L (https://eur-­lex.europa.eu/legal-­content/EN/TXT/PDF/?uri=CELE X:32020L2184&from=EN). Standards are also different in other countries: 50 μg/L (class 2) or 20 μg/L (class 1) for South Africa, and 10 μg/L for Australia and New Zealand. Regulatory wastewater discharge standards for selenium also vary from country to country. There is a lack of clarity in national, European and international regulations concerning selenium (Fordyce 2013; Santos et  al. 2015; Crini 2017; Kumkrong et al. 2018; LeBlanc et al. 2018). In France, various departments such as Seine-et-Marne, Essonne, Loiret, Vienne and Marne have selenium issues because rocks and soils are naturally rich in selenium are concerned. The quality limit for selenium in tap water intended for human consumption is 10 μg/L according to the Code de la Santé Publique (Public Health Code, Order of 11 January 2007). However, this value is exceeded in several regions for certain drilling waters (AFSSA 2007; Vilaginès 2010). In some seleniferous areas, natural water concentrations can reach values of 50–300  μg/L (Vilaginès 2010). Over the past ten years, French public authorities have thus taken on the dimension of the selenium phenomenon (Crini 2017). A French Mayor must request an operating derogation if the selenium content is 10–40 μ/L, with a restriction on use, particularly for children under 4 years of age if the content exceeds 20 μ/L. This three-year derogation must allow time to provide a technical solution to this excess of selenium. Many cities are waiting for the standard to be raised because treatment would be very expensive. L’Agence Nationale de Sécurité, The National Health Security Agency, has set the value at 30 μg/L in October 2012 without any consumption restrictions. For several years, the World Health Organization has also recommended changing the threshold from 10 to 40 μg/L, it has provisionally set the value at 40 μg/L in 2011. To solve the problem of selenium in drinking water, some American regions have chosen to mix water to reduce material requirements and costs, others have chosen low selenium supply sources or to remove excess selenium using treatment processes in public distribution systems or at home (Crini 2017). For industry, selenium is also considered as an emerging hazardous substance. In Europe, the release of selenium-contaminated water into the environment through industrial processes is currently a regulatory, environmental and health concern (Crini 2017).

3.4  Methods to Remove Selenium from Water 3.4.1  Main Treatment Methods There is actually no single method to ensure adequate treatment, and, in practice, a combination of different methods is used to achieve the targeted water quality in the most economical way; for example, achieving residual concentrations below the European Union regulation limit of 10 μg/L for drinking water. Selenium removal methods are classified into three categories: chemical, biological and physical

3  Technologies to Remove Selenium from Water and Wastewater

221

Technologies available for selenium removal

Chemical methods -

coprecipitation coprecipitation and adsorption reduction techniques iron reduction and coprecipitation zero valent iron coagulation/flocculation ferric coagulation and precipitation electrocoagulation electrodialysis cementation photoreduction

Physical methods -

adsorption reverse osmosis nanofiltration ion-exchange evaporation

Biological methods -

microbial reduction bacterial treatment algal-bacterial removal wetlands biochemical reactors bioremediation phytoremediation biosorption biomass

Fig. 3.1  Classification of the technologies for selenium removal

technologies (Fig. 3.1, Koren et al. 1992; Twidwell et al. 1999; Shamas et al. 2009; Sandy and DiSante 2010; Moore and Mahmoudkhani 2011). Technologies can also be classified into conventional methods, e.g. coprecipitation, reduction-adsorption and oxidation-reduction; established removal processes, e.g. adsorption, ion-­ exchange, membrane filtration, and emerging removal methods such as fluidized bed reactors, algal-bacterial removal and catalyzed cementation (Crini 2017). Nonetheless, only few methods are commonly used by the industry, mainly for economic reasons. In general, the removal of contaminants from effluents is done by chemical and biological means, with research focusing on effective and less costly combinations. Table  3.3 summarizes advantages and disadvantages of the main technologies for the treatment of selenium-contaminated water and industrial wastewater.

3.4.2  Coprecipitation The main technology used at the industrial scale to remove metals and metalloids from wastewaters is direct precipitation. However, selenium is difficult to precipitate because it is mainly found in the form of oxyanions in wastewaters, which are highly soluble and perfectly stable (Plotnikov 1960). In addition, similar to selenocyanate, selenite and selenate may also form a variety of stable complexes with several transition metals. One solution is to use iron coprecipitation coupled, if necessary, with a coagulation-­flocculation step (Fig. 3.2). This technology is also known as ferrihydrite adsorption and some authors mention a double technology: coprecipitation and adsorption (Merrill et al. 1986, 1987; MSE 2001; Twidwell et al. 2005; Gingerich

Chemical technologies Main Technology characteristic(s) Coprecipitation Uptake of selenium Iron and separation of coprecipitation the products formed ferrihydrite A two-step process: adsorption precipitation with precipitation-­ simultaneous adsorption adsorption of chemical selenium on precipitation ferrihydrite and cupric ferric hydroxide coprecipitation surfaces References

Merrill et al. (1986, 1987); Koren et al. (1992); Kapoor et al. (1995); Kashiwagi and Kokufuta (2000); MSE (2001; Fujita et al. (2002); Twidwell et al. (2005); Gingerich et al. (2018); He et al. (2018b)

Disadvantages Chemical consumption (FeCl2, FeSO4) Physicochemical monitoring of the effluent: pH-dependent (optimal conditions in the range of pH 4 to 6) Requires primary treatments: coagulation and flocculation Ineffective in removal of selenium at low concentration (less than 5 μg/L) Not able to remove selenate Competition from phosphate, silica and other elements (vanadium present in flue gas desulfurization effluents) High sludge production, handling (thickening and dewatering) and disposal problems (management, treatment, cost) Toxicity of sludge considered as hazardous waste Requires tertiary treatment in some cases: media filtration, pH adjustment Requires a step of oxidation of the selenocyanate to selenite before removal Potential release of selenium from ferrihydrite residuals

Advantages

Technologically simple: the contaminants are sequestered in the mineral during crystal growth Both economically advantageous and efficient (to treat wastewater): short treatment time Integrated physicochemical process Established by US EPA as best demonstrated available technology for selenite removal Widely implemented in industry Very efficient for selenite Significant reduction of other contaminants

Table 3.3  Comparison of chemical, biological, physical and physicochemical technologies for the treatment of selenium-contaminated water and industrial wastewater

Reduction-adsorption Reduction of Reduction selenium, uptake techniques Ferrous hydroxide and separation of the products formed Reduction-­ A two-step process: adsorption ferrous iron is Reduction-­ added resulting in adsorption-­ the reduction of precipitation selenate to selenite Adsorption and the subsequent adsorption and/or coprecipitation of selenite by ferric hydroxide or ferrihydrite

Coagulation/Flocculation A two-step process: Coagulation-­ formation of flocs precipitation and separation of Coagulation-­ the products formed adsorption

Relatively simple and low cost method Well-known technology similar to the coprecipitation method Widely implemented in industry Efficient to treat water or wastewater: chemical reduction of Se(IV) to Se(0)

Process simplicity Widely implemented in industry with coprecipitation technology Good sludge settling and dewatering characteristics Significant reduction of suspended solids Well-known mechanism: coagulation, formation of hydroxide flocs, adsorption, and precipitation Chemical consumption pH-dependent: optimal pH between 8 and 9 Ineffective in removal of selenium at low concentration (less than 5 μg/L) Not as effective at the reduction of selenate to selenite as zero-valent iron Presence of dissolved oxygen, bicarbonate and nitrate can interfere with reduction High sludge production, handling and disposal problems (management, treatment, cost) Toxicity of sludge considered as hazardous waste

Chemical consumption: high coagulant dosage Physicochemical monitoring of the effluent: pH-dependent Fe-based coagulants are much more efficient that Al-based coagulants in selenium removal Large volume of sludge generated

(continued)

Zingaro et al. (1997); MSE (2001); Geoffroy and Demopoulos (2011); Sharrad et al. (2012); Ling et al. (2015); Stefaniak et al. (2018)

Koren et al. (1992); Kapoor et al. (1995); Hu et al. (2015); Santos et al. (2015); Staicu et al. (2015a, 2015b); He et al. (2018b)

Chemical technologies Main Technology characteristic(s) Zero-valent iron Chemical reduction Reduction-­ and adsorption adsorptionRedox reactions coprecipitation Oxyanions are converted to elemental selenium by redox reactions Zero-valent iron used as reducer and catalyst

Table 3.3 (continued)

References MSE (2001); Sandy and DiSante (2010); Tang et al. (2014a, b); Ling et al. (2015); van Hullebusch (2017); Gingerich et al. (2018); He et al. (2018b); Stefaniak et al. (2018); Das et al. (2019); Sharma et al. (2019)

Disadvantages Laboratory scale Technology is temperature and pH dependent: temperature strongly affects the reaction kinetics; in some cases, pH adjustment may be necessary with longer residence times The reduction of selenium is highly dependent on the surface characteristics of iron and dissolved oxygen concentration in groundwater Presence of other oxyanions such as sulfate, phosphates, nitrates, nitrites and carbonates can interfere with reduction The presence of oxygen can compete with reduction of selenium Requires tertiary treatment to remove ferrous iron: aeration followed by clarification for example; generation of a large sludge volume (management, cost); spent zero-valent iron must be removed, replaced and disposed (hazardous waste)

Advantages

A relatively inexpensive and moderately strong reducing agent Efficient to remove selenite and selenate from different salt containing solutions to low concentrations For wastewaters, Se(IV) is favored over Se(VI) Adapted to low pollutant loads although kinetics are slow Well established procedures (temperature, pH, residence time) and maintenance Relative easy physicochemical monitoring of the effluent: efficiency is a function of pH (optimal pH between 6 and 7) and equilibrium concentrations of ferrous and ferric iron Well-known reduction mechanisms; also provides ferric and ferrous iron products for precipitation and adsorption of selenium Significant reduction of other contaminants Used as pre-treatment for constructed wetlands

Effective treatment for the reduction, coagulation and separation of selenate Generation of coagulants in situ without chemical addition (iron or aluminum sacrificial anodes) pH control is not necessary Adaptation to different pollutant (including Se) loads and different flow rates May coagulate other colloids further enhancing the coprecipitation of selenium Efficient elimination of suspended solids and metals Effective in treatment of drinking water supplies for small or medium sized communities Widely used in the mining industries

Catalyzed reduction Efficient to remove selenite and selenate to low concentrations Well-known technology similar to the zerovalent iron method Significant reduction of other contaminants

Electrocoagulation Electrolysis Electrochemical Ferrous iron is method produced by Electrolysis applying direct electrical current to the effluent where an iron anode oxidizes to form ferrous iron that can reduce selenate and the resultant ferric iron can coprecipitate selenite

Cementation Catalyzed reduction Galvanic cementation Modified zero-valent technology Enhanced cementation

Initial cost of the equipment Cost of maintenance (sacrificial anodes, requires an electrical current, cleaning of cathodes plates) Requires addition of chemicals (salts, flocculants) in polymetallic wastewaters Variability in wastewater salinity vary the oxidation of the iron anode to ferrous ion resulting in potential perturbations Anode passivation and sludge deposition on the electrodes that can inhibit the electrolytic process in continuous operation Cost of sludge treatment Requires post-treatment to remove high concentrations of iron

Long residence times In some cases, pretreatment (pH) is required Cost of maintenance (sacrificial anodes) The presence of high concentrations of interfering anions can be a problem Media requires periodic replacement and disposal (hazardous waste) Sludge volume generation (management, cost) Requires tertiary treatment to remove copper or nickel discharges

(continued)

Mollah et al. (2004); Mavrov et al. (2006); GOLDER (2009); Sandy and DiSante (2010); Gingerich et al. (2018); He et al. (2018b); Stefaniak et al. (2018); Hansen et al. (2019); Kazeem et al. (2019)

MSE (2001); Twidwell et al. (2005); GOLDER (2009); Shamas et al. (2009); van Hullebusch (2017); He et al. (2018b)

Chemical technologies Main Technology characteristic(s) Electrochemical methods ElectroAs in remediation electrocoagulation, ions released from sacrificial electrodes react with selenium to form insoluble precipitants The technique uses a low-level direct current as cleaning agent through several transport mechanisms and electrochemical reactions to form ferrous hydroxides Photoreduction A two-step process: Chemical chemical reduction reduction and by irradiation and adsorption adsorption method advanced oxidation

Table 3.3 (continued)

MSE (2001); GOLDER (2009); Shamas et al. (2009); Sandy and DiSante (2010); Labaran and Vohra (2014); He et al. (2018b); Mohapatra and Kirpalani (2019); Sharma et al. (2019)

Efficient to remove selenite and selenate to low concentrations Method uses irradiation of ultraviolet at a certain wavelength and in presence of titanium dioxide to convert oxyanions into elemental selenium

Laboratory scale Emerging technology Formation of toxic hydrogen selenide gas

Sandy and DiSante (2010); Laboratory scale Baek et al. (2013); Gingerich Emerging technology et al. (2018) High electricity consumption Not effective in removing high selenate concentrations

Oxidation of iron anode prevents water oxidation and formation of oxygen, and produces ferrous and ferric ions leading to the formation of hydroxides in a mixed cell Not dependent on dissolved oxygen The method can be applied to both anoxic and oxic system to remove selenate

References

Disadvantages

Advantages

Biological technologies Main Technology characteristic(s) Biological methods Use of biological Biological cultures: anaerobic conversions or aerobic bacteria, Biological algae, fungi, plants volatilization On basis of Biological conversion of reduction soluble oxyanions Bioreduction to elemental Microbial insoluble selenium reduction Uptake, reduction Heterotrophic and volatilization microbial processes Reduction Bioremediation Bioreactors Biofilm reactors Sludge-based Bioreactors Activated sludge Constructed wetlands Phytoremediation Microbial reduction Enhanced in situ Reduction of reduction oxidized forms of selenium using inoculated microorganisms in anoxic conditions

MSE (2001); Yee et al. (2007); GOLDER (2009); Hunter and Manter (2009); Zhu et al. (2009); Vriens et al. (2014); Dessì et al. (2016); Tan et al. 2016); Mal et al. (2017); van Hullebusch (2017); He et al. (2018b); Stefaniak et al. (2018); Tan et al. (2018); Gebreeyessus and Zewge (2019); Paul and Saha 2019); Rene et al. (2019); Zhang et al. (2019a); Zeng et al. (2019)

Necessary to create an optimally favorable environment Requires management and maintenance of the microorganisms and/or physicochemical pretreatment Large space requirements High operating time Performance is difficult to control (e.g. treatment of flue gas desulfurization effluents) Some systems have low chloride tolerance (94d

References Malhotra et al. (2020) Malhotra et al. (2020) Malhotra et al. (2020) Richards et al. (2011) Richards et al. (2011) Richards et al. (2011) Richards et al. (2011) Kharaka et al. (1996) Chung et al. (2010) Chung et al. (2010) He et al. (2016a, b) He et al. (2016a, b) Chehayeb and Lienhard (2017) Chehayeb and Lienhard (2017) He et al. (2018a, b, c) He et al. (2018a, b, c) Cingolani et al. (2018)

Commercial membrane Laboratory made membrane c Data from the third stage of a multistage disc tube reverse osmosis (DT-RO) system d No information about the rejected Se species a

b

Effect of Speciation on Selenium Rejection  In water, selenium can occur in different oxidation forms (-II, IV, VI) which depends on the redox potential, pH, dissolved organic matter, and microbial activities (Kumar and Riyazuddin 2011). Se(IV) and Se(VI) species are the predominant Se forms in water (Chand and Prasad 2009). Se(IV) species are found in moderately oxidizing environment (moderate

3  Technologies to Remove Selenium from Water and Wastewater

273

concentration of dissolved O2 as electron acceptor), while Se(VI) species are mostly present in oxidizing environment of high dissolved O2 concentration (Sharma et al. 2019). For instance, in environment conditions of pH 8.1 and with electron activity (pE) of 12.5, the concentration of Se(VI) would be estimated to be 10 times higher than Se(IV) concentration (Sharma et al. 2019). In nanofiltration/reverse osmosis, the speciation chemistry of a contaminant affects the rejection mechanisms in both size and charge mechanisms (Richards et al. 2009). The influence of the pH on the speciation of Se is shown in Fig. 3.7. He et al. (2016a, b) have investigated the effect of the speciation on Se(IV) and Se(VI) rejection by varying pH during the filtration experiments with a laboratory made nanofiltration membrane (see Fig. 3.7). Higher rejection (>90%) of Se(IV) was observed at pH  >  9, where the divalent SeO32-

Fig. 3.7  Speciation of selenium (Se) and its rejection with laboratory-made TFN membrane, adapted from He et al. (2016a, b); speciation of Se(IV) and Se(VI) was calculated at 25°C and CO2 partial pressure of 3.9·10-4 bar, by using the software MINTEQ v.3.1, KTH, Sweden

274

E. Lichtfouse et al.

becomes the dominant species as shown in Fig. 3.7a. In the other hand, the rejection of Se(VI) was higher than in the case of Se(IV) at all pH values due to the presence of the divalent SeO42- species in water from pH > 4 (see Fig. 3.7b). Similar behavior was observed with commercial nanofiltration membranes used to remove Se(VI) (Malhotra et al. 2020). Further, it is noted that the increase in Se(VI) rejection with pH is influenced by the surface charge characteristic of the membrane that becomes more negative with increasing the pH, like most polymeric nanofiltration/reverse osmosis membranes. To sum-up, pH variation in water and wastewater affect the efficiency of nanofiltration/reverse osmosis process especially when using loose nanofiltration membranes where the dominant separation mechanism is charge exclusion. In addition to water chemistry, the separation efficiency of nanofiltration/ reverse osmosis processes further depend on the operating conditions, such as pressure and recovery that are reviewed in the subsequent section. Effect of Operating Conditions  – Transmembrane Pressure and Recovery  In addition to water chemistry, nanofiltration/reverse osmosis performance is also influenced by the operating conditions, such as the applied (transmembrane) pressure and recovery (Ballet et  al. 2007; Lee et  al. 2015). Transmembrane pressure variations affect the water transport and hence the diluting effect (Verliefde et al. 2013), and determines flux and recovery. Regarding the recovery, which is the ratio of permeate production by feed volume, is also known by its influence on permeate quality and membrane fouling/scaling (Lee et al. 2015). Malhotra et al. (2020) have investigated the effect of transmembrane pressure on the Se(VI) rejection with loose (NF2) and tight (NF2) nanofiltration membranes, whereas, Chung et al. (2010) have studied the effect of the recovery and Se(VI) rejection with nanofiltration/reverse osmosis membranes. Results from both studies are shown in Fig. 3.8. Results shows that, the rejection of Se(VI) increased with the transmembrane pressure and from 15 bar the rejection remained constant for both nanofiltration membranes (Fig. 3.8a). This behavior is explained by solution-diffusion mechanism where the solvent flux increases with the transmembrane pressure resulting in high rejection (Verliefde et al. 2013). On the other hand, the observed decrease in rejection with recovery (see Fig. 3.8b) has been attributed to the increase of Se(VI) concentration at the membrane surface (concentration polarization) which enhances the solute transport through the membrane (Chung et  al. 2010). In addition, at higher recovery, the rejection of Se(VI) with the nanofiltration was more affected than with the reverse osmosis membrane where the ions transfer is known to be less convective and more diffusive (Pontié et al. 2008). In summary, the rejection of Se with nanofiltration/ reverse osmosis membranes increases with transmembrane pressure whereas it decreases with the recovery. However, further in-depth investigations are still needed to work at high recovery and thus to decrease the retentate quantity without affecting the efficiency of the nanofiltration/reverse osmosis process. Gaps in Knowledge in Selenium Removal by Nanofiltration/Reverse Osmosis  Besides the studies on the effect of pH and ultimately the speciation on the rejection of selenium with nanofiltration/reverse osmosis, future studies that

3  Technologies to Remove Selenium from Water and Wastewater

275

Fig. 3.8  Rejection of Se(VI) with nanofiltration/reverse osmosis membranes as a function (a) the transmembrane pressure (TMP) and (b) the recovery (source: (a) adapted from Malhotra et  al. (2020): Se(VI) 1600 μg/L and pH 8; (b) adapted from Chung et al. (2010): S(VI) 326 μg/L, pH 7.2 and pressure 8 bar)

investigate the impact of the solute-solute interactions. The presence of other compounds and how these interact with selenium, such as phosphates (in case of wastewater), hardness and organic matter (in case of natural water) has to date not been investigated. In this context, further research is needed to fully understand, for instance, the solute-solute interactions that can impact the rejection of Se and the complex transport mechanisms through the nanofiltration/reverse osmosis membrane. Regarding the impact of the operating conditions on selenium rejection with nanofiltration/reverse osmosis, futures studies need to focus on the velocity variations (in case of cross flow systems) that effect mass transfer. For the process itself, much more dedicated work focused on system design and the optimization of the operating parameters to achieve zero liquid discharge and eventually lower specific energy consumption.

3.5.5  Selenium Removal Using Electrodialysis Operating Principles of Electrodialysis  Electrodialysis is an electro-membrane process, in which the driving force is the electrical potential over the membrane stack that generates a direct electric current. Anions are moved from the cathode to the anode, while cations move in the other direction. While counter-ions can pass through the ion exchange membranes, co-ions are repulsed from similarly charged membranes and cannot be transported across. The target in electrodialysis is to only remove charged ions, therefore electrodialysis is generally incapable of removing non-charged elements from water streams. Figure  3.9 shows the basic concept of an electrodialysis process. Electrodialysis is more energy efficient compared to reverse osmosis for desalinating brackish water. Specifically, in desalinat-

276

E. Lichtfouse et al.

Diluate

Cl anion

2-

Se species (IV, VI)

Na cation

Concentrate Anode

CEM

AEM

CEM

AEM

CEM

AEM

-

+

Anion Exchange Membrane (AEM)

Cathode

Electrochemical potential (driving force)

+ + + + + + + + + + + +

Cation Exchange Membrane (CEM)

Diluate

2-

-

2-

Feed

Fig. 3.9 Principle of an electrodialysis process. (Source: Mehran Aliaskari, Eggenstein-­ Leopoldshafen, Germany)

ing brackish water with salinity of 4 g/L total dissolved solids or less; this higher efficiency also depends on the recovery of the process and extent of salt removal (Karimi et al. 2015; Patel et al. 2021). Both selenium species, selenite (Se(IV)) and selenate (Se(VI)) are anions in the neutral pH range. In consequence, electrodialysis can in principle remove common selenium species from water. Electrodialysis can indeed remove different trace inorganic contaminants including fluoride, nitrate and arsenic(V) from brackish water to some extent (Onorato et  al. 2017). However, reports on selenium removal by electrodialysis are to date very sparse. Effect of Operative Parameters on Selenium Removal by Electrodialysis  In the absence of data on selenium removal by electrodialysis, one can consider the removal of similar ions. Kim et al. (2012) have investigated the competitive separation of single- versus double-charged ions by electrodialysis in different flowrates. The results suggest that at higher flowrates, double-charged ions show a somewhat increased transport compared to single charged ions. The higher transport of double-­ charged ions in higher flowrates is attributed to the thinner boundary layer, as the limiting factor for ion transport in electrodialysis is this layer, in which diffusivity and mobility of the ions play an important role in transport. Selenate is analogous to sulfate and has similar chemistry and diffusivity in water (see Table 3.5), therefore similar removal by electrodialysis can be expected for them. Sosa-Fernandez et al. (2019) show that an increase in current intensity had little impact on the final sulfate removal, and high sulfate removals up to 97% can be achieved when diluate solution is depleted of other anions such as chloride. However, Onorato et al. (2017) have investigated the effect of applied electrical potential on selenium removal. By an increase in electrical potential from 12 to 18 v, selenium removal is improved from

3  Technologies to Remove Selenium from Water and Wastewater

277

Table 3.5  Diffusion coefficient of different anions in water at 25°C and 1 atm Diffusion coefficient Se (IV)a Se (VI)b Sulfatec Chloridec Nitratec Fluoridec Ionic mobility (×1012 mol s/kg) 0.36 0.38 0.43 0.82 0.77 0.59 Diffusivity (×109 m2/s) 0.89 0.94 1.06 2.03 1.90 1.47 Iida et al. (2011) Yuan-Hui and Gregory (1974) c HDR Engineering Inc (2002) a

b

33% to 48%. Karimi and Ghassemi (2015) have shown that electric potential has a bigger impact on divalent ions compared to monovalent ions. Figure 3.10 shows the speciation of selenium compounds over a pH range of 1 to 13. As the pH of the sample water tested by Onorato et al. (2017) is reported at 7 to 9, selenium is divalent in this pH. Effect of Speciation on Selenium Removal with Electrodialysis  In a general study with brackish water that contained some Se, Onorato et al. (2017) have also investigated Se removal as a function of pH. The brackish water had an electric conductivity of 8290 μS/cm (≈5.3 g/L NaCl). In this, selenate (Se(VI)) was the dominant Se species at a relatively low concentration of 20  μg/L (WHO guideline: 40  μg/L (WHO 2017)), while no data for selenite Se(IV) was reported. Removal of Se(VI) is shown in Fig. 3.10, along with speciation simulation for both Se(VI) and Se(IV) in pH 1 to 13 by Minteq (v3.1, KTH, Sweden). Removal of selenate (Se(VI)) at neutral pH was about 40%, and indeed relatively low. Unfortunately, no removal data for selenite (Se(IV)) is available. However, Based on the speciation of Se(IV) in Fig. 3.10, it is expected that Se(IV) removal would be maximum in between pH 4 to 7, where selenium is dominantly monovalent. This is expected due to the higher diffusivity and lower hydration number of monovalent species compared to divalent ions (Marcus 1997; Tanaka et al. 2013). The increased removal of Se(VI) in pH 8 to 11 (see Fig. 3.10) was attributed to the presence of calcium, which may have caused a co-precipitation of various ions (i.e. CaSeO4) (Onorato et al. 2017). The complexity of the real water used in these experiments makes it difficult to draw meaningful conclusions about selenium removal by electrodialysis from these results. In the case of specific wastewaters that require pretreatment steps before the electrodialysis process, the pH of the feed must be considered as selenium removal is pH-­ dependent and hence will be affected strongly by pH variation. For example, Gingerich et al. (2018) investigated selenium removal from wastewater of coal-fired power plants. After a pretreatment process necessary for removing other contaminants, the pH of the wastewater is high and therefore the solution pH may have an impact on selenium removal by electrodialysis based on the very limited observations from Onorato et al. (2017). Effect of Competing Anions on Selenium Removal by Electrodialysis  While electrodialysis can in principle remove trace selenium concentrations, presence of other anions in water may hinder Se removal through competition. While there is no data available for Se, this process is likely from observations with other similar ions. For

278

E. Lichtfouse et al.

Fig. 3.10  Speciation of selenium VI (top) and IV (bottom) over pH 1 to 13 and removal of selenium VI, adapted from Onorato et al. (2017); electric potential 12 V; feed salinity 5.3 g/L; membranes Neosepta CMX-SB & AMX-SB, Tokuyama Soda Ltd., Japan; the highlighted pH range shows the real pH of sample water used which was then modified by adding HCl and NaOH)

example, in experiments removing sulfate from saline water, Sosa-Fernandez et al. (2019) showed that the removal of sulfate was always lower than that of chloride. It is expected that oxoselenium anions removal follow the same trend as sulfate because selenium and sulphur are both in the 15th group of the periodic table and have chemical similarities, e.g. ionic mobility and diffusion coefficient in Table 3.5. Ionic mobility and diffusion coefficient of selenium species and different anions are presented in Table 3.5. Lower mobility and diffusivity of both Se(IV) and Se(VI), suggest that an increase in the salinity of the feed water may result in a lower removal of selenium due to the competition of other present anions in water to be transported.

3  Technologies to Remove Selenium from Water and Wastewater

279

Gaps in Knowledge in Selenium Removal by Electrodialysis  Very few studies on Se removal by electrodialysis have been reported to date. Several handbooks (HDR Engineering Inc 2002) and patents (Wallace 2013a, b) mention electrodialysis as a process to remove selenium from water, in the reviewed documents no references or experimental results other than reports of the charged state of selenium species in water was observed. Moreover, in mentioned patents (Wallace 2013a, b), it is suggested to have a selenium removal process after electrodialysis to ensure complete selenium removal from product water. This implies that no complete removal was expected. Additionally, the complex chemistry of real brackish groundwater in the only published study (Onorato et al. 2017) makes it difficult to conclude the effectiveness of electrodialysis for selenium removal. Presence of other compounds (e.g. hardness, organic matter, multivalent ions and salinity) in water may affect selenium removal by electrodialysis. More systematic experiments are needed to understand the mechanisms and the ability of various electrodialysis membranes and operating parameters to achieve effective Se removal.

3.5.6  R  emediation of Solutions Containing Selenium by Chitosan-Enhanced Ultrafiltration Pressure-driven membrane processes such as reverse osmosis and nanofiltration are able to reject species with very low molecular weight such as metal ions. In reverse osmosis, the separation of various species of a mixture is related directly to their relative transport rates within the membrane, which are determined by their diffusivity and solubility in the membrane material. In nanofiltration, the separation of solutes results from a complex mechanism including steric hindrance dependent on the relative sizes of the pores and the solutes (Ferry 1936), Donnan exclusion resulting from the Coulomb interaction between charged solutes and the membrane fixed charge (Donnan 1995), dielectric exclusion in terms of both Born dielectric effect, resulting from the solvation energy barrier due to the decrease of the dielectric constant of the solution inside the membrane pores (Bowen et  al. 1997; Déon et  al. 2012) and image charge effects due to the interaction between the ions and the polarization charges, induced by the ions, at the interface between the membrane matrix and the solution inside the pores (Yaroshchuk 2000; Szymczyk et al. 2005). Reverse osmosis is generally used when a total retention of ions is desired whereas nanofiltration is rather dedicated to partial demineralization of waters. However, even if reverse osmosis and nanofiltration processes are capable of rejecting ions, unfortunately, they produce relatively low permeation fluxes and requires high transmembrane pressures to obtain significant flux (Table  3.6), which result in a high energy cost. Oppositely, ultrafiltration requires lower applied pressures while providing higher permeation fluxes, but the small solute removal performances are much lower due to larger pores.

280

E. Lichtfouse et al.

Table 3.6  Comparison of different pressure-driven membrane processes

Molecular weight cut-off (kDa) Pore diameters (nm) Transmembrane pressure (bar) Permeation flux (L/ hm2) Exclusion mechanisms

Reverse osmosis nonylphenol >4-n-nonylphenol ~4-n-octylphenol >4-nonylphenol. If the values obtained for 4-tert-octylphenol and 4-nonylphenol are compared, an important decrease is observed, suggesting that the difference in the chemical structure between these two substances may explain the different obtained values. On the contrary, sucrose- and starch-based polymers were not able to remove alkylphenols (