Water Pollution and Remediation: Heavy Metals [1st ed.] 9783030524203, 9783030524210

Table of contents :
Front Matter ....Pages i-xvii
Sources and Health Risks of Rare Earth Elements in Waters (Willis Gwenzi, Nyarai M. Mupatsi, Munyaradzi Mtisi, Allan A. Mungazi)....Pages 1-36
Removal of Heavy Metal Pollutants from Wastewater Using Zerovalent Iron Nanoparticles (Ambavaram Vijaya Bhaskar Reddy, Muhammad Moniruzzaman, Gajulapalle Madhavi)....Pages 37-72
Water Treatment Chemicals for Pollution Minimization and Management (Vinod P. Sharma)....Pages 73-84
Advanced Treatment of Real Wastewater Effluents by an Electrochemical Approach (Govindaraj Divyapriya, Jaimy Scaria, T. S. Anantha Singh, P. V. Nidheesh, D. Syam Babu, M. Suresh Kumar)....Pages 85-122
Unconventional Adsorbents for Remediation of Metal Pollution in Waters (Md. Mostafizur Rahman, Rubaiya Akter, Mashura Shammi)....Pages 123-146
Desalination Technology for Water Security (Mashura Shammi, Md. Mostafizur Rahman, Mohammed Mofizur Rahman)....Pages 147-176
Nanotechnology for the Remediation of Heavy Metals and Metalloids in Contaminated Water (Roop Singh Lodhi, Subhasis Das, Aiqin Zhang, Paramita Das)....Pages 177-209
Hybrid Treatment Technologies for the Treatment of Industrial Wastewater (Vikas S. Hakke, Murali Mohan Seepana, Shirish H. Sonawane, Anand Kishore Kola, Ramsagar Vooradi)....Pages 211-241
Removal of Heavy Metals in Biofiltration Systems (Andreas Aditya Hermawan, Amin Talei, Babak Salamatinia)....Pages 243-258
Contamination and Health Impact of Heavy Metals (Naveen Patel, Deepak Chauhan, Shraddha Shahane, Dhananjai Rai, Md. Zafar Ali Khan, Umesh Mishra et al.)....Pages 259-280
Tin-Based Compounds for Water Remediation (Ivana Aguiar, Daniela Oreggioni, María E. Pérez Barthaburu)....Pages 281-312
Some Effective Methods for Treatment of Wastewater from Cu Production (Vesna Krstić)....Pages 313-440
Heavy Metal Removal from Wastewater Using Adsorbents (Reena Malik, Bhaskaran, Meena, Suman Lata)....Pages 441-469
Electroanalytical Techniques for the Remediation of Heavy Metals from Wastewater (Muhammad Altaf, Naila Yamin, Gulzar Muhammad, Muhammad Arshad Raza, Munazza Shahid, Raja Shahid Ashraf)....Pages 471-511
Mechanisms and Approaches for the Removal of Heavy Metals from Acid Mine Drainage and Other Industrial Effluents (Vhahangwele Masindi, Muhammad S. Osman, Memory Tekere)....Pages 513-537
Removal of Dyes and Heavy Metals with Clays and Diatomite (Nadia Tahari, Houwaida Nefzi, Abdelkader Labidi, Sameh Ayadi, Manef Abderrabba, Jalel Labidi)....Pages 539-569
Back Matter ....Pages 571-581

Citation preview

Environmental Chemistry for a Sustainable World 53

Inamuddin Mohd Imran Ahamed Eric Lichtfouse  Editors

Water Pollution and Remediation: Heavy Metals

Environmental Chemistry for a Sustainable World Volume 53

Series Editors Eric Lichtfouse , Aix-Marseille University, CNRS, IRD, INRAE, 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 Journals Environmental Chemistry Letters http://www.springer.com/10311

More information about this series at http://www.springer.com/series/11480

Inamuddin • Mohd Imran Ahamed Eric Lichtfouse Editors

Water Pollution and Remediation: Heavy Metals

Editors Inamuddin Department of Applied Chemistry Aligarh Muslim University Aligarh, India

Mohd Imran Ahamed Department of Chemistry Aligarh Muslim University Aligarh, India

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

ISSN 2213-7114 ISSN 2213-7122 (electronic) Environmental Chemistry for a Sustainable World ISBN 978-3-030-52420-3 ISBN 978-3-030-52421-0 (eBook) https://doi.org/10.1007/978-3-030-52421-0 © 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

Water pollution is a major health issue for living organisms and ecosystems as a result of increasing industrialisation and urbanisation. In particular, toxic metal pollution originates from many wastes such as industrial waste, radioactive waste, agricultural waste, sewage, plastics and oil leakage. This book summarises recent basic and applied knowledge on sources, health risks and remediation methods of metals and metalloids. Rare earth elements, as emerging pollutants from the electronic industry, are reviewed in Chap. 1 by Gwenzi et al. in terms of sources, behaviour, human intake, risk of exposure and mitigation. Chapter 2 by Reddy et al. presents the application of zerovalent iron nanoparticles to remove heavy metals (Figure). Chemicals used for treatment of water pollution are reviewed by Sharma in Chap. 3. Chapter 4 by Divyapriya et al. reviews electrochemical advanced oxidation processes to treat real wastewater and discusses hybrid processes in treating waste from textile, petrochemical, paper and tannery industries and from urban effluents and landfill leachates. Chapter 5 by Rahman et al. depicts the removal of heavy metals, arsenic (As), lead (Pb), cadmium (Cd) and chromium (Cr) using unconventional low-cost novel sorbents such as waste materials, biochar, industrial wastes, nanomaterials and metal-organic frameworks.

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Preface

Chapter 6 by Shammi et al. addresses desalination to improve water security. Nanomaterials-based methods for the remediation of heavy metals and metalloids are reviewed in Chap. 7 by Lodhi et al. Chapter 8 by Hakke et al. presents conventional and hybrid methods for treating the pharmaceutical and textile industrial effluents, with focus on hydrodynamic cavitation techniques with the microand ultra-filtration, photo-catalysis with ultra-filtration, and the use hydro-gels. Chapter 9 by Hermawan et al. reviews heavy metals retention in biofiltration and discusses selection of the filter materials. Lead, chromium, cadmium, copper and arsenic are reviewed in terms of toxicity and remediation in Chap. 10 by Patel et al. Chapter 11 by Aguiar et al. details the use of tin-based nanocompounds for water remediation. Nanoparticles synthesis and photocatalytic properties for metals, dyes and microbial decontamination are discussed, with focus tin sulfide nanoparticles. Remediation of acid wastewater from Cu production by adsorption, nanomaterials, photocatalysis, nano zero valent iron and phytoremediation is reviewed in Chap. 12 by Krstić. Metal pollution includes Cu, Pb, Cd, Cr, As, Zn and Hg. Chapter 13 by Malik et al. discusses adsorption to remove metals from water, with materials such as plants, biomass, carbon forms and composites, nanocomposites containing polymers, metal oxides, zeolites and metal-organic frameworks. Remediation techniques of wastewater polluted by metals are presented in Chap. 14 by Altaf et al. Chapter 15 by Masindi et al. reviews mechanisms and approaches employed for the removal of heavy metals from acid mine drainage and

Preface

vii

other industrial effluents. Clays and diatomaceous earth for removal of dyes and heavy metals are presented in Chap. 16 by Tahari et al., with focus on surface modification to improve adsorption. Aligarh, India Aligarh, India Aix-en-Provence, France

Inamuddin Mohd Imran Ahamed Eric Lichtfouse

Contents

1

Sources and Health Risks of Rare Earth Elements in Waters . . . . . Willis Gwenzi, Nyarai M. Mupatsi, Munyaradzi Mtisi, and Allan A. Mungazi

2

Removal of Heavy Metal Pollutants from Wastewater Using Zerovalent Iron Nanoparticles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Ambavaram Vijaya Bhaskar Reddy, Muhammad Moniruzzaman, and Gajulapalle Madhavi

3

4

Water Treatment Chemicals for Pollution Minimization and Management . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Vinod P. Sharma Advanced Treatment of Real Wastewater Effluents by an Electrochemical Approach . . . . . . . . . . . . . . . . . . . . . . . . . . . Govindaraj Divyapriya, Jaimy Scaria, T. S. Anantha Singh, P. V. Nidheesh, D. Syam Babu, and M. Suresh Kumar

1

37

73

85

5

Unconventional Adsorbents for Remediation of Metal Pollution in Waters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 123 Md. Mostafizur Rahman, Rubaiya Akter, and Mashura Shammi

6

Desalination Technology for Water Security . . . . . . . . . . . . . . . . . . 147 Mashura Shammi, Md. Mostafizur Rahman, and Mohammed Mofizur Rahman

7

Nanotechnology for the Remediation of Heavy Metals and Metalloids in Contaminated Water . . . . . . . . . . . . . . . . . . . . . . 177 Roop Singh Lodhi, Subhasis Das, Aiqin Zhang, and Paramita Das

ix

x

Contents

8

Hybrid Treatment Technologies for the Treatment of Industrial Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 211 Vikas S. Hakke, Murali Mohan Seepana, Shirish H. Sonawane, Anand Kishore Kola, and Ramsagar Vooradi

9

Removal of Heavy Metals in Biofiltration Systems . . . . . . . . . . . . . 243 Andreas Aditya Hermawan, Amin Talei, and Babak Salamatinia

10

Contamination and Health Impact of Heavy Metals . . . . . . . . . . . . 259 Naveen Patel, Deepak Chauhan, Shraddha Shahane, Dhananjai Rai, Md. Zafar Ali Khan, Umesh Mishra, and Vinod Kumar Chaudhary

11

Tin-Based Compounds for Water Remediation . . . . . . . . . . . . . . . . 281 Ivana Aguiar, Daniela Oreggioni, and María E. Pérez Barthaburu

12

Some Effective Methods for Treatment of Wastewater from Cu Production . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 313 Vesna Krstić

13

Heavy Metal Removal from Wastewater Using Adsorbents . . . . . . . 441 Reena Malik, Bhaskaran, Meena, and Suman Lata

14

Electroanalytical Techniques for the Remediation of Heavy Metals from Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 471 Muhammad Altaf, Naila Yamin, Gulzar Muhammad, Muhammad Arshad Raza, Munazza Shahid, and Raja Shahid Ashraf

15

Mechanisms and Approaches for the Removal of Heavy Metals from Acid Mine Drainage and Other Industrial Effluents . . . . . . . . 513 Vhahangwele Masindi, Muhammad S. Osman, and Memory Tekere

16

Removal of Dyes and Heavy Metals with Clays and Diatomite . . . . 539 Nadia Tahari, Houwaida Nefzi, Abdelkader Labidi, Sameh Ayadi, Manef Abderrabba, and Jalel Labidi

Index . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 571

About the Editors

Inamuddin Ph.D., is an assistant professor at the Department of Applied Chemistry, Aligarh Muslim University, Aligarh, India. He has extensive research experience in multidisciplinary fields of analytical chemistry, materials chemistry, electrochemistry, renewable energy, and environmental science. He has published about 175 research articles in various international scientific journals, 18 book chapters, and 110 edited books with multiple well-known publishers. His current research interests include ion-exchange materials, a sensor for heavy metal ions, biofuel cells, supercapacitors, and bending actuators. Mohd Imran Ahamed received his Ph.D. degree on the topic “Synthesis and characterization of inorganic-organic composite heavy metals selective cationexchangers and their analytical applications,” from Aligarh Muslim University, Aligarh, India, in 2019. He has published several research and review articles in the journals of international recognition. His research works include ion-exchange chromatography, wastewater treatment and analysis, bending actuator, and electrospinning. Eric Lichtfouse is a biogeochemist at Aix Marseille University who has invented carbon-13 dating, a molecular-level method allowing to study the dynamics of organic compounds in temporal pools of complex environmental media. He is the Chief Editor of the journal Environmental Chemistry Letters and the book series Sustainable Agriculture Reviews and Environmental Chemistry for a Sustainable World. He is the author of the book Scientific Writing for Impact Factor Journals, which includes an innovative writing tool: the Micro-Article.

xi

Contributors

Manef Abderrabba Laboratory of materials, molecules and applications, IPEST, Preparatory Institute of Scientific and Technical Studies of Tunis, University of Carthage, La Marsa, Tunisia Ivana Aguiar Área de Radioquímica, Departamento Estrella Campos, Facultad de Química, Universidad de la República, Montevideo, Uruguay Rubaiya Akter Department of Environmental Sciences, Jahangirnagar University, Dhaka, Bangladesh Muhammad Altaf Department of Chemistry, GC University Lahore, Lahore, Pakistan Raja Shahid Ashraf Department of Chemistry, GC University Lahore, Lahore, Pakistan Sameh Ayadi Laboratory of materials, molecules and applications, IPEST, Preparatory Institute of Scientific and Technical Studies of Tunis, University of Carthage, La Marsa, Tunisia D. Syam Babu CSIR-National Environmental Engineering Research Institute, Nagpur, Maharashtra, India Bhaskaran Department of Chemistry, University of Delhi, Delhi, India Vinod Kumar Chaudhary Department of Environmental Sciences, Dr Ram Manohar Lohia Avadh University, Ayodhya, Uttar Pradesh, India Deepak Chauhan Department of Civil Engineering, BIET, Jhansi, Jhansi, Uttar Pradesh, India Paramita Das Department of Chemical Engineering, Indian Institute of Science Education and Research Bhopal, Bhopal, Madhya Pradesh, India Subhasis Das Environmental and Industrial Biotechnology Division, The Energy and Resources Institute, New Delhi, India xiii

xiv

Contributors

Govindaraj Divyapriya Environmental Water Resources Engineering Division, Department of Civil Engineering, Indian Institute of Technology Madras, Chennai, Tamilnadu, India Willis Gwenzi Biosystems and Environmental Engineering Research Group, Department of Soil Science and Agricultural Engineering, Faculty of Agriculture, University of Zimbabwe, Harare, Zimbabwe Vikas S. Hakke Chemical Engineering Department, National Institute of Technology, Warangal, Telangana, India Andreas Aditya Hermawan Discipline of Civil Engineering, School of Engineering, Monash University Malaysia, Subang Jaya, Selangor, Malaysia Md. Zafar Ali Khan Department of Civil Engineering, Govt. Polytechnic College, Gonda, Gonda, Uttar Pradesh, India Anand Kishore Kola Chemical Engineering Department, National Institute of Technology, Warangal, Telangana, India Vesna Krstić Mining and Metallurgy Institute Bor, Bor, Republic of Serbia Technical Faculty, University of Belgrade, Bor, Republic of Serbia M. Suresh Kumar CSIR-National Environmental Engineering Research Institute, Nagpur, Maharashtra, India Abdelkader Labidi Laboratory of materials, molecules and applications, IPEST, Preparatory Institute of Scientific and Technical Studies of Tunis, University of Carthage, La Marsa, Tunisia Chemistry Department, El Manar University, University of Sciences of Tunis, Tunis, Tunisia Jalel Labidi Biorefinery Processes Research Group, Department of Chemical and Environmental Engineering, University of the Basque Country (UPV/EHU), Donostia-San Sebastian, Spain Suman Lata Department of Chemistry, Deenbandhu Chhotu Ram University of Science and Technology, Murthal, Haryana, India Roop Singh Lodhi Department of Chemical Engineering, Indian Institute of Science Education and Research Bhopal, Bhopal, Madhya Pradesh, India Gajulapalle Madhavi Department of Chemistry, Sri Venkateswara University, Tirupati, Andhra Pradesh, India Reena Malik Department of Chemistry, Deenbandhu Chhotu Ram University of Science and Technology, Murthal, Haryana, India

Contributors

xv

Vhahangwele Masindi Magalies Water, Scientific Services, Research & Development Division, Brits, South Africa Department of Environmental Sciences, School of Agriculture and Environmental Sciences (CAES), University of South Africa (UNISA), Florida, South Africa Council of Scientific and Industrial Research (CSIR), Built Environment (BE), Hydraulic Infrastructure Engineering (HIE) Group, Pretoria, South Africa Meena Department of Chemistry, Deenbandhu Chhotu Ram University of Science and Technology, Murthal, Haryana, India Umesh Mishra Department of Civil Engineering, NIT, Agartala, Agartala, Tripura, India Muhammad Moniruzzaman Centre of Research in Ionic Liquids, Universiti Teknologi PETRONAS, Seri Iskandar, Perak, Malaysia Department of Chemical Engineering, Universiti Teknologi PETRONAS, Seri Iskandar, Perak, Malaysia Munyaradzi Mtisi Biosystems and Environmental Engineering Research Group, Department of Soil Science and Agricultural Engineering, Faculty of Agriculture, University of Zimbabwe, Harare, Zimbabwe Gulzar Muhammad Department of Chemistry, GC University Lahore, Lahore, Pakistan Allan A. Mungazi Biosystems and Environmental Engineering Research Group, Department of Soil Science and Agricultural Engineering, Faculty of Agriculture, University of Zimbabwe, Harare, Zimbabwe Nyarai M. Mupatsi Biosystems and Environmental Engineering Research Group, Department of Soil Science and Agricultural Engineering, Faculty of Agriculture, University of Zimbabwe, Harare, Zimbabwe Houwaida Nefzi Laboratory of materials, molecules and applications, IPEST, Preparatory Institute of Scientific and Technical Studies of Tunis, University of Carthage, La Marsa, Tunisia Chemistry Department, El Manar University, University of Sciences of Tunis, Tunis, Tunisia Biorefinery Processes Research Group, Department of Chemical and Environmental Engineering, University of the Basque Country (UPV/EHU), Donostia-San Sebastian, Spain P. V. Nidheesh CSIR-National Environmental Engineering Research Institute, Nagpur, Maharashtra, India Daniela Oreggioni Departamento de Desarrollo Tecnológico, Centro Universitario Regional del Este- Universidad de la República, Rocha, Uruguay

xvi

Contributors

Muhammad S. Osman Council of Scientific and Industrial Research (CSIR), Built Environment (BE), Hydraulic Infrastructure Engineering (HIE) Group, Pretoria, South Africa Naveen Patel Department of Civil Engineering, NIT, Agartala, Agartala, Tripura, India María E. Pérez Barthaburu Área de Radioquímica, Departamento Estrella Campos, Facultad de Química, Universidad de la República, Montevideo, Uruguay Md. Mostafizur Rahman Department of Environmental Sciences, Jahangirnagar University, Dhaka, Bangladesh Mohammed Mofizur Rahman Alexander Von Humboldt International Climate Protection Fellow, Institute for Technology and Resources Management in the Tropics and Subtropics (ITT), TH Cologne – University of Applied Sciences, Cologne, Germany Dhananjai Rai Department of Civil Engineering, BIET, Jhansi, Jhansi, Uttar Pradesh, India Muhammad Arshad Raza Department of Chemistry, GC University Lahore, Lahore, Pakistan Babak Salamatinia Discipline of Chemical Engineering, School of Engineering, Monash University Malaysia, Subang Jaya, Selangor, Malaysia Jaimy Scaria CSIR-National Environmental Engineering Research Institute, Nagpur, Maharashtra, India Murali Mohan Seepana Chemical Engineering Department, National Institute of Technology, Warangal, Telangana, India Shraddha Shahane Department of Civil Engineering, NIT, Agartala, Agartala, Tripura, India Munazza Shahid Department of Chemistry, University of Management and Technology, Lahore, Pakistan Mashura Shammi Department of Environmental Sciences, Jahangirnagar University, Dhaka, Bangladesh Vinod P. Sharma CSIR-Indian Institute of Toxicology Research, Lucknow, India T. S. Anantha Singh Department of Civil Engineering, School of Technology, Pandit Deendayal Petroleum University, Gandhinagar, Gujarat, India Department of Civil Engineering, Natinal Institute of Technology, Calicut, Kerala, India Shirish H. Sonawane Chemical Engineering Department, National Institute of Technology, Warangal, Telangana, India

Contributors

xvii

Nadia Tahari Laboratory of materials, molecules and applications, IPEST, Preparatory Institute of Scientific and Technical Studies of Tunis, University of Carthage, La Marsa, Tunisia Chemistry Department, El Manar University, University of Sciences of Tunis, Tunis, Tunisia Biorefinery Processes Research Group, Department of Chemical and Environmental Engineering, University of the Basque Country (UPV/EHU), Donostia-San Sebastian, Spain Amin Talei Discipline of Civil Engineering, School of Engineering, Monash University Malaysia, Subang Jaya, Selangor, Malaysia Memory Tekere Department of Environmental Sciences, School of Agriculture and Environmental Sciences (CAES), University of South Africa (UNISA), Florida, South Africa Ambavaram Vijaya Bhaskar Reddy Centre of Research in Ionic Liquids, Universiti Teknologi PETRONAS, Seri Iskandar, Perak, Malaysia Quality Control División, Ultra International Limited, Ghaziabad, Uttar Pradesh, India Ramsagar Vooradi Chemical Engineering Department, National Institute of Technology, Warangal, Telangana, India Naila Yamin Department of Chemistry, GC University Lahore, Lahore, Pakistan Aiqin Zhang State Laboratory of Surface and Interface Science and Technology, Henan Collaborative Innovation Center of Environmental Pollution Control and Ecological Restoration, Zhengzhou University of Light Industry, Zhengzhou, People’s Republic of China

Chapter 1

Sources and Health Risks of Rare Earth Elements in Waters Willis Gwenzi , Nyarai M. Mupatsi, Munyaradzi Mtisi, and Allan A. Mungazi

Contents 1.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2 Rare Earth Elements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2.1 Global Occurrence and Production of Rare Earth Elements . . . . . . . . . . . . . . . . . . . . . . 1.2.2 Properties and High-Technology Applications of Rare Earth Elements . . . . . . . . . . 1.3 Occurrence and Behavior of Rare Earth Elements in Aquatic Systems . . . . . . . . . . . . . . . . . . . 1.3.1 Anthropogenic Sources . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.3.2 Dissemination and Behavior of Rare Earth Elements in Aquatic Systems . . . . . . . 1.3.3 Environmental Health Risks . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.4 Assessment and Mitigation of Health Risks . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.4.1 Assessment of Human and Ecological Health Risks . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.4.2 Prevention and Control of Health Risks . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.5 Future Research Directions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.5.1 Source Partitioning and Behavior in Aquatic Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.5.2 Ecotoxicology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.5.3 Human Toxicology and Epidemiology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.5.4 Human Exposure and Health Risks in Developing Countries . . . . . . . . . . . . . . . . . . . . . 1.6 Summary, Conclusions, and Future Directions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

3 6 6 7 9 9 10 12 19 19 20 22 22 23 23 24 24 26

Abstract Anthropogenic rare earth elements widely used in high-technology applications are prevalent in the aquatic environment, thus constituting emerging contaminants. Yet reviews on the anthropogenic sources, behavior, and potential health risks of rare earth elements remain limited. The current chapter seeks to (1) highlight anthropogenic sources, behavior, and human intake pathways of rare earth elements, (2) discuss the human and ecological health and exposure risks of rare earth

W. Gwenzi (*) · N. M. Mupatsi · M. Mtisi · A. A. Mungazi Biosystems and Environmental Engineering Research Group, Department of Soil Science and Agricultural Engineering, Faculty of Agriculture, University of Zimbabwe, Harare, Zimbabwe e-mail: [email protected] © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature Switzerland AG 2021 Inamuddin et al. (eds.), Water Pollution and Remediation: Heavy Metals, Environmental Chemistry for a Sustainable World 53, https://doi.org/10.1007/978-3-030-52421-0_1

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W. Gwenzi et al.

elements, (3) present a conceptual outline for assessing and mitigating health risks, and (4) identify the key thematic areas for further research. Anthropogenic hotspot sources of rare earth elements include wastes and wastewaters from medical facilities, rare earth elements mining and mineral processing, high-technology electrical and electronic industries, petroleum refineries, rare earth elements-enriched fertilizers and livestock feeds, and recycling plants for postconsumer electronic and electrical goods. The dissemination of rare earth elements from sources into the various environmental compartments is controlled by anthropogenic (industrial discharges) and hydrological processes. Human exposure occurs via occupational inhalation in rare earth elements-based industries, ingestion of contaminated food, and medical applications including magnetic resonace imaging. To date, evidence exists documenting rare earth elements in human body parts including the brain, hair, nails, milk, serum, and sperms. High concentrations of rare earth elements reduce plant growth and nutritional quality, impaired biochemical function in plants, and induce neurotoxicity, acute and chronic health effects, and genotoxicity in aquatic animals. The uptake, partitioning, and bioaccumulation of rare earth elements may also occur along the trophic levels in aquatic ecosystems. Human health risks include (1) severe damage to nephrological systems and nephrogenic systemic fibrosis induced by gadolinium-based contrast agents used in medical applications, (2) induced sterility and anti-testicular effects in males, (3) dysfunctional neurological disorder and reduced intelligent quotient, (3) fibrotic tissue injury, (4) pneumoconiosis, and (5) oxidative stress and cytotoxicity. In developing countries, the health risks of rare earth elements may be considerably high due to the following: (1) weak and poorly enforced environmental and public health regulations, (2) overreliance on untreated drinking water, and (3) lack of human health surveillance systems for early detection and treatment of human health effects. However, limited empirical data exist to establish the relationship between rare earth elements in the aquatic environments and their health effects. A conceptual outline for assessing and mitigating the health risks and thematic areas for further research were highlighted. Keywords Anthropogenic sources · Ecological effects · Ecotoxicology · Environmental reservoirs · Human health effects · Human intake pathways · Lanthanides

Acronyms Ce CeCl3 Cu Dy Er Eu Gd

Cerium Cerium trichloride Copper Dysprosium Erbium Europium Gadolinium

1 Sources and Health Risks of Rare Earth Elements in Waters

Ho La LaCl3 Lu Nd P680 Pm Pr REE(s) Sc Sm Tb Tm Y Yb Zn

1.1

3

Holmium Lanthanum Lanthanum trichloride Lutetium Neodymium Photosystem II, one of the photosynthetic reaction centers Promethium Praseodymium Rare earth elements Scandium Samarium Terbium Thulium Yttrium Ytterbium Zinc

Introduction

Rare earth elements consist of 17 elements, which include 15 lanthanides and 2 other elements (scandium, Sc, and yttrium, Y) (Li et al. 2013a, b). The 15 lanthanides are lanthanum (La), cerium (Ce), dysprosium (Dy), europium (Eu), gadolinium (Gd), lutetium (Lu), praseodymium (Pr), neodymium (Nd), promethium (Pm), samarium (Sm), terbium (Tb), holmium (Ho), erbium (Er), thulium (Tm), and ytterbium (Yb). Yttrium (Y) and scandium (Sc) exhibit physicochemical properties and behavior similar to that of lanthanides, hence are included among rare earth elements. Promethium, a radioactive element, has no naturally occurring long-lived and stable isotope; thus, it is excluded in environmental studies (Hu et al. 2006). Thus, the current chapter only focuses on lanthanides (lanthanum to lutetium) and yttrium and scandium. Although the term “rare earth elements” is used in this chapter for consistency with other literature, the term is a misnomer because the crystal abundances of rare earth elements are often comparable to, and in some cases even higher than, those of common metals such as silver and gold (Wedepohl 1995; Schüler et al. 2011; Henderson 2013). In this context, “rare” implies that rare earth elements occur in very low concentrations but are highly dispersed in most geological systems compared to common elements such as metals. The unique properties of rare earth elements, including permanent magnetism and high reactivity, make them ideal for high-technology applications including electrical and electronic devices and equipment, advanced ammunition systems and platforms, renewable energy, and medical applications. The increasing industrial applications of rare earth elements could lead to a corresponding increase in the

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release of rare earth elements into the environments via industrial wastes and wastewaters, thus raising human and ecological health concerns. An increasing body of literature has reported anthropogenic rare earth elements in aquatic systems in various countries (Hatje et al. 2016; Hissler et al. 2016). Anthropogenic rare earth elements, specifically gadolinium, have been used as transient chemical tracers for anthropogenic pollution in aquatic systems (Brünjes et al. 2016). Fraum et al. (2017) and Rogowska et al. (2018) present recent reviews on gadolinium and gadolinium contrast agents, while a review on the occurrence, behavior and fate, and health risks is presented in one of the current author’s papers (Gwenzi et al. 2018). Thus, rare earth elements can be considered as among emerging contaminants posing human and ecological health risks. Yet the ecotoxicology, mechanisms of action, and epidemiology of emerging contaminants are still poorly understood; thus, emerging contaminants remain unregulated (Petrović et al. 2003). In addition, until now, appropriate analytical methods were unavailable to detect the occurrence of emerging contaminants in the environment due to very low concentrations (Petrović et al. 2003; Gwenzi et al. 2018). Until recently, literature on emerging contaminants was dominated by harmful biological agents including antimicrobial resistance genes and the host antimicrobial-resistant microbes and synthetic organic contaminants particularly industrial solvents, fire retardants, illicit drugs, food additives and colorants, pharmaceuticals, and personal care products (Petrie et al. 2015; Gwenzi and Chaukura 2018). However, recent studies including one review by the current author have drawn attention to rare earth elements as emerging contaminants of human and ecological health concerns (Gwenzi et al. 2018). Rare earth elements qualify to be classified among emerging contaminants for the following reasons: (1) maximum guideline limits to safeguard human and ecological health are nonexistent, rendering the regulation of rare earth elements problematic (Kulaksız and Bau 2013); (2) routine environmental and human health monitoring and surveillance systems often exclude rare earth elements; (3) the concentrations in environmental media are often extremely low (ng/mL to μg/ml), which can only detected using highly sensitive advanced analytical equipment; and (4) the ecotoxicology and human toxicity including modes of actions and epidemiology, which are important for formulating maximum guideline limits, are still poorly understood. Recent field studies and reviews have documented the occurrence and health risks of rare earth elements of anthropogenic origin in aquatic systems (Tepe et al. 2014; Hatje et al. 2016; Gwenzi et al. 2018). These studies further point that rare earth elements are a new group of contaminants, which is currently overlooked by policy makers, the public, and researchers. The purpose of this chapter is to present an overview of anthropogenic sources, behavior and fate, and potential health risks associated with rare earth elements. Figure 1.1 presents an overview of the focus of the chapter. The specific objectives of the current chapter are to (1) highlight anthropogenic sources, behavior, and human intake pathways of rare earth elements, (2) discuss the human and ecological health and exposure risks of rare earth elements, (3) present a

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High Technology Applications of Rare Earth Elements REEs in green energy devices

Agrochemical industries (e.g. REE Fertilizers

Consumer electronic products

Medical applications

Livestock & aquacultural systems

Petroleum industries

Waste and Wastewaters from REE Industries

Environmental Reservoirs of rare earth elements: Industrial effluents

Electronic wastes

Mine wastes

Landfills

Wastewater systems

Discharges & Hydrological Processes Rare earth elements in aquatic systems Geochemical behaviour: Sorption Redox reactions Sedimentation Resuspension

Uptake by aquatic organisms

Uptake

Uptake

Human Exposure Pathways Ingestion of contaminated aquatic foods and drinking water

Health Risks of REEs Ecological risk: Phytotoxicity Genotoxicity Neurotoxicity Cytoptoxicity Oxidative stress

Human risks: Nephrogenic systemic fibrosis Damage to nephrological system Dysfunctional neurological disorder Fibrotic tissue injury Pneumoconiosis Anti-testicular effects and male sterility

Fig. 1.1 Summary depiction of the high-technology applications, environmental reservoirs, behavior, and health risks of rare earth elements in aquatic systems (modified after Gwenzi et al. 2018). REE(s), rare earth element(s); MRI, magnetic resonace imaging

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conceptual outline for assessing and mitigating health risks, and (4) identify the key thematic areas for further research.

1.2 1.2.1

Rare Earth Elements Global Occurrence and Production of Rare Earth Elements

Rare earth elements occur in geological systems and undergo mobilization and are transported as a coherent group; thus, the geogenic concentrations of rare earth elements exhibit the characteristic saw-tooth behavior (Gupta et al. 2014). Rare earth elements are highly reactive; hence, they do not occur as pure elements or uninterrupted ore bodies, but exist within other host mineral ores (Charalampides et al. 2015). The predominant rare earth oxides are xenotime, monazite, and bastnasite (Humphries 2012), while less common ones are gadolinite, allanite, ancylite, euxenite, parisite, lanthanite, yttrotungstite, yttrotantalite, stillwellite, fergusonite, samarskite, yttrialite, loparite, chevkinite, cerite, britholite, fluocerite, and cerianite (Haque et al. 2014). Rare earth elements may also coexist with base metals in mineral ores. The recovery of rare earth elements from the host ores is achieved through costly extractive and metallurgical processes that can release the rare earth elements from mineral ores and subsequently isolate them from the complex solution of various elements. Rare earth oxide reserves exist in Australia, Brazil, the Dominican Republic, the United States, Russia, and several African countries, including South Africa, Burundi, Kenya, and the Democratic Republic of the Congo, among others (Zhanheng 2011; Massari and Ruberti 2013). Globally, China has nearly half (48%) of the total reserves of rare earth elements, followed by the United States (12%), Commonwealth States (17%), India (3%), and then Australia (1%), while other countries including those in Africa account for the remainder 19% (United States of America Geological Survey 2010). The global supply chain of rare earth elements is dominated by China, which accounts for 98% of the total production, while other countries contribute the remainder (2%) (Alonso et al. 2012). Mining of rare earth elements in China started around 1990 and since then has played a dominant role in the rare earth elements’ supply chain, and several countries including the United States are wholly dependent on imports from China (United States of America Geological Survey 2008; Stone 2009). The global consumption of rare earth elements in 2015 was about 119,650 metric tons, and the annual growth rate is estimated to rise to 5% in 2020 as the global demand increases (Zhou et al. 2017). Rare earth elements are regarded as critical resources for high-technology applications because of their strategic and economic importance, coupled with the high risk of the global supply chain (Du and Graedel 2011; Graedel et al. 2015). Substantial literature exists on the global

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occurrence, production, and applications of rare earth elements in high-technology engineering (Du and Graedel 2011; Massari and Ruberti 2013). Recent advances in high-technology applications have witnessed an upsurge in mining, production, and industrial applications of rare earth elements.

1.2.2

Properties and High-Technology Applications of Rare Earth Elements

Rare earth elements and their compounds exhibit unique physicochemical properties, which are critical for their high-technology applications. These properties include the following: (1) rare earth elements with odd atomic numbers have lower relative abundances than those with an even ones, a phenomenon referred to as the “Oddo–Harkins rule” (Binnemans et al. 2013a, b); (2) unique electronic configuration, which accounts for the high reactivity of rare earth elements; thus, they readily react with several nonmetallic elements such as sulfur, oxygen, and hydrogen, thereby forming ionic complexes of rare earth elements with high coordination number often exceeding 6 and in some cases 12 (Tang and Johannesson 2003; Nockemann et al. 2006); (3) an increase in atomic number is accompanied by a decrease in the cationic radius, a phenomenon termed “lanthanide contraction” (Ramos et al. 2016); and (4) rare earth elements are considered as soft elements, which are malleable and ductile, with excellent catalytic, chemical, electrical, and permanent magnetism and optical properties including high luminescence (Redling 2006; Gai et al. 2013). Rare earth elements are essential components of high-technology electronic and engineered applications (Kulaksız and Bau 2011a, b). In summary, Table 1.1 shows that rare earth elements are used in various industrial applications, including (1) miniaturized electronic devices and appliances including mobile phones, (2) advanced weapon systems and platforms, (3) solar panels and wind turbines for renewable energy, (4) electronic and electronic applications including conductors and supercapacitors, and (5) contrast agents in magnetic resonance imaging. According to literature, the proportion of rare earth elements used in various industrial applications decreases in the order: 21% lightweight permanent magnets, followed by 20% catalysts, 18% alloys, 12% powders, and 7% phosphors (Paulick and Machacek, 2017). Specifically, neodymium is a critical component of high-performance neodymium (Nd)–iron (Fe)–boron (B) permanent magnets for generators and motors, while yttrium (Y) is used for the production of superconductors and laser technology (Du and Graedel 2011; Charalampides et al. 2015). Neodymium, yttrium, and other rare earth elements are also used in laser technology as dopants, while yttrium is also used to develop yttrium–aluminum–garnet lasers (Ancsombe 2002). Moreover, rare earth elements are key components of autocatalysts, lighting and display systems, wind turbines, and solar panels (Du and Graedel 2011; Haque et al. 2014) and are also widely applied in pharmacology and

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Table 1.1 An overview of industrial applications of high-technology rare earth elements Rare earth elements Ce, Eu, Gd, Tb, tm Er, Yb, Eu, Y Ce, La, Nd, Pr, Sc, Yb Dy, Gd, Nd, Pr, Tb, Tm, Pr, Nd, Gd, Tb, Dy, Sm La, Ce, Pr, Nd Ce, Dy, Er, Gd, Ho, La, Pr, Nd, Y, Ce, Dy, Eu, Gd, La, Tb, Y Y Er, Ho Nd, Y, Ce, La, Nd, Pr, Sm Y, Sm Sc, Tm Lu, Gd

Industrial applications Flat-screen displays Optical fibers Metallurgy

References Resende and Morais (2010) and Humphries (2012) Eliseeva and Bünzli (2011) Stegen (2015), Paulick and Machacek (2017)

Medical imaging

Naumov (2008), Xie et al. (2014), Stegen (2015)

Lightweight strong permanent magnets Auto-catalysts in petroleum refining Ceramics and glass additives

Haque et al. (2014), Xie et al. (2014), Charalampides et al. (2015) and Stegen (2015) Navarro and Zhao (2014)

Phosphors Oxygen sensors, medicines and drugs, radar Lasers, superconductors and capacitors Battery alloys Microwave filters Electron beam tubes Crystal scintillators

Naumov (2008), Campbell and Keane (2010), United States of America Geological Survey (2014) and Stegen (2015) Tan et al. (2015) Townley (2013), Charalampides et al. (2015), Lu et al. (2017) Ancsombe (2002) Charalampides et al. (2015) Naumov (2008) Naumov (2008) Naumov (2008)

biomedical applications (Thomsen 2017). For example, gadolinium is a well-known component of gadolinium-based contrast agents that are widely utilized for clinical and diagnostic medical applications including magnetic resonance imaging (Lu et al. 2017). For example, yttrium is utilized in the development of anticancer drugs such as TheraSphere®, which contain microspheres of yttrium-90 (Kulik et al. 2006; Lu et al. 2017). In agriculture, low doses of rare earth elements are used in fertilizers and livestock feeds as crop and livestock growth promoters (He et al. 2008, 2010; Wu et al. 2013). For example, on an annual basis, fertilizers doped with nitrates of lanthanum, neodymium, and cerium are applied to approximately more than six million hectares of agricultural land to increase crop yields and quality (Wang et al. 2001; Migaszewski and Gałuszka 2015a, b). The increase in mining, production, and subsequent applications of rare earth elements in various high-technology systems could be accompanied by increased release of rare earth elements in industrial wastes and wastewaters. In fact, the increased occurrence of rare earth elements in aquatic environment has been attributed to the increased production and applications of rare earth elements in high-technology applications (Klaver et al. 2014; Khan et al. 2016).

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1.3 1.3.1

9

Occurrence and Behavior of Rare Earth Elements in Aquatic Systems Anthropogenic Sources

Although the current chapter focuses on aquatic systems, it is noteworthy that such aquatic systems are connected to other environmental compartments (contaminated soils, tailings, wastewaters) through material flows and hydrological processes. Thus, the connectivity between aquatic systems and the contiguous environmental compartments may form a continuum of rare earth contamination. Accordingly, the concentrations of rare earth elements in mine wastes such as tailings (Li et al. 2010) and other rare earth reservoirs (Sloof 1995) exceed the baseline concentrations in the earth crust by several orders of magnitude. Once in the environment, hydrological processes (e.g., runoff) and anthropogenic activities (e.g., wastewater and effluent discharges) play a key role in the dissemination of rare earth elements into aquatic environments. To date, anthropogenic rare earth elements have been detected in aquatic environments in several countries such as Australia, Europe, the United Kingdom, and North America, with gadolinium being the most documented rare earth elements (Bau et al. 2006; Kulaksız and Bau 2011a, b; Hatje et al. 2016). Table 1.2 presents an overview of some of the rare earth elements detected in aquatic systems. Several anthropogenic sources emit rare earth elements into solid waste repositories (landfills, non-engineered waste dumps, mine waste rock dumps, and tailings dams and wastewater treatment plants, which in turn act as rare earth reservoirs). Specific anthropogenic hotspot sources of rare earth elements include (1) mining and processing operations of rare earth elements (Paulick and Machacek 2017; Victoria et al. 2017); (2) hospitals and clinical and diagnostic medical imaging facilities (Lawrence et al. 2009; Lawrence, 2010; Kulaksız and Bau 2011a, b; Möller et al. 2011, 2014); (3) pharmaceutical and drug industries (Hutchinson et al. 2004); (4) rare earth elements-enriched livestock feeds and fertilizers applied to croplands and aquacultural systems (Wang et al. 2001; Redling 2006; He et al. 2010; Migaszewski and Gałuszka 2015a, b); (5) recycling plants for waste electronic and electrical equipment such as computer monitors, television screens, and fluorescent tubes (Resende and Morais 2010; Erfort et al. 2017); and (6) petroleum refineries where rare earth elements are used in catalysts (Kulaksız and Bau 2013). A review of the anthropogenic sources of rare earth elements is presented in an earlier paper (Gwenzi et al. 2018).

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Table 1.2 Nature and sources of rare earth elements reported in various aquatic systems Rare earth elements Eu

Gd

La

Heavier rare earth elements Lu

Ce

1.3.2

Source Fiber optics

Gadopentetic acid, Gd (DTPA)2
, which is used in magnetic resonance imaging Metallurgy

Crystal scintillators

Polishing powders and metallurgy

Aquatic system Sedimentary rocks, bed sediments, seawater rivers in the Chambal River (1.296 μg/g) and Deccan Trap basalts Oceans, river water, and seawater in Yamuna River at a concentration of 4.8 μg/g Oceans, river water, and seawater in Yamuna River (39.6 μg/g) Zircon and garnet in Yamuna River (14.6 μg/g) Sediments such as quartzites, zircon, monazite, and water in Kali Sindh (0.48 μg/g) and Chambal River (0.38 μg/g) Bed sediments in Yamuna and Chambal rivers and fresh water, granites in Hanuman, Chatti, and Deccan Trap basalts (76.6 μg/g)

References Allègre et al. (1996) and Kümmerer and Helmers (2000) Knappe et al. (1999) Knappe et al. (1999) Byrne and Kim (1990) Johannesson et al. (1995) and Aubert et al. (2001) Braun et al. (1990), Sholkovitz (1995) and Dalai et al. (2004)

Dissemination and Behavior of Rare Earth Elements in Aquatic Systems

Rare earth elements may be disseminated into aquatic systems as a coherent group via hydrological processes and anthropogenic activities (Weltje et al. 2002). Specifically, hydrologically driven processes such as infiltration, groundwater recharge, runoff, and erosion processes mobilize and transfer rare earth elements from hotspot sources into various environmental compartments (Cao et al. 2017). The release of untreated and partially treated effluents and wastewaters from rare earth element industries contributes to the occurrence of rare earth elements in aquatic systems. For example, excessive precipitation and overtopping of tailing dams transfer rare earth elements from catchment to surface aquatic systems through runoff (Khan et al. 2016). Rare earth elements may enter groundwater via three pathways (Keasler and Loveland 1982): (1) mobilization and transport of readily soluble rare earth elements by infiltration and recharge, (2) as rare earth element-contaminated water migrating through soil layers into groundwater, and (3) via surface–groundwater interactions. Some studies report that the loading of rare earth elements in aquatic systems

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decreases in the order: riverine systems followed by brackish and then seawater (Herrmann et al. 2016). This trend reflects the hydrochemical processes such as dilution and mixing of fresh and salty water in estuaries (Herrmann et al. 2016). In addition, coagulation and subsequent sedimentation of rare earth elements bound to particulate and colloidal materials such as organic matter and iron and manganese oxyhydroxides may also occur. Gadolinium originating from gadolinium-based contrast agents widely applied in magnetic resonace imaging has been detected in aquatic systems in several developed countries. These include Germany (Kulaksız and Bau 2013), Australia (Lawrence et al. 2009; Lawrence 2010), Switzerland (Vriens et al. 2017), Canada (Macmillan et al. 2017), the United Kingdom (Thomsen 2017), and the United States (Bau et al. 2006; Hatje et al. 2016). Gadolinium contrast agents have a low residence time in the human body; thus, they are rapidly excreted into wastewater conveyance and treatment systems and other aquatic systems. Gadolinium contrast agents are highly stable, and almost inert; thus, they are not effectively removed by most unit operations in wastewater treatment processes, hence the occurrence of gadolinium contrast agents in surface and groundwater systems (Knappe et al. 2005). Rare earth elements have also been reported in drinking water systems including tap water (Kulaksız and Bau 2011a, b; Lindner et al. 2015) and some aquatic organisms such as sea urchins (Merschel et al. 2015). In the Netherlands and Germany, anthropogenic rare earth elements detected in aquatic systems include samarium and lanthanum, which are used as catalytic cracking catalysts in petroleum refineries (Kulaksız and Bau 2013). In the same studies, gadolinium largely occurred in a dissolved form, while samarium and lanthanum were in nanoparticulate or colloidal forms depending on pH and the available ligands (Kulaksız and Bau 2011a, b, 2013). However, currently missing in existing literature are data pertaining to rare earth elements in aquatic systems in developing regions including Africa, Asia, and South America. Rare earth elements in aquatic systems may exhibit complex behaviors, which are controlled by several biogeochemical conditions. These conditions include type and speciation of the rare earth elements and geochemical conditions including solution pH, ionic strength, ligands, redox potential, natural organic matter content, and the presence of aquatic plants (Johannesson et al. 2004; Wilke et al. 2017). In aquatic environments, rare earth elements may undergo sorption–desorption processes, ion exchange, plant uptake and bioaccumulation, and liquid–solid phase partitioning among colloidal materials such as organic matter and minerals and pore and bulk water and plant uptake (Chakhmouradian and Wall 2012; Verplanck 2013; Klaver et al. 2014). The biogeochemistry of rare earth elements in aquatic systems, including solution chemistry, and complexation processes with both inorganic and organic ligands has been discussed in an earlier papers (Johannesson et al. 2004; Migaszewski and Gałuszka 2015a, b). Mineral phases including manganese and iron oxyhydroxides and organic matter exhibiting high surface areas behave as efficient scavengers for rare earth elements, thus controlling the adsorption–desorption reactions. The adsorption of rare earth elements on solid matrices may also occur, followed by sedimentation and resuspension through bioturbation and hydraulic drift. Rare earth elements may also undergo redox reactions, characterized

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by highest oxidation states in surface waters containing high concentrations of oxygen, while lower oxidation states may dominate under anoxic conditions in deepwater layers (Weltje et al. 2002). Compared to lighter rare earth elements, heavier rare earth elements have a tendency to form complexes with inorganic anions; thus, they remain in the solution, while chlorine complexes often mobilize the rare earth elements (Williams-Jones et al. 2012). Aquatic plants may take up rare earth elements, which may then undergo trophic bioaccumulation and biotransformation, phenomena similar to those occurring in soils (Li et al. 2013a, b; Lindner et al. 2013; Khan et al. 2017; Amyot et al. 2017). Plants tend to have a differential uptake of rare earth elements, characterized by a relatively higher affinity for trivalent lanthanides than divalent rare earth elements (Weltje et al. 2002). However, the effects of uptake and bioaccumulation of rare earth elements on plants appear to depend on the type and speciation of the rare earth elements and plant species. One study evaluated the ability of four aquatic plants (Elodea nuttallii, E. canadensis, Ceratophyllum demersum, Lemna gibba) to act as biofilters and take up two gadolinium-based contrast agents (i.e., Dotarem, Omniscan) from aqueous systems (Braun et al. 2018). Although the study showed no significant bioaccumulation of gadolinium-based contrast agents, tissue concentration reached at peak between days 1 and 4, before being released back into water (Braun et al. 2018). In the same study, uptake by the four plant species had negligible effect on the removal of gadolinium in water at concentrations investigated (i.e., 1–256 μg/ L). Accordingly, Braun and co-workers concluded that biofiltration by macrophyte species studied had limited capacity to remove gadolinium-based contrast agents in aquatic systems. Overall, the biogeochemical behavior and fate of rare earth elements in aquatic environments are governed by complex processes involving surface chemistry, solution complexation processes, and plant uptake, which in turn, depend on various environmental and biotic factors.

1.3.3

Environmental Health Risks

Ecological Health Risks Compared to studies documenting the occurrence of rare earth elements in aquatic systems, limited data exist on the effects of rare earth elements on aquatic ecology at species, population, community, and ecosystem levels. However, the few ecotoxicological laboratory studies and data largely drawn from other terrestrial ecosystems in China (e.g., agroecosystems) point to potential ecological effects (Liang et al. 2014; Zhuang et al. 2017a, b). Rare earth elements and their ecological effects have been reported in soils, soil organisms, and plants (Zhang and Shan 2001; Liang et al. 2014) and aquatic systems (Yuan et al. 2003; Trifuoggi et al. 2017). The dominance of literature drawn from China may reflect the country’s major role in the production and industrial applications of rare earth elements. The ecotoxicological effects of rare earth elements can be traced back to the late 1940s and 1970s (Burkes and

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Table 1.3 Summary ecological health risks of rare earth elements Rare earth elements Several rare earth elements La3+, Yb3 +

Yb3+ Ce4+

Gd Ce CeO2

Health risks Affects embryo development in zebra fish

Remarks Rare earth elements attach in place of Ca2+ and affect physiological functions regulating Ca2+ in zebra fish

Delayed larval development

La3+ and Yb3+ delayed larval development, reduced hatching and survival rates, and induced tail deformities in zebra fish Reduced activity of enzyme catalase

Induced mortality in goldfish Induces mortality and infertility of sea urchin embryo Neurotoxicity in rats Reduced life expectancy in fruit flies Contamination of food chain

Nd

Decreased growth in wheat and rice

Li

Necrosis of older leaves

La and Ce

Decreased growth in maize

La3+

Inhibits seed germination

N, Er, and Y

Reduced seed germination rate

Sea urchin experience mortality when exposed to Ce4+ and sperm development are also affected Rats’ exposure to different doses of Gd resulted in neuronal death Ce reduced the lifespan of Drosophila melanogaster at high exposure rates In its nanoparticle form, CeO2 accumulates in zucchini, crickets, and wolf spiders food chain. This results in trophic transfer and food chain contamination by CeO2 A decreased growth was observed in rice and wheat grown in Nd concentration of 10 and 25 mg/L of Nd Death of older leaves was observed in lettuce when tissue concentration above 1000 mg was detected Application of high concentrations of La and Ce reduced growth of maize relative to the control Pre-soaking of seeds in 1–10 mM of La3+ and rare earth element nitrate solution for 4 h inhibited seed germination in plants Higher concentrations reduced seed germination of R. sativus and S. lycopersicum

References Tang and Johannesson (2003) and Cui et al. (2012) Cui et al. (2012)

Oral et al. (2010), Martino et al. (2017) Xia et al. (2011) Huang (2011) Hawthorne et al. (2014)

Basu et al. (2016)

D’Aquino et al. (2009)

Carpenter et al. (2015); Thomas et al. (2014)

McCleskey 1947; Gale 1975). The bacteriostatic effects of cerium and lanthanum were first reported as early as 1947 by Burkes and McCleskey (1947). Since then, evidence showing the adverse effects of rare earth elements in both terrestrial and aquatic organisms has been increasing (Table 1.3). High concentrations of rare earth elements in aquatic systems may have adverse ecological effects (Table 1.3). Rare earth elements have been reported to inhibit plant uptake of essential nutrients such as calcium, an effect attributed to the fact that

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calcium cationic radius and those of rare earth elements are similar (Hu et al. 2003). In some plant species, rare earth elements also alter the channels associated with the endoplasmic reticulum responsible for calcium release (Klüsener et al. 1995). In turn, the reduced intake of calcium may interfere with the biochemical functions of calcium in plants, including formation of cell walls, root growth, photosynthesis, and flowering. Cerium has similar physical properties as calcium; hence, it is easily absorbed by plants and is more phytotoxic as compared to the rest of rare earth elements. According to Burda et al. (1995), cerium competes with calcium for the same binding sites in P680 or photosystem II, one of the photosynthetic reaction centers. In the symplast of roots, ions of rare earth elements disrupt ionic channels responsible for xylem exudation (Schwenke and Wagner 1992). Lanthanum reversibly inhibits the phototaxis and photophobic response in some green and blue–green algae via blockage of the calcium pump and reduction of conductivity (Herrmann et al. 2016). Lanthanum also reduces calcium and magnesium concentrations in plants, which in turn reduce the concentrations of carotenoids and chlorophylls a and b while enhancing the activities of antioxidant enzymes (Xu et al. 2012). In addition, at a concentration of 11.1 mg/L, lanthanum is reported to induce alterations of the ultrastructure of cell organelles such as mitochondria, chloroplasts, and nucleus (Xu et al. 2012). The same studies attributed the reduction in chlorophyll synthesis to lanthanum inhibiting the functions of magnesium and calcium. In algae (Chlorella pyrenoidosa), which is a common aquatic plant, high concentrations of rare earth elements inhibited both growth and reproduction (Hu et al. 2003). In the same study, the toxicity decreased in the order: neodymium followed by praseodymium, cerium, and lanthanum and then the mixtures of the four rare earth elements. The uptake and bioaccumulation of rare earth elements have documented several aquatic plant species such as water hyacinth, a process also common with metals (Singh and Kalamdhad 2012; Zheng et al. 2016). Other mechanisms accounting for the toxicity of rare earth elements in plants include generation of reactive oxygen species causing oxidative stress and alteration of photosystem II (Pang et al. 2002; Xia et al. 2011). In some plant species, cerium, lanthanum, and praseodymium concentrations exceeding 50 mg L
1 inhibit photosynthesis via reduction of the activity of photosystems (Pang et al. 2002). The toxicity of rare earth elements may also depend on geochemical conditions such as redox potential, pH, ionic strength, and the prevailing ligands. A study conducted by Thomas et al. (2014) on seed germination showed that at low pH, cerium had harmful effects on A. syriaca, P. virgatum, R. sativus, and S. lycopersicum, while Y appeared to have phytotoxic effects on D. canadense and S. lycopersicum at higher doses. However, at high pH, lanthanum and cerium had no significant effect on the percentage germination at all doses, but lanthanum reduced plant biomass relative to the control (Thomas et al. 2014). In the same study, all five plant species that were studied showed that Ce accumulated in both plant shoot and root, with the roots accounting for the most accumulation. Oxygen evolution rate in seeds treated with rare earth elements was higher than in non-treated seeds, indicating greater metabolism. Application of low rates improved root growth and fresh weight in Eriobotrya japonica; thus, low rates were found to

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increase the chlorophyll content and vice versa. Application of lanthanum at high rates was found to have an inhibitory effect. Yb3+ and Eu3+ inhibit shoot growth and even cause death at high concentrations. Lanthanum was observed to indirectly reduce algal growth by forming an insoluble phosphate precipitate, which reduced plant uptake and growth due to shortage of phosphate (Stauber and Binet 2000). Jin et al. reported that a lanthanum concentration of 7.2 μmol L
1 significantly reduced the growth of cyanobacteria and microalgae. One beneficial effect of rare earth elements to plants is that rare earth elements can inhibit bioavailability and uptake of toxic metals possibly via competition. For example, lanthanum inhibits uptake of Pb2+ in plants grown in lead-polluted areas. However, the ecological effects of rare earth elements may also depend on other stress factors, plant species, and growth stage. Rare earth elements may also adversely affect aquatic animals including fish and crustaceans, similar to the ecological effects reported in terrestrial animal bioassay species (e.g., rats, mice (Pagano et al. 2012, 2015). In one study conducted on water fleas (Daphnia carinata), lanthanum delayed maturation (Barry and Meehan 2000). Another study reported lack of antagonistic and synergistic effects in fish (carp) exposed to rare earth elements (Qiang et al. 1994). High concentrations of gadolinium in aquatic systems were positively correlated with increased mortality of the daphnids (Perrat et al. 2015). The inhalation of cerium dioxide (CeO2) nanoparticles obtained from diesel fuel catalysts promoted IFN-γ and IL-12 production by alveolar macrophages (Ma et al. 2014). Although data pertaining to aquatic animals is limited, the age dependence of the toxicity of rare earth elements has been reported in terrestrial animals. For example, in one study, 200 or 500 mg cerium trichloride (CeCl3)/kg body weight was administered via gavage for adults, milk for neonatal mice, and placenta transfusion for fetal mice (Kawagoe et al. 2008). On the one hand, multiple toxicity effects manifested in adult mice, including neutrophil infiltrations, necrosis, pulmonary venous congestion, pulmonary hemorrhage, thickened alveolar septae, and hepatic. On the other hand, toxicity in fetal mice occurred in the form of hepatic and pulmonary congestion (Kawagoe et al. 2008). The same authors reported that cerium trichloride (CeCl3) administered via gavage resulted in damage of the liver and respiratory system. These effects were more severe in adult mice that in neonatal and fetal mice. In some studies, 40 mg LaCl3/kg body weight administered to rats via gavage caused behavioral changes and promoted enzyme activity of Ca2+ ATPase in the hippocampal cells (He et al. 2008). On the contrary, significant decreases were noted in the activities of catalases, glutathione peroxidase, and superoxide dismutase relative to the control (Feng et al. 2006). Rare earth elements are highly redox reactive; thus, they may induce oxidative stress that damages biomolecules including deoxyribonucleic acid. As reported in plants, oxidative stress has also been reported in bioassay animals subjected to gavage administration of La3+, Ce3+, and Nd3+ (Huang 2011). In ecotoxicological studies, oxidative stress is evaluated using endpoints such as lipid peroxidation, concentrations of reactive oxygen species, and changes in antioxidant activities of associated enzymes such as superoxide dismutase, glutathione peroxidase, and

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catalases (Liang and Wang 2014; Pagano et al. 2015). Some recent studies report that gadolinium originating from gadolinium-based contrast agents is toxic to marine organisms including bivalves and sea urchins (Perrat et al. 2017; Martino et al. 2018; Rogowska et al. 2018). Some ecotoxicological studies show that rare earth elements are characterized by biphasic or hermetic dose–response relationships (Pagano et al. 2015). Such dose–response relationships show beneficial or stimulatory effects at low doses, while toxic or inhibitory effects are observed at high doses (Pagano et al. 2015). This biphasic or hormetic behavior forms the basis for the applications of rare earth elements as growth promoters in plants, aquaculture, and livestock production (Wang et al. 2000). Accordingly, low doses of rare earth elements are used in livestock feeds and fertilizers as substitutes for zinc and copper in the pretext that the use of rare earth elements reduces the health risks of zinc and copper (Redling 2006). At ecosystem level, studies show that high concentrations of rare earth elements alter ecological functions and reduce biodiversity of insects of the order Dermaptera and family Carabidae (Li et al. 2010). However, some species (e.g., Stibaropus formosanus) and genera (e.g., Formicidae) are less sensitive (Li et al. 2010), suggesting that the ecotoxicological effects are species- and genera-dependent. Rare earth elements potentially alter biogeochemical cycling in aquatic systems, but limited data exist on this aspect. Liu and Wang (2001) showed that soil application of rare earth elements at a rate of 5 mg/kg significantly decreased the concentration of available nitrogen, possibly by inhibiting mineralization of nitrogen. This could in turn induce nitrogen deficiency and alter ecosystem function by reducing primary productivity in aquatic systems. Rare earth elements may also pose a radioactivity hazard to aquatic organisms (Akiwumi and D’Angelo 2017). In summary, rare earth elements may cause diverse ecological effects, which depend on speciation, recipient organism and intake route, environmental conditions, nature, and concentration of rare earth elements in aquatic systems.

Human Exposure and Health Risks Human intake of rare earth elements occurs via medical applications in magnetic resonace imaging and administration of rare earth element-based pharmaceuticals, occupational exposure, and dietary intake (Kanda et al. 2017; Zhuang et al. 2017a, b; Khan et al. 2017). A number of studies have reported rare earth elements and their health risks in the human body including the brain, hair, nails, serum, milk, and sperms (Gomez-Aracena et al. 2006; Wei et al. 2013; Poniedziałek et al. 2017). For example, four rare earth elements (praseodymium, erbium, neodymium, lanthanum) were observed in human milk sampled from women in a hospital in Poznań, Poland (Poniedziałek et al. 2017). Occupational exposure to rare earth elements may occur via inhalation during mining, refining, and industrial production of rare earth elements-based products (Rim et al. 2013; Gambogi 2016). Once in the human body, rare earth elements pose several human health risks (Table 1.4). Human health risks associated with occupational exposure to rare earth elements include

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Table 1.4 Summary of human health risks associated with rare earth elements Rare earth elements Eu, Dy, Pr

Health risks Breast cancer

Gd, La

Nephrogenic systematic fibrosis

Ce

Myocardial infarction Genotoxicity

Several rare earth elements Several rare earth elements Several rare earth elements Several rare earth elements Several rare earth elements

Pneumoconiosis

Several rare earth elements

Failure for blood to clot

Sterility

Arteriosclerosis

Memory loss

Remarks Rare earth elements were detected in neoplastic cells Gd and La damage the structure of the renal tubule. Persistence of Gd in the body resulted in the release of toxic Gd from gadodiamide transmetalation, which triggers nephrogenic systemic fibrosis Elevated Ce concentrations in toe nails increase the risk of myocardial infarction Rare earth elements accumulate in the bone structure, changes the bone texture, and enhances bone marrow micronucleus Rare earth elements cause damage to the cells of the lungs, resulting in pneumoconiosis Accumulate in serum and decrease percentage of normal sperm cells Rare earth elements in blood raise cholesterol levels and production of a lot of lipoprotein Rare earth element accumulation in brain tissues lowers intelligence quotient levels. Rare earth elements also cause dysfunctional neurological behavior, leading to impaired learning ability Rare earth elements lower the total blood bilirubin, glucose, and albumin

References Roncati et al. (2018) Thomsen (2006)

Chen and Zhu (2008) and Zaichick et al. (2011) Pagano et al. (2015)

MarzecWróblewska et al. (2015) Migaszewski and Gałuszka (2015a, b) Zhu et al. (1996) and Zhuang et al. (2017a, b)

pneumoconiosis and interstitial lung disease which have been reported to be prevalent among movie projectionists and photogravers. A detailed database of the case reports of the human health risks associated with occupation exposure is presented in an earlier study (Pagano et al. 2015). A study applying a scanning electron microscopy coupled to energy-dispersive X-ray detected cerium particles in the lungs of polishers for optical lens (Yoon et al. 2005). Dietary intake of contaminated drinking water and foods also contribute to human intake of Rare earth elements. In China, two studies investigated the occurrence of rare earth elements in vegetable and cereal crops and the human health risks of rare earth elements in mining relative to non-mining (control) areas (Zhuang et al. 2017a, b). In the same studies, the total concentrations of rare earth elements in cereals were 74.22 μg/kg for mining area and 47.83 μg/kg for the control (Zhuang et al. 2017a), while the corresponding concentrations in vegetables were on average 94.08 μg/kg (mining area) versus 38.67 μg/kg (control) (Zhuang et al. 2017b). In

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terms of total rare earth elements, this resulted in maximum permissible human daily intakes of approximately 70 μg/kg body weight per day for both vegetables and cereal crops. The same authors hinted that human health risks could be particularly high in children due to continuous exposure. In another study, the total concentrations of rare earth elements in taro and water spinach treated with rare earth elements exceeded Chinese food standards (Li et al. 2013a, b). The same authors noted that the concentrations of rare earth elements were significantly lower in a non-leafy vegetable (taro) than a leafy one (water spinach). Consequently, the concentrations of rare earth elements detected in the human hair and blood were significantly higher in areas contaminated with rare earth elements (424.76–1294.8 μg/L) than in the corresponding samples from the control (Li et al. 2013a, b). This study shows that continuous application of rare earth elements to vegetables may enhance uptake and accumulation in edible parts and transfer of rare earth elements into the human food chain. Hutchison and Albaaj estimated that acceptable rare earth element nitrate concentration in humans range from 0.2 to 2 mg/kg, beyond which adverse human health effects may occur. The human health risks of rare earth elements via dietary intake depend on the concentrations of rare earth elements in food and age and body weight of individuals. Although aquatic systems are sources of human food and water, limited data exists linking human health conditions to rare earth elements in aquatic systems. Thus, further research is required to investigate this aspect. Therefore, the bulk of literature on human health risks of rare earth elements is drawn from literature on medical applications and applications of rare earth elements-enriched fertilizers in crop production systems. Once in the human body, rare earth elements may induce a wide range of human health effects (Table 1.4). Nephrogenic systemic fibrosis is the most severe human health risk attributed to gadolinium-derived gadolinium-based contrast agents used in medical applications (Broome 2008; Thomsen 2017). Gadolinium applications in magnetic resonace imaging and the associated human health effects are among the most studied rare earth elements of anthropogenic sources. Gadolinium-linked incidences of nephrogenic systemic fibrosis, including gadolinium deposition in the human body, and immediate toxicity during pregnancy and lactation and global best practices to minimize the human health risks have been discussed in an earlier review (Fraum et al. (2017). Gadolinium from gadolinium-based contrast agents may migrate across the blood–brain barrier, resulting in gadolinium accumulation in the brain, where it may cause severe damage to the nephrological system (Kanda et al. 2017; Vergauwen et al. 2018). The speciation, partitioning, and human health risks of gadolinium derived from gadolinium-based contrast agents in the brain are discussed in an earlier review (Kanda et al. 2017). Other human health risks such as genotoxicity, fibrotic tissue injury, and bone alteration attributable to rare earth elements have also been reported in literature (Chen and Zhu 2008; Jenkins et al. 2011). Lanthanum causes neurological disorders and reduced the intelligence quotient particularly in infants (Zhu et al. 1996; Gwenzi et al. 2018), while cerium induces pneumoconiosis (Porru et al. 2000).

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In males, rare earth elements have been reported to contribute to male sterility and anti-testicular effects (Chen et al. 2015; Marzec-Wróblewska et al. 2015). Zhu et al. (1996) report that high concentrations of rare earth elements in the human blood system promote the formation of arteriosclerosis via immunogenic damage to the vascular wall. Accumulation of rare earth elements in the human body alters blood properties such as serum triglyceride, β globulin, and albumin while decreasing glutamic pyruvic transaminase and total protein and increasing levels of cholesterol (Khan et al. 2017). However, some studies reporting concentrations of rare earth elements exceeding background values in human tissues due to environmental and dietary exposures failed to detect any corresponding apparent abnormalities in human health conditions (Li et al. 2013a, b; Zhuang et al. 2017a, b). As Pagano et al. (2015) indicated, case–control epidemiological evidence derived occupational exposure to rare earth elements is scanty. Therefore, the determination of maximum permissible concentrations of rare earth elements in food and environmental media, and the corresponding acceptable human intake values of rare earth elements, and chronic effects warrant further research.

1.4 1.4.1

Assessment and Mitigation of Health Risks Assessment of Human and Ecological Health Risks

Health risk assessment is a critical step for evaluating human and ecological health risks of rare earth elements. According to Peduzzi et al. (2009) and Cardona et al. (2012), the notion of health risk entails three aspects: (1) the existence of a potential health threat or hazard, derived from the toxicity of rare earth elements; (2) risk of exposure to the hazard, including the exposure routes and daily intake, which are related to the occurrence and concentrations of anthropogenic rare earth elements in the environment; and (3) vulnerability, which is indicative of the predisposition or propensity of human and ecological populations to a hazard or harm, and other adverse health outcomes such as morbidity and mortality (Gwenzi and Chaukura 2018). These aspects underpin the approach for evaluating the health risks of rare earth elements in aquatic environments (Gwenzi et al. 2018). Health risk assessments include risk analysis, involving the identification and determination of concentrations of rare earth elements in aquatic media such as sediments, pore water, aquatic organisms, and surficial and deep bulk waters (Gwenzi and Chaukura 2018; Gwenzi et al. 2018). Subsequently, the evaluation of health risk may involve qualitative and quantitative analysis, including ecotoxicological studies, modelling, and rating to develop a priority list of rare earth elements that warrant mitigation (Milić et al. 2012). Qualitative risk evaluation or ranking entails the estimation of the probability, likelihood or frequency of occurrence of the event, and the magnitude of the harm or consequences on ecological or human health (Gwenzi and Chaukura 2018). Qualitative ranking and classification of the health risks use categories such as: “low”, “moderate”, “high”, and “extremely high” risk (Gwenzi and Chaukura

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2018; Gwenzi et al. 2018). Such ranking facilitates the development of a priority list, which then assist in identifying health risks that warrant mitigation and target resources. Quantitative risk assessment involves modelling and determination of quantitative metrics for determining the human health risks, which are often probabilistic results. However, in the case of rare earth elements, and other emerging contaminants, quantitative risk assessments are constrained by lack of maximum permissible guideline limits for both humans and the environment. Detailed discussions of health risk assessment procedures for various contaminants including emerging ones are presented in earlier papers and environmental guidelines (Organisation for Economic Co-operation and Development 2003, 2007; Gwenzi and Chaukura 2017; Gwenzi et al. 2018; Sanganyado and Gwenzi 2019). Indeed, several health risk assessment protocols exist at country and regional levels. These protocols include the US Environmental Protection Agency (2017) environmental risk assessment and Organisation for Economic Co-operation and Development (2015) guidelines widely used in the United States and the European community, respectively. These environmental risk assessment protocols apply known sensitive bioassay organisms to determine the dose–response relationships, including ecotoxicological threshold values (Organisation for Economic Co-operation and Development 2015). However, such protocols should be extended to include mixtures of rare earth elements and potential synergistic interactions among rare earth elements and other health stressors using realistic and environmentally relevant concentrations. Such risk assessment should consider daily intakes, multiple exposure pathways, nature and speciation of the rare earth elements, and age and nature of the target organisms exposed to rare earth elements.

1.4.2

Prevention and Control of Health Risks

Given that aquatic systems provide ecological services, and act as sources of human food and water, the remediation of rare earth elements in aquatic systems is a high priority. Potential preventative and control measures may be grouped into three classes: (1) “soft” engineering interventions, (2) “hard” engineering interventions, and (3) strategies aimed at minimizing the demand for raw rare earth elements. “Soft” engineering interventions include raising awareness on the human and ecological health risks of rare earth elements associated with their industrial applications and subsequent disposal. Such awareness may include educational campaigns, policy briefings, and mass media programs targeting policy makers, environmentalists, researchers, and the general public. Practicing proper housekeeping including implementing proper occupational hygiene and safety, health, and environmental programs in the rare earth element supply chain system may minimize waste and wastewater discharges and anthropogenic releases of rare earth elements into aquatic systems. At industrial plant level, principles of cleaner production can also be

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adopted to recover rare earth elements from industrial waste and wastewaters, thus attaining a closed-loop system and zero waste discharge scenario. “Hard” engineering interventions include the development and applications of synthetic materials in engineered systems designed to remove rare earth elements. Such “hard engineering technologies include those based on chemical, biological, biochemical, and biosorption processes and (3) physicochemical processes. Among these technologies, the most prominent ones are (1) adsorption using natural or synthetic adsorbents (Tay et al. 2018; Zhang et al. 2018; Barros et al. 2019; Madbouly et al. 2019); (2) ion exchange using natural and synthetic ion-exchange resins (Wang 2018); (3) biosorption and biorecovery using live organisms such as fungi, bacteria, and microalgae (Kang et al. 2019; Furuhashi et al. 2019) and reverse osmosis (Lawrence et al. 2010); and (4) solid–liquid (Kumar et al. 2010, 2011) and liquid–liquid extraction using ionic liquids (Atanassova et al. 2018; Hunter et al. 2018; Habib et al. 2019). The highlighted technologies have some limitations, including sludge production, some requiring large amounts of chemical reagents, membrane fouling, formation of secondary pollutants, and challenges in separation of adsorbents from aqueous solution. Due to low cost, environmental friendliness, simplicity of application, and readily available feedstocks for the development of adsorbents, adsorption is the most dominant and widely used process for removal of rare earth elements in aqueous systems (Fiket et al. 2018). However, one drawback of adsorption is that it cannot be used for many cycles due to the degradation of the adsorbent structure by the stripping solutions. To overcome this limitation, immobilization on silica and polymers has been used to render adsorbents more efficient and reusable (Gupta et al. 2019). The capacity of several methods to remove rare earth elements occurring in aqueous systems, including the mechanisms and the drawbacks involved, is discussed in detail in the respective papers (Atanassova et al. 2018; Furuhashi et al. 2019; Hunter et al. 2018; Kang et al. 2019; Tay et al. 2018; Wang 2018; Zhang et al. 2018; Madbouly et al. 2019). For example, Lawrence et al. (2010) reported that reverse osmosis removed more than 99% of gadolinium in aqueous systems. Moreover, to attain even higher removal efficiencies, several methods may be combined in tandem in a train consisting of various unit operations (Carolin et al. 2017). In addition, extractive and separation processes are the methods which may be adapted to recover rare earth elements from ores in aquatic systems. These methods include precipitation and selective leaching targeting specific rare earth elements (Hidayah and Abidin 2017; Innocenzi et al. 2018), solid–solid extraction techniques (Hidayah and Abidin 2017), the use of supercritical fluids (e.g., carbon dioxide) (Sinclair et al. 2017), membrane technologies (Liu et al. 2018), plasma separation methods (Gueroult et al. 2018), and ultrasonic extraction techniques (Diehl et al. 2018). However, limited data exists on the application of membrane filtration, electrochemistry, oxidation, photocatalysis, and adsorption using emerging biomaterials such as biochars and biochar–metal oxide composites. The environmental footprint of the mining, processing, and industrial applications of rare earth elements could also be reduced via various interventions (Binnemans et al. 2013a, b). Specifically, the overreliance on rare earth elements

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for high-technology applications can be minimized by reducing the amount used in high-technology devices and equipment. Future demand for raw rare earth elements can also be reduced by exploring new and emerging industrial applications for rare earth elements which are most readily available and cheap to recover relative to less abundant ones. For example, the high concentrations of the costly dysprosium used in neodymium–iron–boron lightweight permanent magnets may be significantly minimized by using samarium–cobalt magnets, whose performance is comparable to that of the neodymium–iron–boron magnets. Another option is transmaterialization or substitution of rare earth elements with other materials, coupled with the elimination of rare earth elements in products by developing alternative non-rare earth element products. In this regard, transmaterialization may entail the use of carbon nanomaterials derived from abundant carbon, thus reducing the amount of rare earth elements required in the development of hightechnology devices and applications (Arvidsson and Sandén 2017). The demand for raw rare earth elements can also be reduced via recovering, recycling, and reuse of rare earth elements occurring in post-consumer high-technology products and nonconventional sources of rare earth elements including e-wastes and mine wastes such as coal ashes (Dent 2012; Tan et al. 2015). For rare earth elements in motors and generators, an option exists to substitute machines relying on permanent magnet with those based on coil-wound induction. In catalysis, scope exists to develop non-rare earth element catalysts, although such options may require considerable time, financial resources, and research effort (Dent 2012). These interventions highlighted here can be used as part of a conceptual framework for the mitigation of health risks of rare earth elements. Overall, a combination of “soft” and “hard” engineering interventions, coupled with minimizing the demand for raw rare earth elements via the various options highlighted, may reduce the global production and applications of rare earth elements and the associated health risks.

1.5

Future Research Directions

Specific knowledge gaps are highlighted under the following focal areas: (1) source partitioning and environmental behavior, (2) ecotoxicology, (3) human toxicology and epidemiology, and (4) human exposure and health risks in developing regions.

1.5.1

Source Partitioning and Behavior in Aquatic Systems

Literature on rare earth elements in aquatic systems is dominated by only a few elements (e.g., lanthanum, cerium, europium, gadolinium, lutetium), thus monitoring of the other understudied rare earth elements is required. Although several anthropogenic hotspot sources may contribute to rare earth elements detected in

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aquatic systems, limited data exists on source partitioning of rare earth elements among various sources. Moreover, barring studies documenting the concentrations of rare earth elements in aquatic systems, few studies have investigated the following: (1) the dominant hydrological and wind-driven processes accounting for the dissemination of rare earth elements in the environment; (2) the relative contribution of biogeochemical processes including sorption–desorption, complexation, ion exchange, precipitation, and plant uptake and bioaccumulation on the partitioning of rare earth elements in bulk water, pore water, the solid phase, and aquatic organism; and (3) trophic bioaccumulation in aquatic food webs including aquatic foods such as fish and crustaceans.

1.5.2

Ecotoxicology

Existing ecotoxicological data on rare earth elements are drawn from studies investigating the effects of single elements on a target bioassay species, while the synergistic interactions among rare earth elements, and between rare earth elements and other health stressors, remain understudied. Future research should investigate the effects of rare earth elements and the interactions between rare earth elements and other stressors on ecosystem services (e.g., biogeochemical cycling) and population, community, trophic, and ecosystem diversity and functions. Studies are also needed to estimate the ecotoxicological threshold points for various environmental media (e.g., median effect concentration, no observable effect concentration). To facilitate the determination of maximum allowable guideline limits, such studies should be conducted at concentrations considered to be environmentally relevant.

1.5.3

Human Toxicology and Epidemiology

Evidence exists on the human toxicology and health effects of rare earth elements in medical applications particularly gadolinium contrast agents used in magnetic resonace imaging (Thomsen 2006), while other studies conducted in China have reported rare earth elements in scalp hair sampled from children in rare earth element mining sites (e.g., Tong et al. 2004). However, further research is required to better understand the human exposure routes and daily intakes and behavior and fate of rare earth elements once in the human body. Moreover, further epidemiological research is required to establish the relationship between rare earth elements detected in aquatic environments and human health effects such as incidences of morbidity and mortality.

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Human Exposure and Health Risks in Developing Countries

Developing regions, including Africa, are poorly represented in literature on anthropogenic rare earth elements in the environment including aquatic systems. Yet several developing countries, including Kenya, the Dominican Republic, and the Democratic Republic of Congo, among others have large reserves of rare earth elements. Moreover, developing countries also have anthropogenic hotspot sources of rare earth elements including medical facilities, waste repositories, and petroleum refineries. Other sources include recycling plants for waste electrical and electronic equipment and mining and mineral processing industries. In these regions, the bulk of monitoring data in aquatic systems excludes rare earth elements and is often limited to conventional toxic contaminants including toxic metals, pesticides, geogenic arsenic, and fluoride. As highlighted in one of the author’s earlier papers (Gwenzi and Chaukura 2017; Gwenzi et al. 2018), the potential health risks of rare earth elements could be higher in these regions as in developed countries due to several risk factors. These risk factors include (1) weak and poorly enforced human health and environmental regulations, causing the discharge of partially treated and untreated effluents and wastewaters into aquatic environments; (2) the continued use of medical devices, equipment, and reagents long banned elsewhere in developed countries; (3) lack of regular human health surveillance systems for early detection and treatment of human health effects; and (4) unavailability of clean treated drinking water, thus forcing people to rely on potentially contaminated shallow groundwater and surface water sources. Therefore, comprehensive studies are required to investigate the occurrence of rare earth elements and their health risks in these regions. In addition, developing countries provide ideal sites to investigate the capacity of conventional low-cost water treatment (sand filtration, boiling) and emerging technologies such as zero-valent iron (Fe0) and biochar water filters (Gwenzi et al. 2018) to remove rare earth elements in drinking water.

1.6

Summary, Conclusions, and Future Directions

This chapter summarizes the key sources, environmental behavior, human intake pathways, and health risks of anthropogenic rare earth elements as emerging contaminants. Anthropogenic rare earth elements (i.e., lanthanum, cerium, europium, gadolinium, lutetium) are widely reported in aquatic environments in several countries. Dominant hotspot sources of rare earth elements include mining and processing of rare earth elements, medical facilities, petroleum refineries, recycling industries for waste electrical and electronic equipment, industries producing hightechnology products, pharmaceutical industries, and rare earth elements-enriched livestock feeds and fertilizers. The mobilization and transport of rare earth elements from sources and reservoirs occur via anthropogenic activities and hydrologically

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and wind-driven processes. Human exposure occurs via occupational inhalation, intake during medical applications, and consumption of contaminated water and foods. Evidence exists demonstrating the chronic and acute adverse effects of some rare earth elements in humans and aquatic organisms. Rare earth elements occur in human hair, nails, and bio-fluids (i.e., milk and serum), pointing to human intake and potential human health risks. Gadolinium contrast agents used in magnetic resonace imaging may cross the human brain–blood barrier and induce severe damage to nephrological systems and nephrogenic systemic fibrosis. In humans, rare earth elements also reduce intelligence quotient particularly in infants and cause dysfunctional neurological disorders, pneumoconiosis, fibrotic tissue injury, cytotoxicity, and oxidative stress. In males, rare earth elements induce anti-testicular health effects and sterility. However, barring gadolinium in gadolinium contrast agents used in magnetic resonace imaging, epidemiological evidence relating anthropogenic rare earth elements detected in aquatic systems to adverse human health effects is still weak. In plants, rare earth elements decrease both root function and growth and nutritional quality, while genotoxicity, oxidative stress, and neurotoxicity may occur in aquatic animals. Rare earth elements may bioaccumulate in aquatic organisms along trophic levels and cause chronic and acute toxicities and alter ecological functions. Although data drawn from developing countries including Africa are still missing, the human health and ecological risks could be higher in these regions than currently perceived. The risk factors predisposing human and ecological health in developing countries relative to their counterparts in developed countries were discussed. A conceptual outline for assessing and mitigating health risks of rare earth elements was highlighted. Specific remediation techniques for the removal of rare earth elements in aqueous systems were also discussed. Future research directions were highlighted to better understand the following: (1) hotspot reservoirs; (2) behavior and fate of rare earth elements in aquatic systems, including solid–liquid phase partitioning; (3) human toxicology and epidemiology and aquatic ecotoxicology; and (4) remediation of rare earth elements in aqueous systems, including drinking water and wastewaters. An improved understanding of the sources, behavior, and health risks of anthropogenic rare earth elements could contribute toward the development of environmental maximum guideline limits. Acknowledgments WG gratefully acknowledges funding from the International Foundation for Science (IFS), Sweden [grant number C/5266-2] for supporting the project entitled “Development and Application of a Biochar-Based Technology for Decentralised Treatment of Contaminated Drinking Water among Poor Communities in Developing Countries.” The project provided the framework and motivated this book chapter. The authors have sole responsibility over research and the decision to publish this chapter.

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Author Contributions WG originated, conceptualized, finalized, and edited the manuscript; NMM, MM, and AAM contributed equally to data collection and review of literature and drafted the manuscript.

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Chapter 2

Removal of Heavy Metal Pollutants from Wastewater Using Zerovalent Iron Nanoparticles Ambavaram Vijaya Bhaskar Reddy and Gajulapalle Madhavi

, Muhammad Moniruzzaman

,

Contents 2.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2 Presence and Adverse Effects of Heavy Metals in Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3 Zerovalent Iron Nanoparticles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3.1 Synthesis of Zerovalent Iron Nanoparticles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3.2 Characterization of Zerovalent Iron Nanoparticles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3.3 Potential Applications of Zerovalent Iron Nanoparticles . . . . . . . . . . . . . . . . . . . . . . . . . . 2.4 Removal of Selected Heavy Metals Using Zerovalent Iron Nanoparticles . . . . . . . . . . . . . . . 2.4.1 Removal of Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.4.2 Removal of Cadmium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.4.3 Removal of Arsenic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.4.4 Removal of Chromium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.4.5 Removal of Lead . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.5 Toxicity and Risk Assessment of Zerovalent Iron Nanoparticles . . . . . . . . . . . . . . . . . . . . . . . . . 2.6 Conclusions and Future Perspectives . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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A. Vijaya Bhaskar Reddy Centre of Research in Ionic Liquids, Universiti Teknologi PETRONAS, Seri Iskandar, Perak, Malaysia Quality Control División, Ultra International Limited, Ghaziabad, Uttar Pradesh, India M. Moniruzzaman (*) Centre of Research in Ionic Liquids, Universiti Teknologi PETRONAS, Seri Iskandar, Perak, Malaysia Department of Chemical Engineering, Universiti Teknologi PETRONAS, Seri Iskandar, Perak, Malaysia e-mail: [email protected] G. Madhavi Department of Chemistry, Sri Venkateswara University, Tirupati, Andhra Pradesh, India © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature Switzerland AG 2021 Inamuddin et al. (eds.), Water Pollution and Remediation: Heavy Metals, Environmental Chemistry for a Sustainable World 53, https://doi.org/10.1007/978-3-030-52421-0_2

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Abstract Heavy metals are high atomic weight elements that consist five times higher density than water. The widespread use of heavy metals in scientific, agricultural, domestic, industrial, and medical applications accelerated their distribution into water bodies through the environment. The extreme toxicity of heavy metals and their adverse effects on human health have raised concerns for the removal of heavy metals from different water bodies. Considering the severe toxicity, mercury, cadmium, chromium, arsenic, and lead were identified as priority heavy metal pollutants and categorized as human carcinogens by United States Environmental Protection Agency. Over the past few years, zerovalent iron nanoparticles have emerged as potential alternatives for the removal of heavy metals from water and wastewater streams. The superior reactivity and large surface area of zerovalent iron nanoparticles provided greater versatility for the in situ remediation of heavy metals. Therefore, this chapter presents a detailed discussion on the advances reported for the heavy metal remediation using zerovalent iron nanoparticles. It begins with the fate and transport of heavy metals into water bodies and their impact on human health and environment. Additionally, preparation methods, characterization techniques, and inherent applications of zerovalent iron nanoparticles toward the removal of heavy metals from different water bodies are extensively described following the risk assessment studies. Finally, concluding remarks and future prospects that support the effective remediation of heavy metals using zerovalent iron nanoparticles are summarized. Keywords Wastewater · Heavy metal pollution · Remediation approaches · Risk assessment · Toxicity · Human health · Zerovalent iron

2.1

Introduction

The accessibility of potable water is indeed necessary for all human beings throughout the world for their basic needs together with the drinking purpose. However, majority of the water bodies throughout the world are being contaminated with a variety of heavy metals and hazardous chemicals due to the rapid industrialization, globalization, and increased population. Among the variety of pollutants that are commonly popping up in the global water bodies, the appearance of heavy metals has become a serious threat owing to their extreme toxicity and accumulation tendency in food chains. When these heavy metals consumed by the humans through drinking water, they produce serious health issues like cancers, nervous system diseases, organ damage, and death in rare instances (Joseph et al. 2019). These heavy metals generally possess the specific density of higher than 5.0 g/cm3 and seriously influence the ecology and environmental systems. Also, majority of the heavy metals are highly water soluble and act as carcinogens to the fauna and flora of the receiving water bodies (Salawu et al. 2018).

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The prioritized heavy metals including chromium, mercury, cadmium, zinc, arsenic, lead, nickel, copper, vanadium, and titanium occur from the electroplating, milling and etching, anodizing cleaning, and electrolysis deposition processes. Also, the printed circuit board manufacturing industries release a significant amount of heavy metal wastes including tin, lead, and nickel into water bodies. Further, the wood processing industries generate arsenic waste, and inorganic pigment manufacturing industries produce chromium and cadmium sulfide wastes that probably merge into the water bodies (Sharma et al. 2019). In addition, several humaninduced activities including textile industries, petroleum refining, pesticides usage, sewage runoff, battery and paint manufacturers, and printing and photographic industries release considerable amounts of heavy metal waste into the vicinity of water bodies (Sarah et al. 2019). All the above named industries generate enormous amount of wastewater and sludge that contains heavy metal pollutants. The United States Environmental Protection Agency has provided regulations declaring the permitted heavy metals concentration in various drinking water sources (Maurya et al. 2018). Considering the overall toxicity of these heavy metals on human health even at trace concentrations, the development of heavy metal removal technologies to safe hazardous levels in wastewater has gained the researchers attention to safeguard the environment and all living beings on this planet. Several conventional approaches were reported in the literature for the removal of heavy metals from a variety of wastewaters including reverse osmosis, ultra filtration, coagulation, ion exchange, chemical precipitation, complexation, activated carbon adsorption, solvent extraction, electro deposition, cementation, and membrane filtration methods (Atari et al. 2019; Prabhu and Prabhu 2018). But, most of the stated conventional methods are labor-intensive, less efficient, consumes high energy, provides poor removal efficiency, and generates hazardous waste sludge. Although, some of the above stated processes are effective for the heavy metal removal, the high operational cost confined their use in treatment applications (Barakat 2011). For instance, the ion-exchange resins are commercially recognized as effective pollutant adsorbents in wastewater treatments, but the high cost of resins is a bottleneck for their use in large-scale treatments (Al-Enezi et al. 2004). Fortunately, adsorption has emerged as the promising technology by virtue of its simple operation, improved efficiency, low cost, adsorbent reusability, analytes recovery, and availability of vast number of adsorbents (Hegazi 2013). Accordingly, a variety of cost-effective and efficient adsorption approaches have been reported in recent years for the removal of heavy metals from wastewaters (Al-Jlil 2010; Santhy and Selvapathy 2004; Pirajan and Giraldo 2012). Over the years, nanotechnology has provided remarkable applications in polluted water and soil treatments. The utilization of different metal nanoparticles successfully remediated a series of chlorinated organic compounds and heavy metal pollutants from the wastewater and subsurface soils (Liu et al. 2019). The increasing application of nanomaterials in the environmental remediation of contaminated sites has been witnessed over the last couple of decades. The exceptional characteristics of nanomaterials including high surface area and small size extended the reactivity and adsorption capacity. Till date, a variety of nanomaterials offered promising

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results for the heavy metals removal from contaminated soil and waters. Among them, zerovalent iron nanoparticles were regarded as the most promising materials for the treatment of contaminated soil and wastewaters considering their high reactivity, increased stability, and environmental compatibility. These zerovalent iron nanoparticles transfer the electrons to targeted organic and inorganic pollutants and oxidize to form Fe2+ and Fe3+ions, meaning that they act as effective catalysts and reducing agents (Chen et al. 2008). In virtue of the importance of zerovalent iron nanoparticles in heavy metal pollutant remediation, this chapter aims to discuss latest developments regarding the heavy metal pollutants removal using zerovalent iron nanoparticles from wastewaters. The fate and transportation of heavy metals into water bodies and their detrimental effects on human health and ecosystem are discussed in detail. Further, a variety of production methods and characterization techniques of zerovalent iron and their utilization in the organic and metal pollutants removal are clearly described. In addition, recent progresses of zerovalent iron nanoparticles toward the heavy metals removal and their toxicity evaluations are extensively discussed. Finally, the concluding remarks and future perspectives of zerovalent iron nanoparticles as potential applicants for the wastewater treatment are presented.

2.2

Presence and Adverse Effects of Heavy Metals in Water

The acute toxicity and non-biodegradability of heavy metals referred them as most critical pollutants among all other contaminants present in the water bodies and wastewater streams. The increasing disposal of heavy metal wastes particularly mercury, arsenic, chromium, cadmium, and lead from the different human-induced activities has created several environmental impacts on ecology and human health. All metals have tendency to transport with the sediments and cause adverse effects on aquatic flora and fauna; then they bioaccumulate in foods and create immense effects to human health (Gbaruko and Friday 2007). Aquatic environments containing metal concentrations above the water quality standards induce a threatening toxicity to human and animals. Particularly, the impact of heavy metals is strong on fetus and newborn babies, where high level of exposure occurs in human beings. When new born babies exposed to heavy metal pollution, it may leads to damage the brain, disrupt the function of red blood cells, and affect the central nervous system along with other behavioral and physiological complications. Besides, the acute toxicity of heavy metals may also result to develop cancers (Waheed et al. 2014). Apart from human health issues, heavy metals cause serious physiological and morphological changes in plants, which could damage the cell function and subsequently affect the photosynthesis. The toxicity of heavy metals in several plant species produces nutrient deficiency and chlorosis and increased oxidative stress, bleaching, and mutagenic changes. Despite the fact that there are several heavy metals that cause severe toxicity to human beings and plants, we have discussed the most critical and prioritized heavy metal ions, namely, mercury,

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arsenic, chromium, cadmium, and lead in the following sections. The major sources of selected heavy metal ions and their permitted concentrations in water sources along with their human health effects are presented in Table 2.1. Mercury Is a highly toxic heavy metal that occurs from the volcanic eruptions and natural weathering of rocks and soil. It has played a key role in medicine, optics, catalysis, and energy-efficient technologies over the several thousands of years. Further, it is extensively used in different temperature measuring instruments and in a number of industrial chemicals. The anthropogenic mercury mainly generates from the agriculture, mining and processing industries, incineration, and municipal wastewater discharges (Hargreaves et al. 2016). The presence of mercury in aqueous systems is extremely toxic to aquatic animals and human beings. The mercury exists in three major forms, namely, metallic elements, organic compounds, and inorganic salts, among which, its existence in organic form is highly toxic (Fu and Wang 2011). The exposure to mercuric toxicity may result in blindness, brain damage, kidney damage, mental retardation, mitochondrial damage, neurological damage, and paralysis. For instance, in Minamata Bay of Japan, the babies were born with physical and mental disorders to the who mothers consumed mercury-contaminated fish (Harada 1995). Furthermore, mercury toxicity associates with abortions, physiological stress, and tremors. Relatively, the toxicity of methyl mercury is higher than all other forms, and it directly strikes the central nervous system in human. A maximum of 0.002 mg/L concentration of mercury is permitted in the drinking waters in accordance with the Environmental Protection Agency guidelines (Gray et al. 2015). Arsenic Is the prioritized toxic heavy metal that greatly affects the animal and human health. In nature, arsenic exists in both amorphous and crystalline forms and occurs in earth crust deposits. Further, it remains in two major inorganic forms, namely, arsenate and arsenite (Lata and Samadder 2016). In addition to the natural weathering of rocks, several human activities such as smelting and mining processes, over use of pesticides, and combustion of coal release a considerable quantity of arsenic into the environment. The presence of arsenic in groundwater is a matter of concern, and its concentrations were found in the groundwaters of Bangladesh, Brazil, India, Taiwan, and Chile (Herath et al. 2016). The continuous intake of arsenic at higher concentrations through the drinking water may affect the central nervous system, contaminate the blood, cause skin and kidney problems, breathing problems, and nausea, and may induce cancer. The presence of metal oxides of aluminum, manganese, and iron promotes the adsorption of arsenic into the aquatic bodies. According to Environmental Protection Agency, the maximum allowable concentration of arsenic in drinking water is 0.01 mg/L (Ahamed et al. 2006). Cadmium Is usually found in soil and water systems and is considered as the highly toxic element owing to its long half-life. There are several studies reported the carcinogenicity of cadmium even at its trace concentrations. The alloy and metalfinishing industries, electronic and electroplating industries, alkaline batteries, insecticides, coatings, pigments, photography, and several other anthropogenic processes

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Table 2.1 Major sources, permitted concentrations, and human health effects of selected heavy metals in aqueous media Heavy metal Mercury

Major sources Fossil fuels Dental amalgams Old latex paints

Cadmium

Arsenic

Chromium

Incinerators Thermometers Nonferrous metallurgy Electroplating Pesticide and fertilizers Nuclear plants Printing and pigments Chemical industries Acute effects in children Mining and smelting Pesticides and fungicides Coal combustion Industrial arsenic waste Respiratory problems Tanneries Pulp and rubber industry Electroplating

Lead

Dye industry Smoking and automobile Pesticides and paints Mining and batteries Electronic waste Burning of coal

Permitted concentrations (mg/L) Environmental Protection Agency limit 0.002 European community limit 0.001 Water quality regulation of India limit 0.004

Effects on human health Mutagenic effects Mental retardation Neurological damage Abortions and tremors

Environmental Protection Agency limit0.005 European community limit 0.2 Water quality regulation of India limit 0.001

Renal dysfunction Increased blood pressure Gastrointestinal disorder Kidney damage Bone defects Bronchitis and anemia

Environmental Protection Agency limit10 European community limit 10 China health standards 0.01

Environmental Protection Agency limit0.1 European community limit is 0.5 Water quality regulation of India limit 0.1 Environmental Protection Agency limit0.1 European community limit 0.5 Water quality regulation of India limit 0.1

Melanosis, keratosis Hyper pigmentation in humans Nausea and cancers Immunotoxic and genotoxic issues

Skin ulceration Damage to kidneys and lungs Affects central nervous system Toxic to humans, aquatic livestock Hyper tension and brain damage Tiredness and irritability anemia Behavioral changes in children Impact on reproductive system

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are the major sources for cadmium contamination (Idrees et al. 2018). Cigarette smoking is also a source for cadmium poisoning in humans (Ali et al. 2018). The concerns regarding the toxicity of cadmium in the environment are very high as its acute toxicity strikes the livestock and human health even at very low concentrations. The exposure of humans to cadmium causes renal dysfunction, osteomalacia, bone defects, and lung cancer (Johri et al. 2010). Besides, cadmium does not play role in any biological systems, and hence, it is not essential to the body even at trace concentrations. Therefore, it only provides hazardous effects to the human and animals. According to Environmental Protection Agency standards, the permitted cadmium concentration in drinking water is 0.005 mg/L (Burke et al. 2016). Chromium Is another hazardous heavy metal that seriously affects the human health and environment. It occurs with different oxidation states in the environment, among which, chromium (III) and chromium (VI) are the most common forms. The toxicity of chromium (VI) is relatively higher, because of its high solubility and great mobility in water under natural conditions (Vignati et al. 2010). The combustion of oil and coal and usage of fertilizers and other anthropogenic sources from electroplating, tanning, dying, and wood preserving industries are the major sources of chromium contamination. Further, paper, leather, pulp, tanning, and rubber manufacturing industries are the significant contributors of chromium. All these industries release a large amount of wastewater containing hazardous chromium into the water bodies every year (Wang and Choi 2013). The toxicity of chromium greatly affects the biological processes and mainly causes skin ulceration and kidney problems and then affects the central nervous system. The high amount of chromium also reduces the photosynthesis in plant species, and its toxic effects associate with hematological problems and show impact on immune response in freshwater fish. Based on the Environmental Protection Agency guidelines, the allowable concentration of chromium in drinking water is 0.1 mg/L (Zhao et al. 2016). Lead Is the hazardous heavy metal and is being discharged in high quantities into the environment every year. High accumulation and biomagnification tendency of lead within the food chains cause extensive environmental contamination and public health problems throughout the world (Zietz et al. 2010). Most of the lead present in the environment mainly comes from the natural weathering and a variety of anthropogenic activities such as mining, dye industries, pigments and paint manufacturing, battery manufacturing, electroplating industries, and vehicular exhausts. The accumulation of lead in the human body causes several manifestations such as fatigue, hyper tension, anemia, and brain and kidney damage and affects the overall central nervous system (Watt et al. 2000). Human exposure to high dosage of lead results to brain damage and creates encephalopathic symptoms. Besides, exposure to lead rattles the function of gastrointestinal tract and reproductive system. Airborne lead deposits on the agricultural foods and contaminates fruits, seeds, and vegetables. Moreover, newborn babies are more sensitive than adults to lead, and it harms the hemoglobin synthesis and shows severe impact on kidneys (Tang et al. 2017). The Environmental Protection Agency has fixed the standards for maximum lead concentration in drinking water as 0.1 mg/L.

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Zerovalent Iron Nanoparticles

Zerovalent iron nanoparticles used in a wide range of industrial and domestic applications considering their large surface area, small size, high reactivity, charge stabilization, chemical reducibility, surface sorption, and high efficiency. In fact, the combination of all these superior properties raised their tendency to interact, adsorb, and react with the contaminants. In recent years, the reducing properties of zerovalent metals established them as effective remediants for different organic and inorganic pollutants in water treatment (Wenjuan et al. 2018). For instance, zerovalent zinc nanoparticles have been reported to possess excellent degradation properties for dioxins (Bokare et al. 2013). The silver nanoparticles are capable to disinfect water owing to their antimicrobial ability. Next, zerovalent iron nanoparticles have been proved to be the excellent materials for the degradation of all prioritized heavy metal pollutants. Consequently, zerovalent iron nanoparticles have drawn good attention as effective adsorbents for the remediation of a variety of heavy metals comprising mercury, chromium, copper, nickel, cadmium, and lead. Basically, zerovalent iron in nanoparticles acts as reducing agents, while ferric oxide shell provides reactive sites for the electrostatic interaction with heavy metals (Zhang et al. 2019a, b, c, d). The large specific surface area and strong reducing properties of zerovalent iron nanoparticles provide outstanding performance during the removal of heavy metal pollutants from aqueous media. The core shell structure of zerovalent iron nanoparticles representing a variety of removal mechanisms for the heavy metals and chlorinated organic compounds is presented in Fig. 2.1 (O’Carroll et al. 2013). From Fig. 2.1, we can clearly see that the removal mechanism of zerovalent iron nanoparticles varies with the heavy metals in accordance to their standard potential (E0). For instance, the standard potential (E0) of lead (II) is slightly higher than iron (II), and thus the removal mechanism involved is mainly reduction and sorption. Whereas in case of chromium (VI), the standard potential (E0) of it is much greater than iron (II), and thus the removal mechanism involved is reduction and precipitation. Overall, the effective application of zerovalent iron nanoparticles in heavy metal remediation from wastewaters and contaminated sites has sparked a great deal of interest owing to their low cost and high efficiency than the expensive treatment technologies (Zingaretti et al. 2019).

2.3.1

Synthesis of Zerovalent Iron Nanoparticles

Generally, zerovalent iron nanoparticles can be synthesized using a variety of physiochemical approaches. The standard physical methods including grinding, ball milling, inert gas condensation, severe plastic deformation, and lithography are very useful approaches for the synthesis of zerovalent iron nanoparticles. Apart from the abovementioned physical methods, there have been several chemical processes reported for the preparation of zerovalent iron nanoparticles (Vijaya

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Fig. 2.1 The core shell structure of zerovalent iron nanoparticles depicting various removal mechanisms for heavy metals and chlorinated compounds. (Reprinted with permission from O’Carroll et al. 2013, Copyright (2013) Elsevier)

Bhaskar Reddy et al. 2013). The gas–phase reduction, liquid–phase reduction, chemical vapor condensation, controlled coprecipitation, pulse electro deposition, and liquid flame spray methods are some potential chemical methods for the production of zerovalent nanoparticles (Badmus et al. 2018). Besides, zerovalent iron nanoparticles can be synthesized using a reducing agent, i.e., sodium borohydride (NaBH4) in accordance with the Eq. (2.1). Even though the above reaction appears easy to proceed, a high amount of sodium borohydride is necessary to accelerate the process and to ensure the steady growth of zerovalent iron nanoparticles, which is quite expensive and toxic to the environment. Nevertheless, this method offers several difficulties relating the safety issues linked with the consumption of toxic reducing agent (sodium borohydride) as well as with the flammable H2 gas generation during the synthesis. o  þ 4 Fe3þ þ 3 BH 4 þ 9 H2 O ! 4 Fe # þ3 H2 BO3 þ 12 H þ 6H2 "

ð2:1Þ

Other chemical methods to obtain zerovalent iron nanoparticles are (1) the reduction of hematite (α-Fe2O3) or goethite (α-FeOOH) at high temperatures using H2 and (2) the breakdown of iron pentacarbonyl (Fe(CO)5) in organic solvents (Vijaya Bhaskar Reddy et al. 2016). Among the different production methods, chemical reduction is more compatible for the preparation of zerovalent iron nanoparticles due to its simplicity and homogenous product formation. However, zerovalent iron nanoparticles obtained through the chemical methods generate some unavoidable hazardous by-products, which may cause adverse health effects upon their use (Keller et al. 2012). As a result, green chemistry and green synthesis of zerovalent iron nanoparticles have become the cost-effective and environmentally friendly production route that can further reduce the oxidation and agglomeration of resulting nanoparticles. Besides, the green synthesis of zerovalent iron nanoparticles

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is easy to carry out on large-scale production as the process does not require high temperature, high pressure, and additional energy inputs. The first green synthesis process for the production of zerovalent iron nanoparticles was reported by VeruTEK and United States Environmental Protection Agency. In this approach, the plant extracts of green tea, black tea, coffee, lemon, sorghum, and grape were heated to boiling temperature and collected the polyphenolic solution. Later, the received extracts were added to Fe2+ solution, where Fe2+ ions were reduced to zerovalent iron by the addition of polyphenols (Njagi et al. 2010). In a recent study, zerovalent iron nanoparticles were synthesized using green and semi-green approaches and evaluated their performance for the reductive dehalogenation of chlorinated organic compounds in contaminated groundwater samples. In green preparation method, the plant extracts of Virginia creeper, green tea, and coffee acted as reducing and capping agents, while in semigreen method, commercial chemicals were used for the synthesis of zerovalent iron nanoparticles. The authors concluded that, the semi-green method produced smallscale zerovalent iron nanoparticles with greater reducing ability than the green approach. However, an excess amount of sodium borohydride was required for the completion of reduction, which is very expensive and produced hazardous boron compounds that are harmful to the environment (Kozma et al. 2016). In other study, mango peel extracts were successfully utilized to synthesize the zerovalent iron nanoparticles by a low-cost and economically viable approach. The X-ray photoelectron spectroscopy depth profiling studies of the synthesized zerovalent iron nanoparticles showed the growing intensity of Fe0 with decreasing quantity of Fe2 + /Fe3+. The structural analysis of zerovalent iron nanoparticles revealed a layer of polyphenols accompanied by the Fe-oxides and hydroxides onto the metallic Fe providing the shell structure of “Fe3+/Fe2+-polyphenols” complex on core metallic iron (Desalegn et al. 2019). Further, a green process was developed for the zerovalent iron nanoparticle synthesis using green tea extracts as reducing agent and ferric chloride (FeCl3) as iron precursor. The polyphenols were extracted from green tea using microwaveassisted extraction. The study also optimized the most favorable conditions for the synthesis of green zerovalent iron nanoparticles based on the yield, rate of oxidation, and surface area of the metal particles. The optimized conditions reportedly green tea-to-solvent ratio of 1:20, green tea extract flow rate of 8.0 mL/min, and ferric chloride-to-green tea extract ratio of 1:2 produced the highest yield of 86.48% (Banerjee and Chatterjee 2015). In the next study, the authors presented a first time report for the synthesis of zerovalent iron nanoparticles using Urticadioica leaf extracts as reducing and capping agents. Based on the characterization results, the synthesized nanoparticles were completely composed of zerovalent iron without the presence of iron–oxide impurities, and the shape of zerovalent iron nanoparticles was found to be spherical with the particle diameter ranging between 20 and 70 nm. Green synthesized zerovalent iron nanoparticles were completely protected from the oxidation, and it was also found that several zerovalent iron nanoparticles were trapped in the biologic coating and fabricated a complex. These complexes helped to form a stable colloid system with negligible aggregation (Alireza et al. 2017). In

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addition, another method reported for the synthesis of zerovalent iron nanoparticles using the extracts of banana peel and ferrous sulfate in 1:1 ratio. The authors concluded that, in this approach, the zerovalent iron nanoparticles have been synthesized directly from the ferrous sulfate and banana peel extract in aqueous media without adding any capping agents. Although, the size of the resulted zerovalent iron nanoparticles was found >100 nm, the process is environmentally friendly, and the resulted nanoparticles were effective in the remediation of wastewater pollutants (Sunardi et al. 2017). The major advantages of green synthesis methods are the use of nontoxic solvents such as water in place of toxic organic solvents and the utilization of natural materials instead of hazardous commercial products. Further, it is expected that the stability of zerovalent iron nanoparticles synthesized by green methods is relatively higher as the polyphenols collected from the plant extracts act as capping agents (Eslami et al. 2018). Regardless of the numerous advantages of green approaches for the preparation nanoparticles, these processes are still not commonly accepted owing to the lack of knowledge on their reaction mechanism and physiochemical properties like size, agglomeration, and poor reactivity of the resulted nanoparticles (García et al. 2019). Additionally, several other methods including ultrasound- assisted production, carbothermal reduction, precision milling, and electrochemical generation processes are not well accepted at this stage, but they may soon become more popular methods for the synthesis of various metal nanoparticles due to their multiple advantages (Wang et al. 2017).

2.3.2

Characterization of Zerovalent Iron Nanoparticles

A detailed knowledge about the synthesized zerovalent iron nanoparticles regarding their size, specific surface area, surface morphology, crystal structure, and elemental distribution is necessary for the clear understanding of reaction mechanism, kinetics, and the resulted intermediates. Also, the fate and transportation of these zerovalent iron nanoparticles in soil and water environments rely on the above surface characteristics. The combination of spectroscopic and diffraction techniques acts as superior tools for the physical imaging and chemical identification of zerovalent iron nanoparticles. The techniques including transmission electron microscopy, scanning electron microscopy in transmission mode, scanning electron microscopy, energy dispersive X-ray spectroscopy, scanning auger microscopy, and atomic force microscopy were recognized as effective and relatively informative tools for the thorough chemical and structural characterization of zerovalent iron nanoparticles (Madhavi et al. 2014). In addition to the above stated techniques, X-ray diffraction is a nondestructive technique that provides clear information regarding the crystallographic structure, chemical composition, particle size, and molecular data of all metal nanoparticles. The crystallographic study of zerovalent iron nanoparticles carried out by X-ray diffraction displayed the body-centered cubic (bcc) structure (Madhavi et al. 2013). Next, the X-ray photoelectron spectroscopy is another

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important characterization technique that provides (1) the physical state, chemical structure, and valence states of elements based on the peak displacement, (2) the quantitative information of all elements excluding hydrogen and helium, (3) the semiquantitative information for chemical states based on the relative peak intensity ratio, and (4) various elements and their chemical states across the film thickness based on the change of background intensity (Santos et al. 2017). Further, atomic force microscope is the key characterization technique that does not necessitate surface conductivity as it measures the interactions between scanning probe and the sample surface. The inter reactions including van der Waals, electrostatic, surface tension, friction, and magnetic forces surpass the deficiency of scanning tunneling microscopy. Unlike the scanning tunneling microscopy, atomic force microscope mapping discovers the surface heterogeneities and correlate those properties with the topographic properties (Chekli et al. 2016). Besides the above discussed techniques, a variety of other characterization techniques are still accessible for the characterization of zerovalent iron nanoparticles, which includes surface analysis spectroscopy, Mossbauer spectroscopy, and dynamic structure spectroscopy (Sun et al. 2006). At present technology, there are above 100 surface characterization tools existed for the zerovalent iron and other metal nanoparticles. When an individual characterization technique is not capable to provide the sufficient information, it is prerequisite to connect two or more characterization techniques that complement together. The customarily available characterization techniques for the surface characterization of zerovalent iron nanoparticles are shown in Fig. 2.2 (Vijaya Bhaskar Reddy et al. 2016). In general, the reactivity and surface properties of zerovalent iron nanoparticles vary with the time, solution chemistry, and surrounding environments. They are commonly spherical in shape and form chain-like clusters/aggregates. The core– shell structure of zerovalent iron nanoparticles illustrates that the valence iron oxide shells are highly insoluble, and they protect zerovalent iron nano-core from the rapid oxidation (Schmid et al. 2015). The structure and composition of these iron oxide shells vary with the production methods and environmental conditions. For instance, iron oxide nanoparticles produced through sputtering contains maghemite (γ-Fe2O3) or partly oxidized magnetite (Fe3O4). Similarly, nanoparticles obtained by the metallic vapor nucleation also consist maghemite and magnetite with high amount of maghemite in small particles owing to the rapid oxidation and higher surface-tovolume ratio (Ali et al. 2016). In a study, the pristine zerovalent iron nanoparticle morphology was examined using high-resolution transmission electron microscopy and identified the thickness of outer layer of iron oxide and inside core of zerovalent iron as 10 and 20 nm, respectively (Kawamoto et al. 2013). In other studies, zerovalent iron nanoparticles were synthesized and characterized to observe the morphology of different iron surfaces. The size measurements revealed the average size of zerovalent iron nanoparticles ranged between 20 and 50 nm. Due to the strong magnetic forces, zerovalent iron nanoparticles were not well separated and closely distributed forming the core–shell structure. Also, the authors revealed that the zerovalent iron nanoparticles were oxidized and cracked to form 10–30 nm size particles because of the Fe(III) nanoparticles (i.e., Fe2O3) aggregation on zerovalent

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Fig. 2.2 A detailed list of commonly employed characterization techniques for the zerovalent iron nanoparticles analysis. (Reprinted with permission from Vijaya Bhaskar Reddy et al. 2016, Copyrights (2016) Elsevier)

iron nanoparticle surface (Lin et al. 2008; Su et al. 2015). Therefore, it is unclear from the existing literature to infer the effect of shell structure and composition variations on zerovalent iron nanoparticle reactivity, transportation, and aggregation behavior. The fundamental zerovalent iron nanoparticles consist spherical shape particles with an average diameter less than 100 nm. But, the formation of iron oxide shell on the surface of zerovalent iron nanoparticles decreases the reactivity of zerovalent iron afterward.

2.3.3

Potential Applications of Zerovalent Iron Nanoparticles

Over the last two decades, zerovalent iron nanoparticles have received great attention for the remediation of a broad range of environmental pollutants considering their reactivity and higher surface area. The major applications of zerovalent iron nanoparticles were primarily concentrated on their electron-donating properties. These zerovalent iron nanoparticles are sufficiently reactive in aqueous media under ambient conditions and act as good electron donors (Dong et al. 2019). Zerovalent iron nanoparticles are the first reactive materials that were employed in groundwater treatment through the permeable reactive barriers. Considering the small size of zerovalent iron nanoparticles, the slurries can be injected either by gravity or under pressure to the contaminated area. The zerovalent iron nanoparticle

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permeable reactive barriers were employed to remove highly toxic contaminants of concern including organochlorine pesticides, tetrachloroethylene, trichloroethylene, and polychlorinated biphenyls (Henderson and Demond 2007). The effectiveness of zerovalent iron nanoparticle permeable reactive barriers depends on the optimization of experimental parameters. Initially, Gillham and O’Hannesin reported the efficiency of zerovalent iron nanoparticles for the dehalogenation of 1,4-chlorinated ethane (Gillham and O’Hannesin 1994). During the same year, Matheson and Tratnyek investigated the quick dehalogenation of carbon tetrachloride and trichloroethylene using zerovalent iron nanoparticles (Matheson and Tratnyek 1994). In addition, zerovalent iron nanoparticles have successfully degraded a vast range of organic pollutants and pesticides especially vinyl chloride, trichloroethylene, carbon tetrachloride, research department explosive, trinitrotoluene, lindane, and dichlorodiphenyltrichloroethane and provided the evidence for their application in the remediation of other contaminated sites (Song and Carraway 2005; Fu et al. 2014). In the remediation of organohalogen compounds, zerovalent iron serves as powerful electron donor and transfers the electrons from iron surface to the targeted pollutant to facilitate the reductive dehalogenation. In addition to the organohalogen compounds, zerovalent iron nanoparticles have showed their potential for the effective remediation of several other soil and water contaminants. The remediation of textile dyes and heavy metals using zerovalent iron nanoparticles has led the scientific community to adopt this technology rapidly as a new alternate tool for the environmental cleanup (Raman and Kanmani 2016). Over the past few years, several studies successfully demonstrated the remediation of dyes and heavy metals from wastewater samples. Among the dyes and heavy metals, the higher removal efficiency was obtained for the dyes due to their strong adsorption and reduction capacities with the zerovalent iron nanoparticles. However, many parameters including pH, reaction temperature, oxidation–reduction potential, ionic strength of the polluted water, hardness, and the presence of natural organic matter play crucial role to determine the overall efficiency of zerovalent iron nanoparticles (Rahman et al. 2014). Most importantly, azo dyes that are resistant in natural environment were efficiently removed by zerovalent iron nanoparticles (Satapanajaru et al. 2011). Few studies reported the efficient processes for the degradation of various azo dyes using the environmentally friendly zerovalent iron nanoparticles, in which the process parameters, namely, solution pH, particles size, amount of zerovalent iron nanoparticles, and initial dye concentrations were optimized to minimize the chromophoric groups and the content of total organic carbon of the azo dyes, namely, remazol black B, methyl orange, acid orange 8, and acid orange 7. The dye removal processes were proved to be effective alternatives to breakdown various azo dyes in aqueous systems (Pereira and Freire 2006; Zhao et al. 2008). Zerovalent iron nanoparticles are further impressive materials for the removal of different pharmaceuticals, namely, diazepam, amoxicillin, chloramphenicol, tetracycline, ampicillin, and metronidazole under the optimized reactant concentration, pH, temperature, and competitive anions (Mackulak et al. 2016). A study presented for the enhanced degradation of diazepam using zerovalent iron nanoparticles. In

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this study, the authors proposed that Fe0/EDTA/O2 appeared to be a Fenton-like process as it has produced hydroxylated intermediates as a result of OH attack on diazepam rings and promoted the degradation of diazepam (96%) and its mineralization (60%) significantly, denoting the process efficiency for the degradation of complex molecules as pharmaceuticals. The degradation efficiency was increased with increasing zerovalent iron nanoparticles and with the presence of ethylenediamine tetra acetic acid (Bautitz et al. 2012). In another study, the authors proposed the removal of chloramphenicol from aqueous solutions applying zerovalent iron nanoparticles. The authors described that the removal efficiency was found proportional with zerovalent iron nanoparticle dosage. The zerovalent iron nanoparticles catalyzed by O2 showed higher removal efficiency than the process catalyzed by N2 due to the Fenton-like reaction (Siqing et al. 2014). Further, there are two other studies reported for the removal of cytostatic drugs, antibiotics and diagnostic agents using zerovalent iron nanoparticles. The studies proved that the concentrations were substantially decreased for all selected drugs when treated with zerovalent iron nanoparticles. Further, the studies described that redox process is involved in the transformation of selected pharmaceuticals. Also, the removal efficiency of the methods confirmed the ability of zerovalent iron nanoparticles in the remediation process of pharmaceuticals from aqueous samples (Stieber et al. 2011; Machado et al. 2017). A list of common environmental pollutants that have been successfully remediated by zerovalent iron nanoparticles are presented in Table 2.2.

2.4

Removal of Selected Heavy Metals Using Zerovalent Iron Nanoparticles

Heavy metals are nonbiodegradable pollutants, and it is very difficult to eliminate them by natural methods from the environment. The wastewater generated from the wide array of industries including electroplating, metal smelting, electronic and battery manufacturer, and alloy manufacture industries contains complex heavy metal ions that are readily soluble in water and possess high tendency to accumulate in living organisms. Almost all heavy metal ions are found to be toxic and carcinogenic to the biota. In recent years, nanotechnology has successfully introduced different types of nanomaterials for the environment remediation. Some benefits include increased efficiency for contaminant removal, reduced consumption of raw materials, and substitution of more abundant and less toxic materials than ones currently used. Nanomaterials have worked as effective adsorbents for the removal of heavy metals in aqueous systems due to their high surface reactivity and large specific surface area. Over the past few years, zerovalent iron nanoparticles have demonstrated effective remediation approaches for the specific hazardous heavy metals in wastewaters including mercury, chromium, cadmium, arsenic, copper, lead, and nickel (Shaolin et al. 2017). Generally, the high reducing power of zerovalent iron nanoparticles facilitates the rapid heavy metal remediation

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Table 2.2 A list of common environmental pollutants successfully remediated using zerovalent iron nanoparticles Heavy metals Mercury Cadmium Arsenic Chromium Lead Nickel Chlorinated alkenes Terachloroethene Trichloroethene Dichloroethene Vinyl chloride

Organic dyes Orange II Acid red Acid Orange Chrisoidine Orange G

Polychlorinated compounds Polychlorinated biphenyls Dioxins Pentachlorophenol 2,4-Dichlorophenol 2,3-Dichlorophenol

Polyaromatic hydrocarbons Quinoline Pyrene Naphthalene Isoquinoline Phenanthrene

Chlorinated methanes Carbon tetrachloride Dichloromethane Chloroform Chloromethane

Pesticides Aldrin Lindane 2,6-dinitroanilines Malathion Chlorpyrifos Atrazine Chlorinated benzenes Chlorobenzenes Dichlorobenzenes Trichlorobenzenes Tetrachlorobenzenes Petachlorobenzenes Hexachlorobenzenes Inorganic anions Dichromates Nitrates Perchlorates Chlorates Bromates

approaches. But, in some cases, iron oxide on particle shell acts as adsorbent and captures metals or metalloids. A schematic representation for the effective heavy metal remediation in aqueous media using zerovalent iron nanoparticles is depicted in Fig. 2.3. In heavy metal remediation approaches, generally, zerovalent iron contributes high reducing capacity, and iron oxide shell (Fe2O3) provides reactive sites for the electrostatic interaction of heavy metals with zerovalent iron nanoparticles. Moreover, when particle size of zerovalent iron nanoparticles is controlled during the synthesis, a large number of reactive sites will be available for the heavy metal interactions. The availability of heavy metals in wastewater varies with the chemical speciation and reaction conditions at contaminated site such as pH, associated minerals, and concentration of organic matter (Zou et al. 2016). The metallic pollutants containing higher positive E0 than iron will be reduced directly to their elemental state. The abundant reactive sites on zerovalent iron nanoparticles provide high reactivity and greater efficiency for the heavy metal removal from contaminated waters. The action mechanism of zerovalent iron nanoparticles toward the heavy metal remediation involves redox processes, in which they donate electrons to the metallic pollutants for their reduction (Zarime et al. 2018). In recent years, a large number of studies utilized zerovalent iron nanoparticles for the successful demonstration of heavy metals remediation in wastewaters. A few most significant studies have been discussed in the following sections and presented in Table 2.3.

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Heavy metals Mercury (Hg), Cadmium (Cd), Arsenic (As), Chromium (Cr), etc.,

Properties Size, shape, concentration, structural composition, surface activity and physiochemical properties

ZVINPs

Performance Operating conditions, Intrinsic characteristics, Supporting materials

Fig. 2.3 Insights and parameters that affect the removal of heavy metal pollutants from aqueous media using zerovalent iron nanoparticles

2.4.1

Removal of Mercury

Mercury is a highly toxic heavy metal, and often it exists in the form of mercuric sulfide. The volcanoes generate nearly half of the mercury emissions that are released into the atmosphere, and the remaining half is being produced by the various human activities including 65% from combustion, 11% from gold production, 6.8% from metal production, 6.4% from cement industries, and 3.0% each from waste disposal and caustic soda production. At times, liquid mercury was employed as cooling media in nuclear reactors, but it has been substituted with sodium considering the high density of mercury, due to which it consumes high energy to circulate in the reactors. The inorganic mercury compounds are persistent and produce adverse health hazards if accumulated in the drinking water (Kim and Zoh 2012). The drinking water sources close to industries and wastewater streams may often contaminate with mercury discharges and become fatal to the aquatic and human life. Long time exposure to mercury through drinking water shows highly negative impacts on the kidneys, brain, lungs, skin, and eyes. Considering the serious health hazards of mercury and potential ability of zerovalent iron nanoparticles for the remediation of heavy metals, several zerovalent iron-based remediation approaches were established in recent years for the removal of mercury from water bodies. In a recent study, the authors determined the efficiency of zerovalent iron– granular activated carbon to remediate methylmercury in wetland sediments containing plants. The study revealed a lower concentration of methylmercury in

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Table 2.3 List of methods recently reported for the removal of selected heavy metal pollutants using zerovalent iron nanoparticles S. No 1

Material ZVI-GA

Targeted contaminant Me-Hg

Concentration (mg/L) 1–10

2

P-nZVI

Hg, Cr

40–100

3

nZVI-OBA

Hg, Pb

25–50

4

nZVI

Cd, Pb, Cu, Ni

20–100

5

nZVI

Cd, Pb, Cu, Ni

200

6

nZVI/rGOs

Cd

150

Key findings ZVI-GAC reduced mercury concentration in wetland sediments containing plants Adsorption was the dominant mechanism The material was easily regenerated Increased pH enhanced hg removal but gradually decreased Cr removal Successfully removed hg and Pb Aggregation and oxidation of bare zerovalent-iron minimized adding OBA biological support The removal order Pb2 + > Cu2+ > Cd2+ > Ni2+ Formation of FeOOH on ZVNI surface was crucial pH < 6.5 was found ideal for process efficiency Highest removal for Cu (II) and Pb(II) and lowest removal for Cd (II) ZVNI particles were separated easily Graphene addition inhibited the aggregation nZVI/rGOs regenerated by plasma technique The nZVI/rGOs were efficient for metal removal

References Lewis et al. (2016)

Liu et al. (2015)

Amiri et al. (2017)

Azzam et al. (2016)

Danila et al. (2018)

Li et al. (2016)

(continued)

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Table 2.3 (continued) S. No 7

Material nZVI/KBC

Targeted contaminant Cd

Concentration (mg/L) 5–50

8

CS-P-nZVI

Ar

20–100

9

TiO2 + ZVI

Ar

10

10

nZVI/gC3N4

Ar

0.1–500

11

SA-nZVI

Cr

5.0

12

H-nZVI

Cr

1000

Key findings Synergistic effects between biochar and nZVI enhanced the removal of Cd(II) Cd was partly reduced during the removal process The stability of CS-PnZVI was enhanced Ar was removed quickly by CS-PnZVI, Ar was physically adsorbed within 60 min The (TiO2 + ZVI) gave best removal at pH 9.0. Increase of ZVI:TiO2 ratio increased the efficiency of Ar uptake to certain extent The maximum adsorption capacity was 70.3 mg/g Ar removal by nZVI/ g-C3N4 accompanied with arsenic reduction and iron oxidation Sodium alginate enhanced nZVI dispersion SA-nZVI provided higher removal than ZVI provided better stability than the CMC-nZVI H-nZVI provided high efficiency removal of Cr(VI), Mg2+ was extraordinarily enhanced Cr (VI) removal Physical adsorption and reduction involved

References Zhu et al. (2019)

Liu et al. (2016)

Lopez-Munoz et al. (2017)

Chen et al. (2017)

Li et al. (2019)

Fu et al. (2017)

(continued)

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Table 2.3 (continued) Targeted contaminant Cr

Concentration (mg/L) 50

nZVIQuartz

Pb, Ni

25

15

L-cystnZVI

Pb

10–50

16

MgO-nZVI

Pb

5.0–70

S. No 13

Material nZVI-RS

14

Key findings Silicon particles served as support on nZVI Cr(VI) removal was enhanced by siliconrich biochar nZVIRS700 showed the highest Cr (VI) removal nZVI significantly reduced Pb and Ni concentration It is a safe process to treat waste and economical L-cyst-nZVI prepared by chemical reduction method Almost 100% of Pb (II) was removed in 25 min Economical and green process for Pb (II) removal MgO-nZVI showed much better performance Removal proceeds by second-order kinetics

References Qian et al. (2019)

Ghasemzadeh and Bostani (2017)

Bagbi et al. (2017)

Siciliano and Limonti (2018)

zerovalent iron–granular activated carbon-treated samples compared to the untreated samples. They also concluded that the experiments with intact vegetated sediments did not change the methylmercury concentration significantly, because the reduction of methylmercury has occurred through the adsorption but not through the methylmercury demethylation (Lewis et al. 2016). In other study, pumice- supported iron nanoparticles were synthesized and applied for the removal of mercury and chromium in wastewaters. The reported removal efficiencies were 332.4 mg Hg/g and 306.6 mg Cr/g for mercury and chromium, respectively, on pumice-supported iron nanoparticles, which was relatively higher than the efficiency of zerovalent iron nanoparticles. They also reported that higher pH increased the removal of mercury but gradually decreased the removal of chromium ion. Overall, the study described that pumice-supported iron nanoparticles favored the in situ remediation of heavy metals at greater concentration (Liu et al. 2015). In addition, a study reported the modification of zerovalent iron with ostrich bone ash (OBA) and its application for the removal of mercury and lead using a fixed-bed column. The study suggested that

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the rapid aggregation and oxidation of bare zerovalent iron nanoparticles are solved by immobilizing the zerovalent iron nanoparticles on the surface of ostrich bone ash. The removal mechanism of mercury and lead was carried out through the coupled iron–mercury and iron–lead redox reaction (Amiri et al. 2017). Further, there were two studies that compared the efficiency of different commercial zerovalent iron nanoparticles for the immobilization of arsenic and mercury in contaminated soil and groundwater samples. The studies reported that all three commercial zerovalent iron nanoparticles minimized the arsenic and mercury concentrations to significant extent in soil samples and decreased the soil phytotoxicity. The 5.0% zerovalent iron nanoparticles decreased 70% of arsenic, and 10% zerovalent iron nanoparticles reduced nearly 65–90% of mercury depends on the soil type. It has drawn from the results that the interactions between mercury and zerovalent iron were dominated by mercury adsorption, which is the major pathway for mercury removal from soil and solutions. Thus, zerovalent iron nanoparticles have been proposed for the remediation of arsenic and mercury contaminated water and soils optimizing the zerovalent iron concentration and reaction conditions (Gil-Díaz et al. 2017; Vernon and Bonzongo 2014).

2.4.2

Removal of Cadmium

Cadmium is the hazardous heavy metal that is extensively distributed in soil, water, and sediments. It is highly soluble in water than other heavy metal elements, and thus it can directly create the toxicity to human and livestock even at low concentrations. Therefore, cadmium remediation is very essential considering its acute toxicity to animals, plants, and human. The development of nanotechnology has shown its ability to remediate the heavy metals along with other organic pollutants. In recent years, zerovalent iron nanoparticles have provided promising results for the removal of cadmium from polluted water and wastewaters through the adsorption and reduction mechanisms (Arshadi et al. 2014). A recent study reported the synthesis of zerovalent iron nanoparticles using sodium borohydride as reducing agent and examined the efficiency of resulted nanoparticles for the adsorption of cadmium, lead, copper, and nickel metal pollutants. The average particle size of the synthesized zerovalent iron nanoparticles was found to be 43  12 nm with spherical shape. The sorption capacities for selected metal ions followed the order lead>copper>cadmium>nickel. The authors proposed that the formation of an ultrathin FeOOH layer on zerovalent iron surface would be beneficial to provide the interaction between cationic metals and to enhance the adsorption affinity. As a result, the authors suggested pH < 6.5 is best suitable for the simultaneous removal of multiple (Azzam et al. 2016). Other study examined the efficiency of commercial zerovalent iron nanoparticles dispersed in aqueous solutions for the removal of copper, lead, and nickel ions in aqueous solutions. Among the selected metal ions, zerovalent iron nanoparticles have showed highest removal efficiency for copper and lead and lowest removal efficiency for cadmium. Also, the authors recommended using this technique for the decontamination of water from

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heavy metals as it is easy to separate zerovalent iron nanoparticles from water after the remediation process to minimize the redissolution of heavy metals at lower pH values (Danila et al. 2018). Further, a study reported the synthesis of reduced graphene oxide-supported zerovalent iron nanoparticles by the plasma reduction method to enhance the stability and reactivity of zerovalent iron nanoparticles. The results confirmed the higher adsorption capacity of reduced graphene oxidesupported zerovalent iron nanoparticles for cadmium (425.71 mg/g), which was greater than the other adsorbents. In addition, the study described the regeneration of reduced graphene oxide-supported zerovalent iron nanoparticles using plasma reduction technique and reused the materials for several cycles (Li et al. 2016). In a latest study, the synergistic effect of K2CO3-activated porous biochar and zerovalent iron nanoparticles for the removal of cadmium was examined. The results discovered the strong synergistic effects between biochar and zerovalent iron nanoparticles, which promoted the removal of cadmium from aqueous solutions. The coupling mechanism of cadmium with biochar-supported zerovalent iron was found to be complexation–reduction (Zhu et al. 2019). Additionally, in several studies, zerovalent iron nanoparticles were stabilized on different supports including ascorbic acid, carboxymethyl cellulose, and sulfide to evaluate their performance efficiency toward cadmium removal from contaminated waters. The studies were optimized in terms of pH, initial cadmium concentration, contact time, and modified zerovalent iron nanoparticle dosage. The studies revealed that the surface modified zerovalent iron nanoparticles were extremely stable and effective for the removal of cadmium pollutant. Also, the authors suggested that the demonstrated zerovalent iron nanoparticles would be promising adsorbents for the immobilization of cadmium from contaminated water (Savasari et al. 2015; Lv et al. 2018; Huang et al. 2018).

2.4.3

Removal of Arsenic

Arsenic contamination has become the global issue, and it exists predominantly as arsenic (III) and arsenic (V) in groundwater. It is one of the most hazardous heavy metal and is regarded as the first priority pollutant by the World Health Organization. The utilization of zerovalent iron for the removal of arsenic from contaminated waters is one of the effective strategies considering their high adsorption capacity toward arsenic (III) and arsenic (V) ions. The removal of arsenic using zerovalent iron nanoparticles occurred through the adsorption, reduction, surface precipitation, and coprecipitation by forming corrosion products like ferrous/ferric hydroxides. Arsenic (III) is more toxic than arsenic (V), and its removal is relatively difficult than the removal of arsenic (V). The adsorption of arsenic is mainly influenced by the electrostatic interactions, pH, and corrosion phases on zerovalent iron nanoparticles. Additionally, dissolved oxygen also plays an important role in arsenic removal by zerovalent iron nanoparticles (Lata and Samadder 2016). In a study, chitosan–pumice modified zerovalent iron nanoparticles were synthesized and examined their efficiency for arsenic (III) removal from wastewater. The

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characterization studies revealed that chitosan–pumice addition protected the zerovalent iron nanoparticles from aggregation and oxidation. Also, the results revealed that chitosan–pumice modified zerovalent iron nanoparticles removed the arsenic (III) by adsorption following the pseudo-first-order and pseudo-second-order reaction kinetics. The negative ΔG0 and ΔH0 values indicated that the adsorption of arsenic (III) on chitosan–pumice modified zerovalent iron nanoparticles was the spontaneous and exothermic process (Liu et al. 2016). In other study, the removal of arsenic (III) from aqueous solutions using TiO2, zerovalent iron and combined TiO2zerovalent iron was evaluated under UV irradiation. The results confirmed that the solution pH influenced the rate of oxidation of arsenic (III) and the effectiveness of overall arsenic (III) uptake. Also, the study suggested that acidic pH of the solution promoted the uptake of produced arsenic in the presence of metallic iron. The combination of zerovalent iron and TiO2 provided highest arsenic removal from the contaminated waters (Lopez-Munoz et al. 2017). Next in a study, ternary zerovalent iron/phosphotungstic acid/g-C3N4was prepared via photo reduction of Fe(II) ions assisted by phosphotungstic acid and examined its efficiency for the removal of arsenic (III) and arsenic (V) in wastewater. The better performance was obtained for the removal of arsenic (V) over arsenic (III) using zerovalent iron/ phosphotungstic acid/g-C3N4particles. The study suggested that the removal of arsenic (V) proceeds by a simultaneous reduction of arsenic (V) and oxidation of zerovalent iron (Chen et al. 2017). Further, two other methods reported for the synthesis of zerovalent iron nanoparticles supported on a mesoporous type of carbon matrix starch-derived material (Starbon) and Fuller’s earth immobilized zerovalent iron by borohydride reduction method. Both the materials have been evaluated for the removal of arsenic and provided higher removal efficiency for the removal of arsenic (III). They discussed the function of surface OH groups of iron oxide in arsenic removal and the crucial effect of pH on removal efficiency (Baikousi et al. 2015; Yadav et al. 2016).

2.4.4

Removal of Chromium

Chromium is largely disseminated in the earth crust, and a high amount of it is entering into the water bodies through a variety of human activities including leather, tanning, pigments, paints, fungicides, ceramic, photography, and chrome plating industries. According to the International Agency for Research on Cancer regulations, chromium (VI) is deemed to be the class-A human carcinogen. It is a hazardous heavy metal element that remains in the wastewater at high concentrations. Furthermore, the presence of chromium at high concentrations in the aquatic environments is an alarming sign to the industries that involved in chromium consumption or manufacture. In water, chromium produces hydroxides and have tendency to adsorb at high pH values. The ratio of chromium (III) to chromium (VI) greatly varies in the surface water, and hence, a high amount of chromium (III) occurs onsite. Even though there are several conventional methods including membrane filtration, electrochemical precipitation, evaporation, ion exchange,

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adsorption, chelating, solvent extraction, electrolysis, and reverse osmosis processes reported for the removal of chromium, adsorption has offered significant advantages with its design flexibility, high efficiency, easy operation, and low cost over other conventional processes (Rivero-Huguet and Marshall 2009). In recent years, zerovalent iron nanoparticles have evolved as better adsorbents for the effective removal of chromium (VI) from water and wastewaters. A huge number of studies reported in recent years for the effective removal of chromium (VI) from different water bodies using zerovalent iron as reduction agent. Generally, chromium (VI) reduction using zerovalent iron proceeds in a series of chemical reactions. The reduction of chromium (VI) by zerovalent iron results to form iron (III) and chromium (III)) ions. The resulted chromium (III) can be removed through the precipitation or coprecipitation on reaction with mixed iron (III) and chromium (III) hydroxides. However, considering the agglomeration, quick sedimentation, and restricted mobility of zerovalent iron nanoparticles in the aquatic environment, they are not widely applied in the groundwater treatment (Madhavi et al. 2013). But, in a recent study, the zerovalent iron nanoparticles were modified using sodium alginate and prepared a well-dispersed sodium alginate modified zerovalent iron nanoparticles. The synthesized nanoparticles were found embedded in the polymer material and existed in amorphous state with the diameter %95%

–

100 A/m2

56.2% 11.7% 18.6% 21.7%

1170 kW h/m3

88 kWh/mg COD 425 kWh/mg COD

UVC 254 nm, 8500 μW/cm2

10–30 mA cm2

COD chemical oxygen demand, DOC dissolved organic carbon, TOC total organic carbon, UVC ultraviolet C

22 mg L1 of COD

171 mg L1 of COD

Reverse osmosis concentrate (municipal wastewater)

Reverse osmosis concentrate (mixed domestic and textile wastewater)

1h

5h (pH 8.8)

1Ah/L

Hurwitz et al. (2014)

Hurwitz et al. (2014)

Van Hege et al. (2004)

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carbon over a wide range of molecular weight. Moreover, reduced halogenated by-product formation and reduced energy consumption were observed in the hybrid process over the individual photochemical and electro-Fenton processes.

4.3.3

Tannery Industry Effluent

Treatment of effluent from the tannery industry by conventional methods is unsuitable due to its highly contaminated and acidic nature (Borba et al. 2018; Webler et al. 2019). Due to its acidic nature, electrochemical methods are the most suitable methods for the treatment (Martínez-Huitle and Ferro 2006). Boye et al. (2009) compared the efficiency of different electrochemical treatments for the pollutant abatement in tanning process effluent. As per their observation, H2O2-assisted electroprecipitation with sacrificial Fe anode and Ti (Pt) cathode is more efficient (90% of chemical oxygen demand removal in 8 h) than electroFenton, photo-electro-Fenton, and diamond-based electrooxidation. Even though electro-Fenton or photo-electron-Fenton with boron-doped diamond anode and O2 diffusion cathode showed high removal efficiency (75–80%), its performance is constrained by the presence of suspended solids and oxalate complex formation in tanning solutions. Diamond-based anodic oxidation achieves 99% destruction of chemical oxygen demand, but limited by the requirement of longer reaction time, more than 20 h. Szpyrkowicz et al. (2005) studied the treatment of tannery wastewater by using the anodic oxidation process. In this study, the author used four different anode materials, Ti/RhOx-TiO2, Ti/PdO–Co3O4, Ti/PbO2, and Ti/Pt–Ir, and stainless steel is the cathode for all the experiments. Among these four, only two anode materials (Ti/Pt–Ir and Ti/PdO–Co3O4) are found suitable for the tannery waste wastewater treatment as these two electrodes are able to generate active chlorine during the process (Szpyrkowicz et al. 2005). Non-suitability of the remaining two electrodes for the treatment may be due to poor reproducibility and scaling of the electrode material. Isarain-Chávez et al. (2014) investigated the treatment of tannery effluent by using electro-Fenton process. Boron-doped diamond electrodes were used as both anode and cathode for this process. The results showed that current density and Fe2+ concentration were positively influencing the pollutant removal. Kurt et al. (2007) reported the treatment of tannery wastewater collected from the outlet of an equalization basin of common treatment plant by a batch electro-Fenton process. The authors investigated the treatability of wastewater at different pH conditions. More than 70% chemical oxygen demand removal was observed at pH 3, whereas at neutral pH, around 60% chemical oxygen demand removal was observed. Isarain-Chávez et al. (2014) investigated the treatment of tannery wastewater by a photo-electro-Fenton process with boron-doped diamond electrodes used as both the anode and cathode for this process. Authors’ studies showed that UVA irradiation

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enhanced mineralization of organic contaminants in the wastewater with two different concentrations of Fe2+, 1.0 and 3.0 mM, and application of current of 65 and 111 A/m2. They observed that there is no much-pronounced difference between the two conditions, and the highest total organic carbon removal of 80% was observed at Fe2+ of 3.0 mM at applied current of 111 A/m2, whereas 75% with Fe2+ of 1.0 mM at 111 A/m2. The authors extended their studies up to combined process: electrocoagulation followed by photo-electro-Fenton process. From this study, it was observed that the removal efficiency of the total organic carbon from this combined process is 79% in 180 min which is almost similar to the removal efficiency of photo-electro-Fenton process alone, and the same process extended further for up to 360 min in which the removal of the total organic carbon was 90% at Fe2+ of 3.0 mM at 111 A/m2. This is mainly due to oxidation ability of boron-doped diamond, •OH and the UVA light in photo-electro-Fenton to effectively mineralize the organic pollutants in the tannery wastewater. Borba et al. (2018) have performed batch photo-peroxi-electrocoagulation for the treatment of tannery industry effluent under the operational conditions of pH 4, applied current density of 34.2 mAcm2, and H2O2 L1 of 6 g for 420 min. It was reported that maximum removal efficiency of evaluated parameters was observed within 120 min, resulting in the following reductions: 80% of chemical oxygen demand, 95% of color, 98% of turbidity, 88% of total suspended solids, 96% of total fixed solids, and 83% of total volatile solids. The higher activity of the process is due to electrochemical and Fenton-associated phenomena. During the process, the authors observed an increase in the hydrogen peroxide concentration in the system which is mainly due to oxygen discharged on the anode, and H+ dissolution on the cathode leads to in situ formation of hydrogen peroxide in the system.

4.3.4

Landfill Leachate

Landfill leachate generating from the municipal solid waste dumping site is a great challenge. These leachates cause soil as well as groundwater pollution. The low biodegradability of stabilized leachate makes its treatment quite difficult compared to young leachate. The anodic oxidation is one of the best strategies for the oxidation of landfill leachate. The anodic oxidation of leachate with boron-doped diamond as the anode and tantalum plate as the cathode could remove the color, biological oxygen demand, chemical oxygen demand, and N-NH4+ effectively. The borondoped diamond electrode during electrolysis helps in the evolution of high concentration of O2, and it also generates hydroxyl radicals (Nidheesh et al. 2019; Panizza et al. 2008; Panizza and Cerisola 2005; Salazar-Banda et al. 2006). During the electrolysis with boron-doped diamond electrode of 25 mA/cm2 of current density, the high change in the redox potential from –420 mV to 330 mV was reported, which confirms the presence of oxygen and other oxidants (Fudala-Ksiazek et al. 2018). The thickness of boron doped on the diamond (boron-doped diamond 0.5 k to boron-doped diamond -10 k) has very less significance in the removal of

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color and chemical oxygen demand. The generation of hydroxyl radicals by borondoped diamond is independent of the thickness of boron at the surface of electrode (Fierro et al. 2011). Irrespective of the boron doping, 99% of color and 81% of the chemical oxygen demand could be removed from the leachate. The comparative less removal of chemical oxygen demand and the color might be because of the resistance offered by the organic compounds (Fudala-Ksiazek et al. 2018). Also the removal of color and chemical oxygen demand happened at the surface of the electrode and out on the limitation of mass transfer (Fudala-Ksiazek et al. 2018). It is also reported that because of the hydroxyl radicals, the organic matter degradation can be affected by other oxidants like SO4-• and HOCl ions, which generates in the electrolytic reactor depending on the morphology of the electrode and chemical combination of the electrode surface (Fierro et al. 2011). The high sp3/ sp2 carbon ratio gives better organic matter degradation compared to that of low sp3/ sp2 carbon ratio (de Araújo et al. 2014). For example, a boron-doped diamond electrode with higher diamond–carbon content shows more organic matter decomposition compared to the anode with higher graphite–carbon content. The increase in current density has a prominent effect on the removal of chemical oxygen demand from the landfill leachate (12.5, 25, and 50 mA cm2) (Fudala-Ksiazek et al. 2018). At low current density, the boron-doped diamond consumes less energy for the generation of O2 and Cl2 in the system compared to that of the high current density independent of the doping thickness, which reduces the rate of decomposition of the organic matter with an increase in current density (Cabeza et al. 2007). The solution pH has a significant effect on the anodic oxidation of organic matter by boron-doped diamond. The acidic pH (pH 3) supports the anodic oxidation of organic matter through the formation of hydroxyl radical and chlorine gas generation. At the nearneutral pH, the presence of carbonates and bicarbonates acts as the scavengers to the hydroxyl radical affecting the removal of chemical oxygen demand (Cossu et al. 1998; Fudala-Ksiazek et al. 2018). Also, the addition of Fe in the system leads to electro-Fenton reactions, thus increasing the removal efficiency at low pH (Kishimoto and Sugimura 2010). The boron-doped diamond anode oxidation can also remove ammoniacal nitrogen from the system effectively. The ammoniacal nitrogen can be removed by direct means (oxidation by hydroxyl radicals) and indirect means (through the formation of Cl2/HOCl at high dissolved solid concentration). The ammoniacal nitrogen gets oxidized into nitrate ions as well as nitrogen gas during the anodic oxidation with boron-doped diamond anode. The generation of nitrogen gas and nitrate ions follows around 7:3 ratio during the boron-doped diamond oxidation (Fudala-Ksiazek et al. 2018; Pérez et al. 2012). Fudala-Ksiazek et al. (2018) reported 41% ammoniacal nitrogen removal from the landfill leachate for boron-doped diamond -0.5k and boron-doped diamond -10k electrodes. The anodic oxidation of ammoniacal nitrogen with boron-doped diamond -0.5k and boron-doped diamond -10k is independent of current density. The ammoniacal nitrogen removal rate increases with increase in the electrolysis time, as mostly the removal of ammoniacal nitrogen happens due to the indirect oxidation by Cl2, HOCl, and hydroxyl radical (Pérez et al. 2012).

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Moreover, the generation of chlorine gas and further decomposition of HOCl depend on the current density (Lacasa et al. 2012). The low sp3/sp2 carbon ratio by boron-doped diamond -10k enhances the conversion of chloride to chlorine gas when compared to high carbon ratio electrodes (Fudala-Ksiazek et al. 2018). As the ammoniacal nitrogen oxidation happens due to indirect oxidation, the pH has an important role. At acidic pH, the indirect oxidizing species turn out to be Cl2 (aq)/ Cl2 , which has less reactability with N-NH4+. At near-neutral pH, the existence of HOCl oxidizing species oxidizes the ammoniacal nitrogen (Fudala-Ksiazek et al. 2018; Kishimoto and Sugimura 2010). The niobium-coated boron-doped diamond anode and titanium cathode could remove 57% of the chemical oxygen demand and 12.3% of the ammoniacal nitrogen at pH of 8.1, voltage of 9.8 V, current intensity of 3 A, and reaction time of 41 h (Zolfaghari et al. 2016). With the pretreated landfill leachate (membrane biological reactor), the niobium-coated boron-doped diamond could remove 84% of the chemical oxygen demand, and it is also reported that ammoniacal nitrogen concentration increased after the anodic oxidation. The increase in ammoniacal nitrogen happens due to the conversion of total nitrogen to ammoniacal nitrogen in the near-neutral pH oxidation. The scavenging of hydroxyl radicals by carbonates and bicarbonates happens also in the niobium-coated boron-doped diamond electrode (Zhou et al. 2016; Zolfaghari et al. 2016). The energy consumption for the removal of organic matter varies from 70 to 193 KWh/kg chemical oxygen demand for different electrodes (Bashir et al. 2013; Panizza and Martinez-Huitle 2013; Urtiaga et al. 2009; Zolfaghari et al. 2016). Boron-doped diamond electrode performs better than dimensionally stable anodes in the oxidation of leachate organic matter, with titanium as the cathode. Boron-doped diamond electrode system removed 81% of color, almost 100% ammoniacal nitrogen and 50% chemical oxygen demand at pH 7.5 and 1.5 Ah/L electric charge (Ding et al. 2018). When compared to TiRuSnO2 and PbO2 electrodes, boron-doped diamond anodes are more efficient for the oxidative removal of organic matter from landfill leachate (Panizza and Martinez-Huitle 2013). Also, titanium coated with RuO2, IrO2, or PtO and boron-doped diamond electrodes are best for the electrochemical oxidation of leachate because of its catalyst property and resistance toward corrosion (Ding et al. 2018; Shestakova and Sillanpää 2017). Landfill leachate contains heavy metals which is enough for the activation of hydrogen peroxide. At the same time, the results reported by Venu et al. (2016) indicate that external addition of ferrous ion in electro-Fenton process is essential for the effective oxidation of organic pollutants present in landfill leachate. ElectroFenton process using boron-doped diamond anode and carbon felt cathode is effective for the treatment of landfill leachate as well as the removal of micropollutants and humic substances (Oturan et al. 2015). Almost complete degradation of volatile organic compounds, polycyclic aromatic hydrocarbons, polychlorobiphenyls, organochlorine pesticides, polybrominated diphenyl ethers, and alkylphenols from leachate was observed by the authors. Combined electroFenton process and biological treatment is a feasible option to treat stabilized landfill leachate as reported by Baiju et al. (2018). The authors used iron molybdophosphate

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as a source for iron and observed an enhancement in biodegradability (0.03 to 0.4) of landfill leachate after electro-Fenton treatment. Thus, hybrid electro-Fenton and biological treatment removed 97% of chemical oxygen demand from the wastewater and reduced its value from 7184 mg/L to 192 mg/L. Similar to this, iron–manganese binary oxide-loaded zeolite is also found as a good heterogeneous electro-Fenton catalyst for treating stabilized landfill leachate (Sruthi et al. 2018). Peroxi-coagulation process is also a promising technology for the treatment of landfill leachate. Stainless steel is found more effective than iron as anode material in peroxi-coagulation process (Venu et al. 2014). In acidic pH conditions, electroFenton process is the dominant pollutant removal mechanism of peroxi-coagulation process, while electrocoagulation process predominates at neutral as well as alkaline conditions (Venu et al. 2016). Increase in solution pH with electrolysis time is a characteristic of peroxi-coagulation process. Therefore, pH regulation is essential for making the electro-Fenton process as predominant treatment process in peroxicoagulation process, as observed by Venu et al. (2016). Solar photo-electro-Fenton and UV photo-electro-Fenton processes are found very much applicable for the treatment of stabilized landfill leachate (Moreira et al. 2016). These processes are more active than the electro-Fenton process, photo-Fenton process, as well as anodic oxidation process (Moreira et al. 2015, 2016). In acidic conditions, the addition of oxalate ions improved its performances significantly (Moreira et al. 2015). Seibert and co-workers (Seibert et al. 2019) found less toxic by-products than the parent compound present in landfill leachate after oxidation with the photo-electro-Fenton process. The authors also observed a significant improvement in biodegradability after the treatment.

4.3.5

Paper Mill Wastewater

The electrochemical anodic oxidation with anionic membrane works better for the treatment of highly concentrated, highly alkaline paper and pulp industry wastewater (Chanworrawoot and Hunsom 2012). The Ti/RuO2 grid plate anode and stainless steel cathode degrade the organic matter present in the wastewater. During electrolysis, the in situ generation of H+ at the anode surface and the generation of OH at the cathode surface help in the removal of pollutants from wastewater. The H+ generated in the anode surface reacts with lignin, tannin, and other inorganic compounds in the paper wastewater and results in precipitation (Chanworrawoot and Hunsom 2012). The hydroxyl compound precipitates the metallic compound present in the wastewater and also leads to the adsorption of organic matters in the wastewater, reducing the color, biological oxygen demand, and chemical oxygen demand. The active oxygen generation at the anode surface leads to the formation of highly oxidizing oxidant hydroxyl radical. At higher pH condition, the generated oxidant OH● forms physically adsorbed active oxygen and chemically active oxygen and decomposes the organic compounds (Chanworrawoot and Hunsom 2012).

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In the electrochemical oxidation membrane reactor, the increase in circulation rate increases the contact time of the wastewater with the active oxidizing species and thus improves the removal efficiency. The current density also has a similar effect on the abatement of pollutants from paper mill wastewater. The increase in current density increases the removal efficiency of biological oxygen demand, chemical oxygen demand, and color up to 90%, 83%, and 90%, respectively, within the limit up to 3 mA/cm2, and a further increase in current density decreases the removal efficiency of chemical oxygen demand, biological oxygen demand, and color. The increase in current density increases the temperature of the electrolysis reactor and leads to the dissolution of the precipitate, and this behavior can be observed in paper mill wastewater as the pH is higher than 12 (Chanworrawoot and Hunsom 2012). In contrast to the aforesaid for boron-doped diamond as the anode, the increase in current density increased the removal of chemical oxygen demand up to 80% in paper mill wastewater from 0.25 to 0.5 A, and then with further increase in the current density, the removal efficiency remains the same (Klidi et al. 2019). The current efficiency decreased considerably with increase in current density compared to the low current density, the fact that boron-doped diamond electrode could generate the required amount of hydroxyl radical at low current density itself. However, in the case of TiRuSnO2 anode, only 60% of the removal happens at highest current density of 1.5 A, which means the generation rate of active hydroxyl radicals for the decomposition of organic matters is very less compared to borondoped diamond electrode. In the case of TiRuSnO2 anode, higher oxidation complex like MOx + 1 can be observed when compared to the MOx complexes in boron-doped diamond (Comninellis 1994; Klidi et al. 2019). The higher oxidation complex compounds are less effective in organic oxidation as the compounds lead to the evolution of oxygen. The removal of chemical oxygen demand increased with increasing the current density with TiRuSnO2 electrode. It indicates that the secondary oxidation is the active mechanism for the removal of organic matters. The generation of HClO and ClO are the active species for the oxidation of organic matter in combination with hydroxyl radical (Klidi et al. 2018; Ltaïef et al. 2018; Vazquez-Gomez et al. 2006), as the TiRuSnO2 anode well behaves as a catalyst. The increase in temperature also results in the generation of foresaid oxidizing species like HClO and ClO in borondoped diamond anode (Klidi et al. 2018; Marselli et al. 2003; Serrano et al. 2002). The presence of NaCl and boron-doped diamond anode converts the chloride to active chlorine and hypochlorite. The hypochlorite gets oxidized by hydroxyl radicals to chlorate and chlorate to perchlorate. At the same condition, TiRuSnO2 electrode oxidizes the hypochlorite to chlorate, and further oxidation to perchlorate is negligible (Klidi et al. 2018). At high pH condition, electrolytic oxidation turns out to be added with precipitation, thus the increase in electrolysis time increases the dissolved solids and suspended solids in the reactor. The increase in dissolved solids is due to the increased dissolution, and increase in suspended solids is due to the increase on destructed organic matters (Chanworrawoot and Hunsom 2012).

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The addition of NaCl can generate the supportive species OCl which can improve the removal of organic matter from paper mill wastewater. At higher pH, OCl generation rate can be influenced by the current density and the presence of chloride ions in the system. The OCl generation occurs at high current density or after at least 1 h of electrolysis at the low current density conditions. The generated OCl converts the organic matter into CO2 and H2O completely (Chen 2004; Rajkumar et al. 2005; Zhang et al. 2006), thus increasing the removal of chemical oxygen demand, biological oxygen demand, and color above 95%. During electrolysis, the chloride content at high pH also leads to the formation of toxic organochlorinated by-products (Chanworrawoot and Hunsom 2012). In contrast to Ti/RuO2 anode behavior, boron-doped diamond electrode as the anode has no impact on the removal efficiency of organic matter from paper mill wastewater with a wide range pH from 3 to 11. The oxidation of organic matter happens with electrogenerated hydroxyl radicals which are very strong oxidants with high potential and is not affected by pH (García-Orozco et al. 2016; Klidi et al. 2018). About 85% of the chemical oxygen demand gets removed for the wide range of pH from 3 to 11 with boron-doped diamond electrode. On the other side, TiRuSnO2 oxidation of organic matter gets highly related to the pH of the solution. The oxidation of organic matters happens with the active species, HOCl, ClO, and ●OH, at pH 3, HOCl and ●OH being the dominant species, which has high oxidation potential when compared to OCl. Thus, the removal efficiency is high at the acidic pH compared to alkaline pH with TiRuSnO2 anode (Deborde and von Gunten 2008; Klidi et al. 2018; MartínezHuitle et al. 2005; Scialdone et al. 2008). Even titanium anode and stainless steel cathode give 80% of the chemical oxygen demand from the paper mill wastewater at 2.5 A/dm2 with NaCl of supporting electrode (Soloman et al. 2009). The lead-coated anode and stainless steel cathode could remove 97% of the chemical oxygen demand (initial chemical oxygen demand ~6000 mg/L) at 6.6 mA/cm2 current density and in the presence of NaCl at pH 8 and 60 min of electrolysis (El-Ashtoukhy et al. 2009). The electro-Fenton process is an efficient process for the removal of organic matter from the paper mill wastewater. The experiments with Ti/IrO2–Ta2O5 anode and carbon felt, modified carbon felt, and gas diffusion electrode cathode show very less removal of organic matter (13%, 15%, and 17%, respectively, at pH 3 and current density of 20 mA/cm2) as the total organic carbon in the absence of Fe2+ (Klidi et al. 2019). The organic matters get oxidized directly at the anode surface and indirectly by the cathode which generated H2O2 (Ma et al. 2019). Due to its high porosity, gas diffusion electrode allows the oxidation of O2 to H2O2 when compared to modified carbon felt and carbon felt (Brillas et al. 1995, 2009; Harrington 1999; Pérez et al. 2017; Pozzo et al. 2005). The addition of Fe2+ increased the removal efficiency from 19% to 41% with Ti/IrO2–Ta2O5 anode and gas diffusion electrode at pH 3. The consideration of the molar ratio of hydroxyl radical for the decomposition of organic matter is a concern, and beyond the molar ratio, the generated hydroxyl radical leads to scavenging (Klidi et al. 2019; Ma et al. 2019). When compared to Ti/IrO2–Ta2O5 anode, boron-doped diamond anode decomposes high concentration of organic matters and leads to higher removal efficiency. The borondoped diamond electrode with modified carbon felt at pH 3 removed 80% of the chemical oxygen demand from paper mill wastewater. The metal oxide anodes

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generate higher oxide chemisorbed hydroxyl radicals when the boron-doped diamond anodes generate physisorbed hydroxyl radicals. The physisorbed radicals combust the organic matter, leading to higher removal efficiency with borondoped diamond anodes (Comninellis 1994; Klidi et al. 2019). As in anodic oxidation, the electro-Fenton process also converts the chloride ions into chlorine, hypochlorous acid, and/or hypochlorite which depend on the pH of the solution, leading to the increase in removal efficiency (Klidi et al. 2018; Ma et al. 2018; Martínez-Huitle et al. 2015). The graphite-graphite electrochemical reactor in the presence of catalyst material Mo–phosphate-modified kaolin loaded with Fe3+ removed 95% of the chemical oxygen demand from paper mill wastewater at a current density of 40 mA/cm2, pH of 4, and 40 min of electrolysis (Ma et al. 2007). In the same reactor, Mo–phosphate-modified kaolin, Mo–phosphate-modified kaolin loaded with Fe3+, and Mo modified with kaolin give 51%, 76%, and 35%, respectively, chemical oxygen demand removal at pH 7 and electrolysis time of 40 min. The addition of phosphate along with Mo in the kaolin replaces the OH compound by PO43, thus modifying the structure of kaolin, which accelerates the removal of organic matter (Ma et al. 2007; Yang and Evmiridis 1994). The loading of Fe3+ on the Mo–phosphate-modified kaolin undergoes Fenton-like process in the reactor and generates hydroxyl radical (Ma et al. 2007; Walling 1975). Also, the presence of Fe3+ in the reactor leads to the generation of flocs at the near-neutral pH, and thus the organic matter gets removed (Can and Kobya 2006; Liu and Jiang 2005). pH plays a crucial role during the degradation of organics in graphite-graphite electrochemical reactor in the presence of Mo–phosphate-modified kaolin loaded with Fe3+ as a catalyst. When compared to pH 8 and 6, pH 4 gives higher removal of paper mill industrial wastewater with 96% removal efficiency (Ma et al. 2007). The generation of hydroxyl radicals is high at acidic pH when compared to that of the near-neutral and alkaline pH (Liou et al. 2005; Liu and Jiang 2005). Also, when compared to the different metal loading on Mo–P-modified kaolin, Fe3+ gives higher removal efficiency compared to Co2+ and Cu2+ (Ma et al. 2007).

4.3.6

Pharmaceutical Industry Effluent

The pharmaceutical effluents are composed of large varieties of compounds such as synthetic or natural organics, inorganics, catalysts, and solvents; thus, the pharmaceutical wastewater contains a wide spectrum of substance to become highly toxic and characterized by high chemical oxygen demand and biological oxygen demand and very low biological oxygen demand/chemical oxygen demand ratio (Panizza 2018). Hence, in order to avoid the negative impacts on human and other animals by the improper discharge of pharmaceutical effluent, some highly advanced treatment processes are necessary. Domínguez et al. (2012) have treated real pharmaceutical effluent from a pharmaceutical industry in Spain through electrooxidation. The main compositions of the

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effluent were aliphatic and aromatic organic compounds, and solvents like ethanol and methanol contribute to total organic carbon/chemical oxygen demand ratio around 0.27 and the chemical oxygen demand around 12,000 mg dm3. The anodic oxidation conducted in batch mode uses boron-doped diamond electrode as the anode and stainless steel as the cathode. The statistical significance of the operating conditions, namely, current density and flow rate at different residence times, is studied by adopting a factorial central composite orthogonal and rotatable design by varying the flow rate from 104.8 to 564.7 cm3 min 1 and current density from 25.7 to 179.4 mA cm2, at residence times from 0 to 570 min. The degradation experiments were conducted by analyzing the variations in total organic carbon. They found that at and above the residence time of 125 min, the influence of both operating parameters tends to diminish and becomes completely decided by residence time. The influence of individual parameters is optimum at around 75 min, where the current density showed the highest positive influence on total organic carbon removal, while flow rate presented a negative influence. By combining both variables adequately, almost 100% of the total organic carbon removal is achievable. The more reliable and efficient pollutant abatement by conductive diamond electrooxidation as compared to Fenton oxidation was proved by Perez et al. (2017) by treating 60 real effluents from different processes in an organic-synthesis pharmaceutical industry. Conductive diamond electrooxidation is found to be more effective than Fenton oxidation in 80% of the samples in terms of chemical oxygen demand removal efficiency, where the initial chemical oxygen demand values were in the range of 2.0 mgdm3 to 100 mgdm3. Conductive diamond electrooxidation analysis was conducted in batch operation mode in single-compartment electrolytic cell under galvanostatic conditions at 30 mA/cm2 and 25  C by using p–Si– boron-doped diamond anode and stainless steel cathode. Both direct and mediated oxidation mechanisms involved in pollutant degradation in the case of conductive diamond electrooxidation might be the reason behind the improved efficiency than Fenton oxidation. Sono- or photo-irradiated conductive diamond electrooxidation can give enhanced oxidation as well as better mineralization of the organic compounds in the industrial real effluent due to the synergic action. This was proved by the treatment of effluent from an organic-synthesis pharmaceutical plant (Martín de Vidales et al. 2017). The electrolysis experiments were conducted in a batch operation mode using a single-compartment electrochemical flow cell. The irradiation is provided either by UVC in the wavelength of 254 nm or by ultrasonic generator emitting at 24 kHz and a maximum ultrasonic power of 200 W (Martín de Vidales et al. 2017). Xia et al. (2019) proposed a modified electrochemical Fenton method to treat the real metronidazole wastewater. This process has advantages of moderate cost and simple operation than Fenton process as a pair of iron electrodes were used in which Fe anode is electro-dissolved supplying stoichiometric Fe2+, and then Fe2+ reacts with H2O2 to generate hydroxyl radical. Under the operational conditions of electrolysis time of 5 min, current density of 15 mA/cm2, 15 mmol/L of H2O2, and pH of 3, biological oxygen demand/chemical oxygen demand ratio improved from 0.17 to 0.35.

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Rajkumar and Palanivelu (2004) employed an electrochemical method for the treatment of pharmaceutical industry effluent possessing phenolic compounds. A mixed triple oxide electrode of titanium electrode coated with RuO2–IrO2 with an effective surface area of 27.7cm2 was used as the anode and graphite carbon as the cathode. The wastewater from bulk drug manufacturing industries with initial chemical oxygen demand of 8879 mg/L was treated electrochemically under the operating conditions of initial pH 9.0, chloride 2500 mg/L, and current density 5.4A/ dm2 to get the total organic carbon and chemical oxygen demand removals around 53.3% and 73.1%, respectively, after the passage of 12 Ah/L charge.

4.3.7

Oil Mill Industry Wastewater

Vegetable oil industry effluents are the acidic aqueous complex of a broad spectrum of organics and inorganics in suspended and dissolved forms. The soluble and stable emulsified organic matter, which the traditional treatment system fails to remove, results in the high biological oxygen demand5 and chemical oxygen demand values (Verla et al. 2014). Based on the type of oil processed and the operating conditions, the amount and characteristics of organic load vary and the chemical oxygen demand values lie in the range 2000 mg/L–30,000 mg/L (Pandey et al. 2003; Sharma and Simsek 2019; Verla et al. 2014). Cañizares et al. (2006) proved that the refractory organics in olive oil mill that cannot be further oxidized by Fenton process can be successfully treated by using electrochemical oxidation using boron-doped diamond electrodes. They used actual industrial wastewater effluent with a residual chemical oxygen demand of around 700 mg dm3 from a wastewater treatment plant. Experiments were carried out with diamond-based material as the anode and stainless steel as the cathode in a singlecompartment electrolytic cell. The increased mineralization efficiency attained in the presence of inorganic salts confirmed the importance of mediated oxidation in boron-doped diamond-based electrooxidation (Cañizares et al. 2006). The authors also compared the treatment efficiency of conductive diamond electrooxidation, ozonation, and Fenton oxidation by using olive oil mill effluent from Spain with an organic load around 840 mg dm3 of the total organic carbon and 3000 mg dm3 of chemical oxygen demand composed of a variety of aliphatic and aromatic compounds. Among the three advanced oxidation processes, only conductive diamond electrooxidation showed the complete mineralization of the waste with high efficiency, as this process combines direct oxidation and mediated oxidation mechanisms. Also, the energy required for the process is found to be lower than that of ozonation. Conductive diamond electrooxidation was done in batch mode in a single-compartment electrolytic cell with stainless steel cathode and boron-doped diamond anode working in natural pH; T, 25  C; and j, 30 mA cm2, (Cañizares et al. 2007). Rodrigo et al. (2010) also found the robust, efficient, and nonselective activity of conductive diamond electrooxidation toward olive oil mill effluent treatment as

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compared to other anodic materials such as lead dioxide (Pb/PbO2), graphite, and dimensional stable anodes. They concluded that the anodic oxidation of industrial wastewater is strongly influenced by the anodic type, organic load, and current density rather than the nature of pollutant by analyzing the pollutant abatement in pharmaceutical, petrochemical, olive oil mill, chemical, and door manufacturing wastewaters. Anodic oxidation in a modified Grignard reactor using platinized expanded titanium electrodes is also effectively detoxifying the olive mill wastewater with a higher organic load of 65 g/L chemical oxygen demand and 25 g/L of the total organic carbon (Belaid et al. 2013). The process is able to achieve the complete removal of all phenolic compounds, 87% of color, and 55% of chemical oxygen demand and total organic carbon (Belaid et al. 2013). Gotsi et al. (2005) investigated the anodic oxidation of olive oil mill wastewaters using a titanium–tantalum– platinum–iridium anode. They found nearly complete degradation of phenols as well as color removal within 60 min, but relatively low mineralization is observed. Electrochemical oxidation can also be used as a posttreatment technique to anaerobic digestion. Gonçalves et al. (2012) electrochemically treated the anaerobically digested olive oil mill effluent over dimensionally stable anodes and found the suitability of RuO2-based anode over IrO2-type DSA for the efficient and complete mineralization. Olive oil mill wastewater with initial chemical oxygen demand of 92,000 mgL1 was completely mineralized by adopting electro-Fenton process (Bellakhal et al. 2006). Electrolysis was carried out using carbon felt cathode and Pt sheet anode where Fe2+ ions were the catalyst at pH of 3 and current density of 8.3 mAcm2 (Bellakhal et al. 2006). Khoufi et al. (2006) studied the applicability of electroFenton process as a pretreatment to biological digestion for olive oil mill wastewater. The electro-Fenton reactor was formed by cast iron plates as anode and cathode with an effective surface area of 0.2 dm2. The optimum H2O2 concentration and current density were found to be 1 g L1 and 7.5 Adm2, respectively, for the removal of 65.8% of the total polyphenolic compounds and toxicity reduction up to 66.9%. Flores et al. (2018) coupled electrocoagulation with electro-Fenton or photoelectro-Fenton in sequential order and evaluated their activity for olive oil mill wastewater treatment. The electro-Fenton and photo-electro-Fenton are conducted in an electrolytic cell with boron-doped diamond anode and a carbon– polytetrafluoroethylene air diffusion cathode with an inter-electrode gap kept at 1 cm. For photo-electro-Fenton process, UVA of λ max of 360 nm was provided. The sequential use of electrocoagulation/electro-Fenton and electrocoagulation/ photo-electro-Fenton is very efficient as compared to single processes as more than 97% of the total organic carbon drop was identified in electrocoagulation/ photo-electro-Fenton process. At neutral pH, both electrocoagulation/photoelectro-Fenton and electrocoagulation/electro-Fenton performed in a similar manner. However, at acidic pH, electrocoagulation/photo-electro-Fenton outperforms electrocoagulation/electro-Fenton. Mostafa et al. (2018) combined high-power ultrasound and electro-Fenton process to treat olive mill wastewater. Pretreatment using high-power ultrasound for electro-Fenton improved the performance as direct ultrasonication for 90 min

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followed by electro-Fenton for 4 h showed chemical oxygen demand reduction up to 75% of olive mill effluent with initial chemical oxygen demand of 19,800 mg/L. Electrocoagulation–peroxidation can be employed as a posttreatment technique for polishing biologically treated palm oil mill effluent before final discharge. The optimized conditions are found to be 40.21 mA/cm2 of current density, 4.4 of initial pH, and 0.5 g/L of H2O2 to achieve removal of 71.3% of chemical oxygen demand, 96.8% color, and 100% total suspended solid within 45 min of contact time. A significant reduction in current density and reaction time was observed when utilizing electrocoagulation–peroxidation as compared to electrocoagulation alone (Bashir et al. 2019).

4.4

Conclusions and Perspectives

Electrochemical advanced oxidation processes were proved to be the promising technology for the treatment of a wide range of real wastewaters. Fast and efficient treatment can be achieved through the optimization of the operational parameters. The process efficiency mainly depends on the number of reactive oxygen species especially hydroxyl radicals generated during the process. Hence, hybrid processes are proved to be much effective than individual processes such as solar photoelectro-Fenton/photo-electro-Fenton. These processes are much effective than the electro-Fenton and UV-based processes alone. Moreover, there are challenges that exist in the electrogeneration of reactive oxygen species in high quantity during the continuous process. Boron-doped diamond anode-based anodic oxidation had been used by many researchers for the effective treatment of the contaminants. However, low-cost, scalable anode materials need to be developed for the real field applications. At the most, the economic feasibility of the electrochemical advanced oxidation processes is the crucial factor for the large-scale operations and the real field implementation. The investment cost including the cost of the reactor, electrodes, and lamps and the operational cost including the electricity consumption for electrochemical, photochemical, and sonolysis process, chemicals used, and maintenance need to be considered. Acknowledgment The authors are thankful to the Director of the Indian Institute of Technology Madras, Chennai; the Director of CSIR-NEERI, Nagpur; and the Director of Pandit Deendayal Petroleum University, Gandhinagar, Gujarat, India, for providing encouragement and kind permission for publishing the article.

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Venu D, Gandhimathi R, Nidheesh PV, Ramesh ST (2016) Effect of solution pH on leachate treatment mechanism of peroxicoagulation process. J Hazard Toxic Radioact Waste 20:4–7. https://doi.org/10.1061/(ASCE)HZ.2153-5515.0000315 Verla AW, Verla EN, Adowei P, Briggs A, Horsfall M (2014) Quality assessment of vegetable oil industry effluents in Port Harcourt, Rivers State, Nigeria. Int Lett Chem Phys Astron 33:179–189. https://doi.org/10.18052/www.scipress.com/ilcpa.33.179 Walling C (1975) Fenton’s reagent revisited. Acc Chem Res 8:125–131. https://doi.org/10.1021/ ar50088a003 Wang C-T, Hu J-L, Chou W-L, Kuo Y-M (2008) Removal of color from real dyeing wastewater by electro-Fenton technology using a three-dimensional graphite cathode. J Hazard Mater 152:601–606. https://doi.org/10.1016/j.jhazmat.2007.07.023 Wang CT, Chou WL, Chung MH, Kuo YM (2010) COD removal from real dyeing wastewater by electro-Fenton technology using an activated carbon fiber cathode. Desalination 253:129–134. https://doi.org/10.1016/j.desal.2009.11.020 Webler AD, Moreira FC, Dezotti MWC, Mahler CF, Segundo IDB, Boaventura RAR, Vilar VJP (2019) Development of an integrated treatment strategy for a leather tannery landfill leachate. Waste Manag 89:114–128. https://doi.org/10.1016/j.wasman.2019.03.066 Xia Y, Zhang Q, Li G, Tu X, Zhou Y, Hu X (2019) Biodegradability enhancement of real antibiotic metronidazole wastewater by a modified electrochemical Fenton. J Taiwan Inst Chem Eng 96:256–263. https://doi.org/10.1016/j.jtice.2018.11.019 Yang S, Evmiridis NP (1994) Synthesis of omega zeolite without use of Tetramethylammonium (TMA) ions. Stud Surf Sci Catal 84:155–162. https://doi.org/10.1016/S0167-2991(08)64109-9 Zhang J, Zheng Z, Luan J, Yang G, Song W, Zhong Y, Xie Z (2007) Degradation of hexachlorobenzene by electron beam irradiation. J Hazard Mater 142:431–436. https://doi. org/10.1016/j.jhazmat.2006.08.035 Zhang Q, Kang B, Xu H, Lin HB (2006) Indirect electrochemical oxidation of 4-Amino-dimethylaniline hydrochloride. Chem Res Chin Univ 22:360–363. https://doi.org/10.1016/S1005-9040 (06)60116-5 Zhou B, Yu Z, Wei Q, Long H, Xie Y, Wang Y (2016) Electrochemical oxidation of biological pretreated and membrane separated landfill leachate concentrates on boron doped diamond anode. Appl Surf Sci 377:406–415. https://doi.org/10.1016/j.apsusc.2016.03.045 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 Zou J, Peng X, Li M, Xiong Y, Wang B, Dong F, Wang B (2017) Electrochemical oxidation of COD from real textile wastewaters: kinetic study and energy consumption. Chemosphere 171:332–338. https://doi.org/10.1016/j.chemosphere.2016.12.065

Chapter 5

Unconventional Adsorbents for Remediation of Metal Pollution in Waters Md. Mostafizur Rahman

, Rubaiya Akter, and Mashura Shammi

Contents 5.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2 Adsorbent Classification . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.3 Unconventional Adsorbents for Trace Metal Removal from Water . . . . . . . . . . . . . . . . . . . . . 5.3.1 Arsenic Removal Using Unconventional Adsorbents . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.3.2 Cadmium Removal Using Unconventional Adsorbents . . . . . . . . . . . . . . . . . . . . . . . . . . 5.3.3 Chromium Removal Using Unconventional Adsorbents . . . . . . . . . . . . . . . . . . . . . . . . . 5.3.4 Lead Removal Using Unconventional Adsorbents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.4 Mechanisms of Adsorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.5 Concluding Remarks . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

124 125 126 126 130 133 134 135 137 138

Abstract Environment conservation through cleaning up of xenobiotics is a global concern. Different technologies have been tried to remove pollutants from the water environment. Among them adsorption is one of the prime candidates. In this chapter, we have discussed the removal of heavy metals: arsenic, lead, cadmium, and chromium using unconventional low-cost novel sorbents, e.g., waste materials, biochar, industrial wastes, nanomaterials, and metal–organic frameworks. We have majorly focused on the introduction of unconventional adsorbents with their maximum adsorption capacity to their target metals from an aqueous environment. Besides the commercial adsorbents such as activated carbon, the unconventional adsorbents showed promising capability to remove metals from water. However, the holistic approach of the multidisciplinary involvement is needed to make these unconventional materials an industrial scale adsorbent to clean up the metals at the source/discharge points.

Md. M. Rahman · R. Akter · M. Shammi (*) Department of Environmental Sciences, Jahangirnagar University, Dhaka, Bangladesh e-mail: [email protected]; [email protected] © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature Switzerland AG 2021 Inamuddin et al. (eds.), Water Pollution and Remediation: Heavy Metals, Environmental Chemistry for a Sustainable World 53, https://doi.org/10.1007/978-3-030-52421-0_5

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Keywords Arsenic · Chromium · Lead · Cadmium · Nanocomposite · Waste · Biochar and adsorption

5.1

Introduction

Cleaning of polluted environment might be one of the greatest global challenges for the next century. Now, pollution has already become a burden for the earth systems. Thousands of synthetic chemicals including metals, organics, dyes, and pharmaceuticals are being discharged into the environment. Most of them have proven human health consequences as well as environmental health outcomes. A great deal of efforts has been taking worldwide to remove pollutants from the environment through technology innovation, and thus recent days are being called as Pollutant Removal Age (Morin-Crini and Crini 2017). Most of the pollutants are finally drained into water bodies. Industrial sectors are the prime contributor to environmental pollution. This sector needs a huge volume of raw water, and after different production processes, the water usually gets polluted with numerous chemicals both inorganic and organics. Thus, industrial effluents are the major contributors to water pollution all over the world. The treatment technologies have been designing based on the end-of-pipe measurement. So far, the trace metals pollution from industrial discharge gained a lot of scientific attention in the remediation point of views. However, the industrial sector has been considering one of the most polluting sectors, although huge efforts were carried out to make the production chains free from pollution for the last three decades (Morin-Crini and Crini 2017). Trace metals are listed as priority pollutants and have already reported for the human health consequences including different forms of cancer, autism spectrum disorders, and other diseases (Rahman et al. 2019; Geir et al. 2018). Continuous efforts have been taking all over the world to find suitable remediation technologies to remove metals from the water including chemical precipitation, sorption, ion exchange, filtration, nanofiltration, phytoremediation, and electrochemical techniques (Sikder et al. 2018). Besides the trace metals, different types of organics including waste organic chemicals, pharmaceutical residues, and chemical additives are of special concern due to their potential deleterious health effects. There are various interfering factors which complicate pollution remediation processes such as concentration, composition, potential hazards, toxicity, cost, time, and treatment levels. However, for the last couple of decades, research on adsorption technologies achieved considerable attention for remediation of water pollution. Moreover, industrial processes also use adsorption techniques in purification, decolorization, detoxification, and separation (Crini and Badot 2010).

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So far, adsorption is considered as the most facile and economically viable option for the water pollution remediation in an industrial scale. For instance, activated carbons are considered as an effective adsorbent. The capacity of activated carbon in the removal of major pollutants including metals, pharmaceuticals, and organics has been widely examined, and further, commercial activated carbon and commercial activated alumina showed tremendous potentials in the water purification. Although the performance is high, the feasibility due to high cost of these conventional adsorbents is not being used commonly. These limitations have driven the interest to search for alternative techniques to purify water pollution such as unconventional adsorbents. This chapter particularly focuses on the unconventional adsorbents in the treatment of metal pollution from water. Moreover, aside from metal adsorption mechanisms, this chapter also demonstrates the advancement of unconventional adsorbents with their classifications.

5.2

Adsorbent Classification

There are a couple of classification systems for adsorbent materials depending on variant parameters such as chemical forms and materials surface geometry. In general, Crini et al. (2018) mention that the solid adsorbent materials are classified in the following five classes: 1. Natural materials; those are directly collected from the nature with/very slight modification. 2. Modified natural materials to incorporate new properties and structure. 3. Fully synthesized adsorbents. 4. Waste and by-product-based materials from both agricultural and industrial sources. 5. Biomaterials. However, in this chapter, the classification by Crini (2005, 2006; Crini and Badot 2007) has been considered for the easiness of discussion and interpretation. According to them adsorbents are simply classified as (1) conventional adsorbent and (2) unconventional adsorbent. The classified list of conventional adsorbent and unconventional adsorbent are presented in Fig. 5.1.

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Fig. 5.1 Conventional and non-conventional adsorbents for the removal of pollutants from wastewaters modified after (Crini 2005, 2006; Crini and Badot 2007). Here, different sources of unconventional adsorbents are listed with specific examples along with the commercially available adsorbents

5.3

Unconventional Adsorbents for Trace Metal Removal from Water

The recent advancement in the development of unconventional adsorbents is discussed in this section. Prior to starting of the evidence-based presentation in the following sections a major list of the unconventional adsorbents capable to remove metals and dyes from aqueous phase are listed in Table 5.1.

5.3.1

Arsenic Removal Using Unconventional Adsorbents

Arsenic is one of the most notorious candidates among the metals. It is responsible for numerous adverse effects on human health including cancers and neurological diseases (Rahman et al. 2019). It is mostly found in the natural environment although anthropogenic activities such as mining wastes, petroleum refining, sewage sludge, agricultural chemicals, ceramic manufacturing industries, and coal fly ash contribute to arsenic levels in both surface and groundwater. From the last couple of decades, scientists have been looking for a suitable arsenic removal technology from the water. Arsenic has four oxidation states, and this speciation is very crucial for their

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Table 5.1 List of unconventional adsorbents which can remove trace metals and dyes from the water environment Unconventional adsorbent Agricultural by-products Agri-food wastes Fruit and vegetable wastes Sawdust Bark Rice husk Wheat Sugar-beet pulp Coconut Opuntia ficusindica Coffee Tea factory waste Peat Lignocellulosic wastes Cellulose Chitin Chitosan Cyclodextrins Industrial by-products Fly ash Municipal wastes Sludge-based adsorbents Nanomaterials

References Oliveira and Franca (2008), Crini and Badot (2010), Sharma et al. (2011), Nguyen et al. (2013), Rangabhashiyam et al. (2014), Lim and Aris (2014), Kharat (2015), Zhou et al. (2015), and Sulyman et al. (2017) Demirbas (2008), Oliveira and Franca (2008), and Kumar et al. (2011) Swami and Buddhi (2006) and Patel (2012) Shukla et al. (2002), Larous and Meniai (2012), Kharat (2015), and Sahmoune and Yeddou (2016) Ahmaruzzaman (2008), Demirbas (2008), Kharat (2015), and Sen et al. (2015) Chuah et al. (2005), Ahmaruzzaman (2011), Nguyen et al. (2013), Dhir (2014), and Sulyman et al. (2017) Ngah and Hanafiah (2008) and Farooq et al. (2010) Ngah and Hanafiah (2008), Ahmaruzzaman (2011), and Dhir (2014) Swami and Buddhi (2006), Patel (2012) Nharingo and Moyo (2016) Anastopoulos et al. (2017) and Sulyman et al. (2017) Ahmaruzzaman (2011) and Sulyman et al. (2017) Vohla et al. (2011), and Raval et al. (2016) Miretzky and Cirelli (2010), Abdolali et al. (2014), and De Quadros Melo et al. (2016) O’Connell et al. (2008), Vandenbossche et al. (2015), and Grishkewich et al. (2017) Yong et al. (2015), Barbusinski et al. (2016), and Anastopoulos et al. (2017) Crini and Badot (2008), Vakili et al. (2014), Yong et al. (2015), Barbusinski et al. (2016), and Azarova et al. (2016) Crini (2005, 2014), Crini and Badot (2010), and Panic et al. (2013) Crini (2005, 2006) and El-Sayed and El-Sayed (2014) Swami and Buddhi (2006), Ahmaruzzaman (2008, 2011), and Raval et al. (2016) Bhatnagar and Sillanpää (2010) Raval et al. (2016) and Devi and Saroha (2017) Crini and Badot (2010), Ali (2016), Kumar et al. (2011), Zhao and Zhou (2016), and Sadegh et al. (2017)

Modified from Crini et al. (2018)

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removal from the water phase. In this section, we have discussed the recent developments of arsenic removal techniques through unconventional adsorbents including nanomaterial-based adsorbents although there are a lot of other methods such as pre-oxidation, ion exchange, separation, and precipitation (Baig et al. 2014). Recently, by-products of different manufacturing industries have been evaluated for arsenic sorption properties. For example, Ociński et al. (2016) found water treatment residuals; a by-product during the deironing and demanganization process of infiltration water showed excellent adsorption properties to remove arsenic (III) (132 mg/g) and arsenic (V) 77 mg/g from the water phase. Water treatment residuals have very high surface area 120 m2/g. Using by-product or waste material as adsorbents has multiple benefits in pollution remediation such as reduction of wastes along with pollutants removal. Biochar has been using as an adsorbent to remove metals from water effectively. Biochar can be defined as a pyrogenic carbon material based on biomass. The processes of preparing the biochar are mainly done using biomass combustion process in an anoxic system. It has charismatic features, for example, high surface area and cation exchange capacity. These properties further allow the application of biochar in soil amendment, soil fertility restoration, and carbon sequestration (Mohana et al. 2014). Previous research showed the capacity of biochar to reduce metals from water environment as well as they also played crucial role in the reduction of metals bioavailability and their mobility in the earth system (Ahmad et al. 2014). The sorption of aqueous arsenic (in anionic forms) of either arsenate (V) or arsenite (III) is relatively low (Beesley and Marmiroli 2011) because most of the surfaces of biochars are predominantly net negatively charged (Mukherjee et al. 2011). Wang et al. (2015a, b) developed magnetic biochar from pinewood and hematite minerals. They found that modified biochar with magnetic properties has higher arsenic removal efficiency than the unmodified one (Wang et al. 2015a, b). Recently, a novel Fe–Mn binary oxide sorbent has been successfully developed by Zhang and his colleagues which revealed with the capacity of removing arsenic (V) and arsenic (III) from the water (Zhang et al. 2013). However, to make this powder-type adsorbent more stable and user-friendly, some stable careers have been studied such as diatomite and polystyrene anion exchanger, chitosan, etc. Although these careers showed adsorption properties, the efficacy is unsatisfactory. Therefore, hybrid forms have got interest to many researchers recently. For instance, Qi et al. (2015) reported a novel sorbent made of Fe–Mn binary oxide impregnated chitosan bead. It showed excellent arsenic removal efficiency from the water environment. Recently, biosynthesized nanoparticles are being used in arsenic removal. A nanoscale zero-valent iron was synthesized using plant extract (Aloe vera) which could remove up to 95% of arsenic under low pH (3) condition (Adio et al. 2017). However, the efficiency varied depending on adsorbent dosage, rotation speed, contact time, and pH. A very in-depth review has been done on the current arsenic removal technologies (Nicomel et al. 2016) and reported different techniques such as oxidation, coagulation–flocculation, and membrane along with nanoparticles. The adsorption-based techniques for arsenic removal are presented in Table 5.2.

20 10 1

7.0 5.0 4.0

4.0 7.0 6.0

Modified chicken feathers Allyl alcohol-treated chicken feathers Chitosan resin

Modified from Nicomel et al. (2016)

5 2

5.0 6.5

10 10 2

Adsorbent dosage (g/L) 5 10

Optimum pH 5.0 5.0

Adsorbent Coconut-shell carbon Coconut-shell carbon pretreated with Fe(III) Coal-based carbon Copper-impregnated coconut husk carbon Rice polish Sorghum biomass Fly ash 20 – 20 20 25 40

– – –

25 30

Temperature (C) 25 25

452 – 0.8 *

1125 206

Surface area (m2/g) 1200 –

Table 5.2 Nicomel et al. (2016) reviewed the adsorbents which are capable of As removal from water

0.13 0.115 4.45

0.14 3.6 –

– – –

0.15 – 30

Ranjan et al. (2009) Haque et al. (2007) Diamadopoulos et al. (1993) Khosa et al. (2013) Khosa et al. (2014) Liu et al. (2012)

Lorenzen et al. (1995) Manju et al. (1998)

– 20.35

4.09 –

References Lorenzen et al. (1995) Lorenzen et al. (1995)

Sorption capacity (mg/g) As(III) As (V) – 2.40 – 4.53

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Rahman et al. (2017b) reported the use of carbon nanotubes for the removal of water pollutants from industrial wastewater without functionalization. Al Omar et al. (2016) provided insights into the deep eutectic solvent functionalized carbon nanotubes used to remove arsenic from water. The deep eutectic solvent-methyl triphenyl phosphonium bromide showed a maximum arsenic removal capacity (23.4 mg/g). The results were well fitted with the Langmuir and Freundlich adsorption isotherm models along with the pseudo-second-order kinetics model (Al Omar et al. 2016). The composition of the graphene oxide/copper ferrite foam has been reported for the excellent adsorbent for the removal of As from water with maximum capacity of 51.64 mg/g for arsenic (III) and 124.69 mg/g for arsenic (V) (Wu et al. 2018). Moreover, the graphene oxide/copper ferrite foam exhibited outstanding reusability potential with eight cycles of use without destruction of major properties (Wu et al. 2018). Vieira et al. (2017) proposed a simple inexpensive method for arsenic removal using brown seaweed (Sargassum muticum) as adsorbent after treatment with ironoxy (hydroxides) which removed arsenic as a rate of 4.2 mg/g (III) and 7.3 (V) mg/g. However, this method has the drawback of iron leaching into the water. Wang et al. (2016) synthesized composite of nanoscale zero-valent iron with pinederived biochar to remove arsenic (V) and found almost 100% removal of arsenic from water under anoxic condition. Moreover, this composite has magnetic properties which facilitate the collection of adsorbents from the solution (Wang et al. 2016). A porous indium metal–organic framework was successfully used for the first time to remove arsenic (V) with a maximum capacity of 103.1 mg/g at neutral pH, which is higher than the commercial adsorbents (usually less than 100 mg/g at neutral pH) (Atallah et al. 2017). Pyrolyzed chestnut shell and magnetic gelatin were used to prepare a green low-cost adsorbent to remove arsenic from wastewater (Zhou et al. 2016). The efficacy (45.8 mg/g) of the adsorbent is much higher than the unmodified biochar. Moreover, a similar effort has also been made by Zhang et al. (2016) to remove arsenic (V), and they use magnetic biochars co-precipitation of Fe2+/Fe3+ on water hyacinth biomass. Azadirachta indica (neem leaves) and Mangifera indica (mango leaves) have been found effective in ashed form to remove arsenic from water (Dorris et al. 2018), and this could work in neutral pH condition.

5.3.2

Cadmium Removal Using Unconventional Adsorbents

Cadmium is a toxic metal with numerous adverse health outcomes upon an elevated level of exposure through induction of different toxic factors including reactive oxygen species (Geir et al. 2018; Rahman et al. 2017a). Removal of cadmium from the water phase is very important to get rid of from the adverse effects. In this section, some recent advancement of cadmium removal using unconventional adsorbents will be presented. Recently, a lot of unconventional adsorbents have been tested to removal cadmium from the aqueous phase. For instance, Kataria and Garg

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(2018) synthesized novel magnetic iron oxide nanoparticles loaded sawdust carbon and ethylenediaminetetraacetic acid modified iron oxide nanoparticles loaded sawdust carbon nanocomposites by low-cost biogenic green synthesis approach and their application for cadmium (II) removal. This adsorbent showed higher capacity of camium adsorption compared to iron oxide nanoparticles loaded sawdust carbon adsorbent (Kataria and Garg 2018). Another nano-based adsorbent was prepared by Sharma et al. (2018) using carboxycellulose nanofibers from Australian spinifex grass to remove cadmium from water. The nitro-oxidized carboxycellulose nanofibers reported maximum cadmium removal capacity 2550 mg/g from water which was the highest among the studied literature (Sharma et al. 2018). Furthermore, copper oxide nanoblades were synthesized to reduce cadmium concentration in water (Bhanjana et al. 2017a). This adsorbent also showed promising efficiency to catch up cadmium (192.30 mg/g) from water phase (Bhanjana et al. 2017a). Sol–gel approach was tested to remove metals from water using cashew nut shell resin-coated magnetic nanoparticles (Devi et al. 2017). The Michael addition reaction helps ferric oxide nanoparticle to covalently bond with cashew nut shell resin which further allows removing cadmium from water (Devi et al. 2017). Xu and McKay (2017) have been represented a comprehensive review of literature of adsorption of cadmium using low-cost adsorbents along with their mechanisms. Table 5.3 presented the outcome of previous studies demonstrating cadmium removal using low-cost unconventional adsorbents (Xu and McKay 2017). Study reported a comparative study for the removal of cadmium water phase using raw oak waste and sodium hydroxide-activated oak waste under similar circumstances. Spontaneous and exothermic behaviors were reported for the adsorbent while removing cadmium. Multiwalled carbon nanotubes were synthesized (60–70 nm (width) and length in microns) using chemical vapor deposition method to remove cadmium from water. They found a maximum adsorption capacity of 181.8 mg/g cadmium from aqueous solution (Bhanjana et al. 2017b). In addition, Borah et al. (2018) studied the Burmese grape leaf extract for the removal of cadmium from natural water. They found a maximum cadmium adsorption capacity 44.72 mg/g of Burmese grape leaf extract, and the reaction followed the pseudo-second-order kinetics along with the Langmuir isotherm model (Borah et al. 2018). Chi et al. (2017) reported camium removal from water by using biochar from corm straw. The cadmium removal efficiency was attained maximum of 38.91 mg/g of biochar. Furthermore, highly selective adsorption of cadmium was carried out using nanofibers synthesized through poly (vinyl alcohol)/chitosan using an electrospun technique (Karim et al. 2019). Maximum cadmium removal capacity was reported 148.79 mg/g from the water phase using the nanofiber composites (Karim et al. 2019). Li et al. (2017) investigated biochars from water hyacinth at 300  C to 700  C for cadmium removal from water. Maximum capacity was found 49.837 mg/g of water hyacinth-based biochar as an adsorbent with in the first 5 h. Anastopoulos et al. (2017) reported in a review that described the potential use of unconventional adsorbents from sugar industry wastes to remove cadmium from

100 mg/L

10–400 mg/L

10–500 mg/L

10–500 mg/L – –

Cd(II)

Cd(II)

Cd(II)

Cd(II)

Cd(II)

Cd(II)

Cd(II) Cd(II) Cd(II)

Chitosan

Chitin

Seaweed

Ceramium virgatum (red algae) Alkali-pretreated Ulva onoi (green algae) Ulva onoi (green algae) Sargassum sp. Laminaria japonica (brown algae)

5.5

–

7.8 5 5

7.8

5

5.7–6.02

4–7

8

pH 8

10–300 mg/L

5–100 mg/L

Initial metal concentration 100 mg/L

Metal ion Cd(II)

Adsorbent material Chemically modified chitosan Chitosan

1 1 1

1

10

–

2

1

5

adsorbent dosage (g/L) 5

12 h 6h 180 min

12 h

120 min

–

24 h

24 h

16 h

Contact time 16 h

Table 5.3 List of some UCAs in the removal of Cd as presented by Xu and McKay (2017)

20 22 25

20

20–50

22–24

25

25

25

Temperature ( C) 25

61.90 mg/g 0.76 mg/g 1.85 mmol/g

90.70 mg/g

39.7 mmol/g

0.85 mmol/g

13 mg/g

0.016 mmol/g

85.47 mg/g

Adsorption capacity (qe) 357.14 mg/g

Suzuki et al. (2005) Sheng et al. (2007) Liu et al. (2009)

Suzuki et al. (2005)

References Sankararamakrishnan et al. (2007) Sankararamakrishnan et al. (2007) Rangel-Mendez et al. (2009). Benguella and Benaissa (2002) Ahmady-Asbchin et al. (2009) Sari and Tuzen (2008)

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water and wastewater. Furthermore, Malik et al. (2016) reviewed literature using very common and very recent articles and confirmed that cellulosic low-cost adsorbents has potential to remove cadmium from water. Biochars produced from different sources have a great potential to remove cadmium. Karunanayake et al. (2017) studied the magnetic biochar produced by magnetite precipitation onto Douglas fir biochar that had been generated by wet fast pyrolysis which could remove cadmium from aqueous solution efficiently (Karunanayake et al. 2017). Naeem et al. (2019) reported two types of biochar (a) wheat straw biochar and (b) acid-treated wheat straw biochar for cadmium remediation. Wheat straw biochar and acid-treated wheat straw biochar demonstrated the maximum adsorption of cadmium from contaminated water by 31.65 mg/g and 74.63 mg/g, respectively. The unconventional adsorbents have the potential to remove cadmium from the contaminated water as well as to clean up contaminated potable water. A composite bead was prepared using carboxymethyl-β-cyclodextrin-functionalized chitosan impregnated with water-insoluble epichlorohydrin cross-linked β-cyclodextrin polymer to remove cadmium from water phase by Sikder et al. (2017). They found excellent cadmium removal capacity of their composite adsorbent which achieved adsorption capacity (Max.) of more than 378 mg/g from aqueous solution (Sikder et al. 2017).

5.3.3

Chromium Removal Using Unconventional Adsorbents

Chromium, especially chromium (VI) ion, has many adverse health effects such as causing liver damage, carcinogenic effects, and inherited gene defects in human. It has a diversified application in the industries and ultimately drained into water. Thus, it is a matter of concern to remove chromium (VI) ions from water due to its toxicity and mobility. In this section, chromium removal unconventional adsorption techniques will be presented. Recently, a lot of efforts have been made to develop low-cost adsorbents to remove chromium from water phase including nanoparticles, biochars, metal–organic frameworks, and others. Kuppusamy et al. (2016) studied the potentiality of dried twigs of Melaleuca diosmifolia for the removal of chromium from aqueous system. It was found that maximum 62.5 mg/g capacity has been obtained using the adsorbent to remove chromium. Guar gum-nano zinc oxide biocomposite has been used successfully to detoxify chromium from water (Khan et al. 2013). About 55.56 mg/g capacity was achieved by Guar gum-nano zinc oxide with very good regeneration capacity of the adsorbent (Khan et al. 2013). Rice straw-based adsorbents were reported by Elmolla et al. (2015); however, the activated form of rice straw-based adsorbent showed significantly better performance in chromium removal under similar operating conditions (Elmolla et al. 2015). Likely, another agricultural waste was tested for adsorption of chromium by Ali et al. (2016). Acrylonitrile-grafted banana peels showed excellent adsorption of chromium from the water; about 96% was removed following the pseudo-second order kinetic model (Ali et al. 2016). Moreover,

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Dehghani et al. (2016) reported treated waste newspaper as potential adsorbent to remove from the water phase. Maximum adsorption capacity 59.88 mg/g was reported at low pH condition and can be explained as low-cost adsorbent (Dehghani et al. 2016). Recently, groundwater treatment residuals have been tested for the reuse potential as an adsorbent by Kan et al. (2017). They used silica sand coated with groundwater treatment residuals for the removal of Cr from water. Though the adsorption capacity is not high (0.27 mg/g), still it has the potential to reduce environmental burdens. Besides, chemically modified palm kernel shell was found very effective in chromium removal (19 mg/g) from water (Hanafiah et al. 2018). Carbon nanotubes have the potential to remove metals from the water phase. Bhaumik et al. (2016) reported polypyrrole wrapped oxidized multiwalled carbon nanotubes nanocomposites for the removal of chromium and achieved excellent adsorption capacity (294 mg/g) (Bhaumik et al. 2016). In addition, magnetic biochars have been studied for environmental remediation. Shang et al. (2016) prepared magnetic biochars with Astragalus membranaceus residue to remove chromium from water. They reported maximum chromium removal capacity (23.85 mg/g) at pH 2 (Shang et al. 2016). Whereas, Samani and Toghraie (2019) reported polyaniline/sawdust/poly ethylene glycol composite which could remove chromium at maximum capacity of 3.2 mg/g at pH 2. Swami and Gupta (2016) reported a comprehensive overview on the adsorption properties of saw dust, bark, and rice Husk. Altun 2019 prepared composite biosorbent from chitosan and sour cherry kernel shells which removes chromium (24.492 mg/g) from the water phase. Meanwhile, Hibiscus cannabinus kenaf was studied by Mohan et al. (2019) to remove chromium from aqueous solution and found 0.538 mg/g maximum adsorption capacity. Another low-cost adsorbent was prepared from spent coffee grounds to remediate chromium from water (Ma et al. 2019). They found about 91.0% of removal of chromium at optimum operating conditions, and the adsorption capacity is found to be as high as 22.75 mg/g. Recently, Hossain et al. (2017) reported successful removal of at pH 3 chromium from tannery effluent using chitosan synthesized from the crab shell. The unconventional adsorbents have been showing great opportunity to remove chromium from water phase.

5.3.4

Lead Removal Using Unconventional Adsorbents

Lead is a well-known environmental pollutant with numerous adverse human health effects. Thus, tremendous efforts have been given on the remediation technology of lead from the water environment. Different nanomaterials have been tested to remove metal pollutants from water. Graphene oxide-based microbots were reported to remove lead from water system (Vilela et al. 2016). Metal–organic frameworks have also been reported for the remediation of environmental contaminants. A comprehensive overview is presented by Kobielska et al. (2018) on the use of

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metal–organic frameworks for lead removal. Mohan et al. (2014) reported energy cane biochar as a potential adsorbent to remove lead from aqueous solutions that could replace the use of the commercial activated carbon. Scots pine (Pinus sylvestris L.) biochar and silver birch (Betula pendula) biochar has been investigated by Komkiene and Baltrenaite (2015) to remove metals including lead from the water phase. Polymer functionalized nanocomposites have been reviewed by Lofrano et al. (2016). Wang et al. (2015a, b) reported a novel approach to prepare biochar from potassium permanganate treated hickory wood through slow pyrolysis (600  C) which could remove lead (153.1 mg/g) efficiently from the water environment. A comprehensive review has been conducted by Li et al. (2017) to examine the potential role of biochars in removing lead from the water environment. A nanoadsorbent prepared by incorporating organic ligand onto the mesoporous silica to remove lead from the water phase was reported to find a maximum adsorption capacity of 169.34 mg/g. In addition, Awual (2016) developed adsorbent by indirect ligand immobilization onto the mesoporous silica for lead (II) ions detection and adsorption from aqueous media. They found the maximum removal capacity of 188.67 mg/g for lead removal from aqueous solution (Awual 2016). Zhou et al. (2017) prepared fresh and dehydrated banana peels. These peels were used as sorbent biochars through a facile one-step hydrothermal carbonization approach. Both fresh and dehydrated banana peels showed excellent lead removal capacity of 359 mg/g and 193 mg/g, respectively (Zhou et al. 2017). Deng et al. (2017) prepared chitosan–pyromellitic dianhydride modified biochar to clean up lead from water. They studied a low-cost approach based on wheat straw pulp fine cellulosic as a biosorbent for the removal of lead in aqueous solutions after nanofibrillation and sulfonation pretreatments.

5.4

Mechanisms of Adsorption

Adsorption usually refers to the surface phenomena to take out target molecules from aqueous or gaseous phase into a solid media (Fig. 5.2). In this process, a remarkable surface–interplay takes place between adsorbent (the solid media/surface where the target molecule/s arrested) and the adsorbate (the target molecule/s which secure position on to the adsorbent surface). However, the process is not so simple due to the involvement of numerous other factors such as affinity between adsorbent and adsorbate, pH, surface area, porosity, contact time, etc. Therefore, the main challenge lays to choose the potential adsorbent focusing on cost-effectiveness, adsorption capacity (Qmax), adsorption rate, selectivity, desorption easiness, and reaction kinetics along with other challenges of the possible mechanism of actions (Crini 2005). Although a huge number of articles have been reported the adsorption processes and their applications, still the adsorption mechanism is not well understood,

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Fig. 5.2 Generalized scheme for adsorption process to clean up polluted water. Porous surface of the adsorbent arrest adsorbate in the aqueous solution then deposited at the bottom of the vessel and finally screened out through filtration

Fig. 5.3 List of pollutant adsorption mechanisms modified after Crini (2005, 2006) and Crini and Badot (2007). Major four mechanisms are such as physisorption, chemisorption, ion exchange, and precipitation involved in the adsorption processes

beceause there is a possibility of a large number of interactions. Figure 5.3 is showing possible interactions between adsorbate and adsorbent (Crini 2005): such as physisorption (physical adsorption), chelation, etc. (Crini 2005). These interactions can be independent and also two or more interactions can be taking place simultaneously (Crini et al. 2018).

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Despite many debates, four main adsorption mechanisms have been proposed (Crini et al. 2018). They are: 1. 2. 3. 4.

Physisorption Phemisorption Ion-exchange Precipitation

A comprehensive discussion on the abovementioned terms has been presented in a review article by Robalds et al. (2016). Several adsorption mechanisms have been discussed by different authors: Srivastava and Goyal (2010), Naja and Volesky (2011), Asgher (2012), Michalak et al. (2013), and Robalds et al. (2016). In case of biosorption, two mechanisms are prominent and well described: chemisorption (ion exchange, complexation and chelation) and physical adsorption (for live species it’s called bioaccumulation). However, for the purpose of the decolorization, the use of unconventional materials generally attributed to the mechanisms of adsorption and ion exchange (Allen and Koumanova 2005; Crini 2006). Shukla et al. (2002) reported ion-exchange and hydrogen bond interactions are the major mechanisms for metal removal using sawdust as adsorbent. The cell wall of sawdust consists of cellulose and lignin, along with hydroxyl groups which are the active agents for the ion exchange. Biopolymers such as starch, cellulose, chitin, and alginates and their derivatives, e.g., chitosan and cyclodextrin, have also been studied for the adsorption process. Sikder et al. (2018) reviewed on native cyclodextrin and modified cyclodextrin to remove pollutants from water phase following adsorption process. For cyclodextrinbased adsorbents, there are two major mechanisms reported: (a) inclusion complex formation and (b) diffusion into the polymer network (Crini 2014; Sikder et al. 2018). The biopolymers are functionally more suitable due to their high selectivity for the adsorption process specifically the chelation and ion-exchange processes for metal removal (Crini and Badot 2008). Moreover, Crini (2015) reported in a review that the dye removal by biopolymers is mainly based on ion exchange, acid–base interactions, precipitation, hydrogen bonding, hydrophobic interactions, and physisorption, although its mechanism is not fully understood (Crini 2015). However, Morin-Crini et al. (2018) recently reported in a review that, despite laboratory-scale success, the newly developed techniques have not been realized in industrial scale along with the shallow mechanistic explanations. Therefore, the mechanistic approach of the study on adsorption processes needs special attention to unveil the underlying processes to invent new pollution remediation technologies based on adsorption.

5.5

Concluding Remarks

Removal of pollutants from the water environment has now become a global concern particularly metal pollution. Tremendous efforts have been made to invent low-cost highly selective adsorbents for the remediation of toxic metals from the water. However, most of these efforts are limited to laboratory-scale experiments rather

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than technology development. The complexity in the adsorption processes and the interfering factors are the pivotal challenge. However, besides the conventional adsorbents (commercial activated carbon, organic resins, activated alumina, zeolites, and sand) in recent years, a great deal of work has been done on unconventional adsorbents including nanomaterials, metal–organic frameworks, biochars, etc. to remove metals from the aqueous phase. In this chapter, different unconventional adsorbents were considered for a discussion focusing on their sources and adsorption capacity in the removal of metals such as arsenic, cadmium, lead, and chromium. The laboratory-scale study reveals that the potential of unconventional adsorbents is really appreciable in terms of their performance for the removal of metals. However, they are not being used in industrial-scale probably due to lack of confidence in material engineering and feasibility. Thus, it is very important to rethink about the multidimensional and multidisciplinary approaches of research in adsorbent development for the industrial and commercial purposes to reduce metal loads in the aquatic environment. Moreover, another issue should be kept in mind that the industrial-scale effluent is not a solvent; rather it’s a huge mixture of heterogeneous chemicals which might have a greater influence on the adsorption processes and selectivity. Finally, great scopes are now existing to implement unconventional solid materials such as biomass, engineered nanomaterials, and chitosan to reduce metals from water. The mechanisms of adsorption should be studied more carefully and should also focus on the simultaneous removal of multiple metals using single adsorbent.

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Chapter 6

Desalination Technology for Water Security Mashura Shammi , Md. Mostafizur Rahman Mohammed Mofizur Rahman

, and

Contents 6.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.2 Water Security and Desalination Technologies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.2.1 Thermal Desalination . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.2.2 Membrane Desalination . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.2.3 Membrane Fouling and Scaling . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.2.4 Prevention of Membrane Foulants: Forward Osmosis and Other Technologies 6.2.5 Inorganic Membranes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.2.6 Ultrasound Application for Membrane Cleaning . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.2.7 Protective Layering on Membrane . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.2.8 Membrane Recycling . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.3 Water-Energy Nexus and Desalination Technologies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.3.1 Global Desalination Market in the Nexus of Water–Energy and Environmental Impacts . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.3.2 Reducing the Carbon Footprint of Desalination . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.3.3 Integrated Hybrid Power Generation for Desalination . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.3.4 Solar Desalination Technologies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.3.5 Nuclear Energy for Desalination . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.3.6 Nano-Enabled Desalination . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.3.7 Desalination for Water Security in Water-Energy Nexus for Developing Country 6.4 Concluding Remarks . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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M. Shammi (*) · Md. M. Rahman Department of Environmental Sciences, Jahangirnagar University, Dhaka, Bangladesh e-mail: [email protected]; [email protected] M. M. Rahman Alexander Von Humboldt International Climate Protection Fellow, Institute for Technology and Resources Management in the Tropics and Subtropics (ITT), TH Cologne – University of Applied Sciences, Cologne, Germany © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature Switzerland AG 2021 Inamuddin et al. (eds.), Water Pollution and Remediation: Heavy Metals, Environmental Chemistry for a Sustainable World 53, https://doi.org/10.1007/978-3-030-52421-0_6

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Abstract Desalination is the technology to provide energy, fresh water, and food security concurrently for the remote, coastal, and energy-lacking countries. In this chapter, we have discussed various techniques and problems of thermal and membrane desalination techniques such as electrodialysis, reverse osmosis, ultrafiltration, and nanofiltration, membrane distillation, and various integrated methods. However, these methods are expensive and energy-intensive, with massive carbon-footprints, and have a serious problem of membrane fouling. Moreover, these technologies are often not sustainable for implementation in many energy-and-water-starved developing countries. We have, therefore, concentrating on research and progress of nanomaterials and energy-efficient membrane development to overcome membrane fouling and scaling prevention. We have also focused on the modification of membrane process such as forward osmosis. We have also discussed the integration of renewable energies such as solar desalination and hybrid power generation such as nuclear desalination to reduce carbon footprint and enhance cost-effectiveness to obtain fresh water. From the viewpoint of water–energy nexus, choosing the right types of desalination techniques and processes should be strategically planned, designed, and implemented to achieve water security. Keywords Carbon footprint · Desalination · Energy security · Forward osmosis · Membrane fouling · Nanofiltration · Photovoltaics-reverse osmosis · Seawater reverse osmosis · Water security · Water–energy nexus

6.1

Introduction

Water plays a major role in all our daily operations and is increasing general consumption each day as human living standards are rising. Freshwater availability is becoming a growing global problem (Gorjian and Ghobadian 2015). In most nations around the globe, groundwater depletion is a prevalent issue. This is due to overcrowding as well as problems with climate change. As a result, the withdrawal rates for groundwater are up to 1–3% annually. Many areas worldwide access deeper groundwater for water demand management (Gude 2018). This enhanced abstraction and pollution of groundwater led numerous problems to ecosystems that depend on it. Contrary to safety for any resource, its main variables are availability, accessibility, safety, affordability, and constancy determined by water security (Gain et al. 2015). Better water resource management is, therefore, undoubtedly allied to improved water security. For a sustainable human community, water and power are the vital components. They are highly interdependent because the production of electricity demands vital amounts of water. Subsequently, vast energy sources are required for the manufacturing, processing, dispersal, and end use of water (Kim et al. 2018). The burdens on easily accessible fresh surface water are intensified during the times of

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drought, which urgently require the consideration of substitute sources for the cities, metropolises, and other entities. Fresh water can thus be transferred over longer distances or obtained through water treatment of substandard quality (Aminfard et al. 2019). It is, therefore, imperative to reconsider sustainable water management that takes financial, societal, and environmental elements into account. Desalination of seawater has continued to be one of the oldest water treatment processes in the human antiquity. Seawater and brackish water desalination are well recognized as an alternative solution for delivering fresh water in many waterstressed regions across the globe (Zhang et al. 2018). It is the procedure of salt exclusion and is a major know-how in easing the water crisis (Babu et al. 2018). It has confirmed to be one of the most dependable adaptive technology for addressing freshwater needs not only for household consumption but for industrial purposes as well (Ullah and Rasul 2018). Although numerous indicators for improving desalination are recognized, there is still a significant margin for improvement. The present methods of desalination (particularly thermal systems) involve significant energy to separate salt from seawater (Ali et al. 2018). The capacity to secure water in the arid and coastal regions through desalination technology covers a variety of working circumstances that have many benefits, which includes the use of the low-quality water resources of seawater and brackish water (Abdelkareem et al. 2018). Nevertheless, the elevated price of standard desalination plants combined with the emissions of greenhouse gas from associated electricity generation processes enforces the identification of inexpensive and ecologically approachable alternative energy sources (Abdelkareem et al. 2018). Further reliable and inexpensive entree to safer water will necessitate scientific and technical novelty (Alvarez et al. 2018). Nanotechnology is also likely to play a key part in offering a variety of exceptional possibilities for improving certain water processing systems, including adaptable treatment systems and equipment adaptable to goals (Alvarez et al. 2018). Water is simultaneously part of the provision for human sustenance, salvation for environments, as well as security for the industrial and economic sector. Energy and water security coexist in similar realms, and their reliance on each other depicts an essential worldwide assembly of security. For instance, water, power, and food supply systems in the Gulf regions are becoming extremely interlinked, and the rise of coproduction of crops and the use of cross-sectoral resource footprints over the past centuries increased considerably (Al-Saidi and Saliba 2019). Yet again, we must remember that the present wastewater managing approaches are not sustainable. As we are aware, the current saline wastewaters usually are disposed of into the ecosystems, dispersed in aquifers, or inserted into the grounds causing contamination of the environmental resources. However, in many coastal as well as arid dry regions, onshore and offshore oil and gas exploration produces highly saline wastewaters. Due to firmer environmental control strategies by the governmental regulatory agencies, saline brines/effluents reuse is necessary (Osipi et al. 2018). Industries require intensive styles for wastewater management that needs lowering of freshwater extractions and abolition of wastewater discharge through zero liquid discharge method (Davenport et al. 2018).

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Providing drinking water in a safe, consistent, and inexpensive way is one of the utmost encounters of the twenty-first century (Alvarez et al. 2018). Energy frameworks support technical resolutions to fulfil the United Nations’ clean water and sanitation goal which is the Sustainable Development Goal six (SDG 6), through consequences intended for forthcoming energy necessities and greenhouse gas discharges. In addition, the energy industry also plays a leading role in water use and has established water efficiency goals in SDG 6 because of important constraints on ongoing energy planning (Parkinson et al. 2019). Consequently, managing water security has numerous facades for many parts of the remote, coastal, and energylacking countries. The best way to address this is by using solar power to desalinate the water supplies but counterbalance many damaging environmental effects of desalination. Desalination will be the technology to provide energy and water security concurrently. This chapter, therefore, discusses the drivers of choosing desalination technologies for water treatment in the light of water-energy nexus which has a likely sustainable solution to the world’s plummeting needs of freshwater scarcity issue.

6.2

Water Security and Desalination Technologies

Enhancing water security requires extraordinary adaptability and flexibility to cope with many avant-garde water systems as well as flowrate. For illustration, storm water, brackish or saline surface and groundwater, industrial grey water were put into miscellaneous treatment methods for diverse uses (Alvarez et al. 2018). It is important to remember that using alternative water supplies are crucial for the accomplishment of SDG 6 of clean water and sanitation (Jones et al. 2019). In terms of water pollution management, regulatory demands and economic drivers for reliable and low-cost access to clean water, better wastewater disposal methods, and recovery are increasingly needed to improve water safety as well as security. Seawater is a vast resource that can provide water security. Unfortunately, seawater usually contains a higher salinity as total dissolved salts ranged from 35,000 to 45,000 ppm (El-Ghonemy 2018). Wastewater also contains higher total dissolved salts. Desalination is the process of eliminating the total dissolved salts from different saline water gradients (Abdelkareem et al. 2018) as well as wastewaters. The five major components of a desalination process are shown in Fig. 6.1 that includes intake, pretreatment, process, posttreatment, and concentrate management (El-Ghonemy 2018; Wilder et al. 2016). The methods of desalination are divided into two primary kinds (Fig. 6.2): (1) thermal and (2) membrane processes. Thermal procedures for desalination are founded on stage variations including (a) multi-phase flash distillation, (b) multieffect distillation, (c) vapor compression distillation, and (d) humidification dehumidification. Membrane desalination uses membranes with unique properties. The major membrane methods are (a) electrodialysis and electrodialysis reversal, (b) reverse osmosis, (d) membrane distillation, and others (Abdelkareem et al.

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Fig. 6.1 (a) The five main stages of a desalination system which include (1) seawater intake; (2) pretreatment by conventional coagulation/ filtration or membrane-based filtration system; (3) the main process of removing dissolved solids, e.g., reverse osmosis; (4) posttreatment which includes the process such as pH adjustment, disinfection, etc.; and finally (5) concentrate management which includes brine discharge, brine dilution, offshore ocean outfall, etc

Fig. 6.2 Schematic diagram of the desalination technologies available in the present world. Based on the energy and membrane usage, the processes are divided into thermal processes and membrane processes, respectively

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2018; Gorjian and Ghobadian 2015). There are also alternative processes in the list together with capacitive deionization, ion exchange resins, and freezing which are beyond the scope of discussion in this chapter.

6.2.1

Thermal Desalination

Thermal desalination is an intensive energy consumption process that uses conventional fossil fuel sources (Al-Othman et al. 2019). Thermal desalination techniques have historically been used for freshwater generation using phase change (for instance, evaporation and condensation) (Qasim et al. 2019). The multistage-flash is an extensive marine desalination plant that is used widely and operates in cogeneration with a power station in the Gulf area (El-Ghonemy 2018). In the 1990s, many multistage-flash plants were established, and most of them are currently in operation in the Middle East Gulf countries (Choi 2016). The functioning temperature in the multistage-flash and multi-effect distillations are 65
C and 112
C, correspondingly. Conventionally, multistage-flash and multi-effect distillation have traditionally demonstrated very high effectiveness in a hostile setting without requirements for seawater pretreatment (Thabit et al. 2019). The multieffect distillation desalination plant is more useful than the multistage-flash form from the standpoint of thermal efficiency, while the multistage-flash type is more appropriate for water treatment for large dimensions. Moreover, locations like energy prices and the interest rate are subject to constraints, which play a key role in full desalination costs and in improving and optimizing the technique. Higher salinity gradients usually account for thermal desalination containing mechanical vapor compression more appropriate (Osipi et al. 2018). Taltape multi-effect distillation in northern Chile can be cited as an example of non-grid desalination treating brackish water with distinct metalloids like mercury and boron at elevated concentrations (Tarpani et al. 2019). The authors identified the effects of the multi-effect distillation on the environment and highly recommended solar desalination as a substitute to counterpart the usage of diesel and biomass energy. The authors (Tarpani et al. 2019) further reported the reduction of arsenic and boron concentration from the treated water and obtained the standards which were suitable for irrigation and livestock consumption.

6.2.2

Membrane Desalination

The membrane-based process can be categorized into two types: (1) pressure-driven processes, and (2) osmotically driven processes. Numerous membrane operation technologies including multi-effect distillation, reverse osmosis, and membrane distillation have been developed because of current growths and advancements in membrane technology. Microfiltration, ultrafiltration, nanofiltration, and reverse

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Table 6.1 The spectrum of filtration techniques for water and wastewater purification and desalination

Ultrafiltration

Pore size (nm) Greater than 100 20–100

Nanofiltration

0.8–3

Small molecules

Reverse osmosis

0.3–1

Ion-exchange resins

0.3–1

Salts, ions, color, low molecular weight species Purifies and changes

Filtration types Microfiltration and beyond

Filter spectrum High suspended solids

Macromolecules

Speciality Sand, silt, organic particles, algae, and fibers even in the presence of fats, oils, and greases Microbiological bacteria, viruses, colloids, silt, and more Virtually all bacteria, viruses, cysts, humic materials, alkalinity, and hardness All bacteria, viruses, cysts, pesticides, radium, and all inorganic contaminants Removes metal ions and mineral contents to soften the water and improve its purity

Modified after Stevens et al. (2017)

osmosis—all are within the pressure-driven process category (Table 6.1), while forward osmosis and pressure-retarded osmosis fall within the category of osmotically driven process. Other membrane-based process included membrane bioreactor and membrane distillation.

Electrodialysis In the electrodialysis process, dissolved ions in liquids are eliminated employing the direct current and the ion-exchange membranes. Cations (positive ions) or anions (negative ions) can pass selectively through the membranes. The electrodialysis reversal method is primarily used according to the same principle as the electrodialysis method. The only exception is that in electrodialysis reversal, both the cathodes and anodes are inverted. In the electrodialysis reversal method, ion-exchange membranes have a longer life and enhanced productivity (Choi 2016).

Reverse Osmosis Reverse osmosis is a principal membrane-based desalination procedure for saline water and wastewater risen significantly in reaction to water shortage and stress (Coutinho de Paula and Santos Amaral 2018; Mito et al. 2019). Reverse osmosis is typically divided into two kinds, namely, (a) brackish water reverse-osmosis and (b) seawater reverse osmosis. The operational procedure comprises units of pretreatment and posttreatment. Pretreated water passes through a semipermeable

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membrane which allows the passage of water to prevent salt particles through a highpressure pump (Abdelkareem et al. 2018). The use of reverse osmosis desalination has an excellent choice for producing clean water for households and industries (Coutinho de Paula and Santos Amaral 2018) and has developed into leading know-how for water treatment (Werber et al. 2018). Although efficacy is steadily being improved, reverse osmosis desalination continues to be an energizing method (Mito et al. 2019), most optimized, and leading membrane-based desalination process in the market (Qasim et al. 2019). High fossil fuel and power consumption; brine waste, with its harmful pollution and negative effects on the aquatic ecosystem; and the elevated capital cost at the beginning of the operation can be portrayed as the main shortcomings (Abdelkareem et al. 2018).

Ultrafiltration and Nanofiltration Ultrafiltration is a pressurized membrane process which distinguishes bacteria and viruses from the water on a size-basis. Therefore, water filtration, disinfection, or many serious waterborne diseases or epidemics can be avoided when ultramembrane-filtered water is being consumed (Arnal et al. 2004). Multistage flash plants can be mutually joined together with other techniques of desalination such as nanofiltration and reverse osmosis (El-Ghonemy 2018). Both nanofiltration and reverse osmosis use cross-linked polyamide membranes for the required water and wastewater separation of salts (Stevens et al. 2017) with ultrahigh permeance and high salt rejection for effective desalination (Wang et al. 2018a, b).

Membrane Distillation Membrane distillation is a dynamic thermal method for the treatment of high salinity waters that is effective for treating industrial wastewater from unusual locations such as hydrocarbons and salt from desalination plants (Deshmukh et al. 2018). Membrane distillation has shown that both freshwater and electricity can be produced independently (Ali et al. 2018). These can produce clean and sustainable energy from different waste streams, including salt- and spoiled water, otherwise accountable for the environment. Membrane distillation has a low energy efficiency compared to reverse osmosis which is advantageous. Membrane distillation can use low-grade heat sources which are excess thermal from power plants and industrial plants (Deshmukh et al. 2018). It further has the potential to desalinate higher salinity gradient waters and wastewaters using lower-grade heat through waste heat recovery at the water–energy nexus. It further has the potential to use in integrated a zero-liquid discharge system which may include a membrane distillation unit along with the reverse osmosis unit to reuse feedwater.

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Integrated Methods Several blending of desalination methods has been implemented, linking forward osmosis, reverse osmosis, assisted reverse osmosis, microfiltration, mechanical vapor compression, and membrane distillations (Osipi et al. 2018). Furthermore, an amalgamation of membrane distillations with pressure-retarded osmosis or electrodialysis reversal can significantly improve the efficiency and deliver a smooth and sustainable path for freshwater and energy production. The notion of high-pressure reverse osmosis which is functioned at a hydraulic pressure greater than 100 bar has been also applied to competently desalinate hypersaline brines (Davenport et al. 2018). High-pressure reverse osmosis has a potential application for treating highsalinity industrial wastewater. This method can decrease the energy needs for desalination. This method also has a likely industrial application of wastewater management by zero liquid discharge. An economical replacement for greater salts concentration could be microfiltration blended with assisted reverse osmosis as well as reverse osmosis. Microfiltration can be further coupled with mechanically based vapor compression or forward osmosis-based vapor compression (Osipi et al. 2018).

6.2.3

Membrane Fouling and Scaling

In all pressure-driven membrane-based processes and osmotically driven membranebased processes, fouling is inevitable (Qasim et al. 2018), including membrane bioreactors and membrane distillations. The biochemical contact between the membrane surface and foulants results in membrane fouling (Choudhury et al. 2018). Membranes accrue both organic and inorganic materials on the membrane surfaces which causes membrane and water quality to foul definitely (Son et al. 2018). Depending on the foulants deposited and analyzed in various studies, membrane fouling can be categorized as (a) fouling due to deposition of colloidal substances, (b) fouling due to the adhesion of organic substances, (c) fouling due to inorganic deposition of sparingly soluble minerals salts, and (d) fouling due to the formation of bacterial biofilm (Qasim et al. 2018). Again, depending on the surface location of membrane fouling, it can be of two types: (a) external surface fouling types and (b) internal surface fouling types. Qasim et al. (2019) and the references therein further referred to the complicated physicochemical interactions among the foulants in the feedwater which are unwanted impurities and membrane surfaces. When external surface fouling occurs, the foulants typically accumulate on the membrane surfaces. On the contrary, internal fouling includes the fouling of the pore spaces of the membranes. Membrane fouling is a major drawback of the reverse osmosis system which must be tackled (Volpin et al. 2018). Fouling of membrane decreases the permeate flux and water yield as well as considerably lessens the membrane’s lifespan while increasing the demand for energy and feeding stress. It further increases the maintenance and replacement costs of the membrane (Qasim et al. 2018). Many reverse

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osmosis systems need to restrict water recovery to prevent fouling and scaling of the membranes. In addition, the salts that are eliminated from the seawater, brackish water, or groundwater usually precipitate on the membrane surface, consequently declining the general competence of the procedure (Aziz and Kasongo 2019). The use of anodic zinc and commercially available antiscalant resulted in higher flux, 30.80 Lm2 h1 and 32.78 Lm2 h1, respectively, of the membrane compared to the untreated membrane flux (25.56 Lm2 h1) that declined continuously (Aziz and Kasongo 2019). Consequently, this characterizes risky operational and economic difficulties in reverse osmosis desalination. Nanofiltration membranes usually eliminate most of the organic compounds and are a decent choice for the pretreatment for organic compounds from wastewater (Jamil et al. 2018). Nevertheless, maximum inorganic compounds pass through the membrane. Special attention should be provided to ultrafiltration membranes as pretreatment for desalination plants. It provides high-quality feedwater for the secure and stable functioning of the desalination units. After treatment and pre-treatment phase is critical for the longevity of ultrafiltration/nanofiltration/reverse osmosis membranes (Al Aani et al. 2018). Moreover, time variable batch or semi-batch procedures such as closed-circuit reverse osmosis and pulse flow reverse osmosis significantly improve fouling resistance in the latest versions of desalination. The shorter residence times of each batch reverse osmosis are usually connected with the series of salinity gradients ranging from low to elevated salinity. These simple improvements of desalination techniques allow for a considerable amount of salt recovery proportions compared to the ongoing reverse osmosis systems for entirely inspected water types including seawaters and brackish-type groundwater. Warsinger et al. (2018) proposed that batch reverse osmosis can operate in high salinity and elevated regeneration during scaling, improved energy efficiency, and superior strength compared with traditional reverse osmosis in inorganic fouling. In addition, a batch reverse osmosis scheme can be upgraded to the existing reverse osmosis schemes and supplied with the reverse osmosis concentrates to recover extra water.

6.2.4

Prevention of Membrane Foulants: Forward Osmosis and Other Technologies

Pretreatment of feedwater is significant for decreasing fouling mechanism. Forward osmosis is a lower level fouling innovation that accomplishes a steady, superiorpermeate for wastewater treatment and reclaim for a long-standing duration (Corzo et al. 2018). Forward osmosis membranes usually comprise pore dimensions lesser than NF membranes. Accordingly, forward osmosis performs excellently in removing inorganic compounds (Jamil et al. 2018). The principles of forward osmosis are growing know-how grounded on typical osmosis of the environment. Solvents such as water permeate over a semipermeable membrane from a low gradient electrolyte solution (the feedwater) to a higher gradient electrolyte (concentrated) “draw”

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solution (Awad et al. 2019). Recovery of the permeated water then requires an additional phase for the segregation of the permeated water from the draw solution that enables the draw solution to be reused. Hybrid forward osmosis–reverse osmosis technology was lately reported by Volpin et al. (2018) to augment the seawater reverse osmosis desalination energy efficacy. The forward osmosis as a “pretreatment” phase was intended to use for thinner seawater concentration along with the recovered wastewater earlier at the desalination stage. The processes ultimately reduced the power needs for the seawater reverse osmosis desalination (Volpin et al. 2018). In seawater reverse osmosis desalination facility, two types of uncertainties are very prevalent, i.e., (i) water consumption fluctuation and (ii) reverse osmosis units malfunction (Wang et al. 2018a, b). Therefore, using hybrid forward osmosis technologies, notably osmotically driven membrane bioreactors, osmotically driven electrodialysis, osmotically driven nanofiltration, or osmotically driven reverse osmosis, can be capable knowhow for greywater reuse in agriculture. These hybrid membrane processes can encounter elevated effluent quality demands, predominantly for boron and/or salt content in wastewater (Corzo et al. 2018). Pressure-retarded osmosis is another innovative technology on the development to yield renewable salinity gradient energy. Yet, intense membrane fouling limits the pressure-retarded osmosis performance expressively when tested with actual wastewater as the feed (Cheng et al. 2018). Cheng et al. demonstrated that fouling inclination can be reduced significantly when forward osmosis utilized as a pretreatment to abstract water from the wastewater rivulet to the inter-circle solution, while thinner inter-circle solution served as a fresh feedstuff to the pressure-retarded osmosis unit. Consequently, the innovative forward osmosis–pressure-retarded osmosis hybrid scheme was skilled of retaining osmotically driven power with the advantages of (1) lower fouling generation, (2) painless scrubbing of deposited salts, and (3) negligible exterior energy requirement that resulted in low carbon emission. Therefore, integration of seawater reverse osmosis with pretreatment by forward osmosis and posttreatment by pressure-retarded osmosis as well as using wastewater reverse osmosis brine will likely to accomplish noteworthy net energy savings. Nevertheless, the fouling of the forward osmosis as well as the pressure-retarded osmosis requires operative management plans and requires extensive optimization of the process (Fane 2018). A pretreatment of feedwater with granular activated carbon adsorption can significantly remove organic compounds (~90%) responsible for organic fouling from reverse osmosis concentrate in wastewater reclamation plants (Jamil et al. 2018), while acid pretreatment at pH 5 can significantly reduce inorganic scaling which is responsible for the decrease of membrane lifespan. It has been clear that forward osmosis–reverse osmosis is the highly expensive method of desalination and often impractically liable to the high level of brine gradient and costs (Osipi et al. 2018).

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Inorganic Membranes

The inorganic membrane has been undergoing quick advancement lately over organic polymer membranes in desalination. Inorganic membranes have several advantages in the list, namely, of resisting high temperature, higher chemical stability, enduring erosion resistance in chemical washing, longer lifespan, and autoclavable in high temperatures and pressures. Consequently, all these exceptional characteristics indicate inorganic membranes to be a good candidate for applications in desalination and water treatment with high fluidity, complete salt refutation, and consequently lower fouling (Elimelech and Phillip 2011; Fard et al. 2018). However, the rate of inorganic membranes manufacturing is very high. Ceramic membranes are comprehensively the most promising at present compared with silica-based membranes or zeolite-based membranes among the different inorganic membranes. Amorphous silica membranes have a higher affinity to adsorb water molecules and structurally degrade in contact with water. Consequently, this limits their application for water treatment. In addition, fabricating a flawless zeolite membrane with proper depth is very difficult. Subsequently, the development of mixed matrix membranes proposes a mutually exclusive solution for fluidity, efficient functioning, and antimicrobial characteristics.

6.2.6

Ultrasound Application for Membrane Cleaning

Application of ultrasound seems to be an operative means of flux enrichment as well as washing and cleaning of membranes to increase the lifespan. Ultrasound embraces an exclusive facility to produce distinct chemical and physical paraphernalia. This consequence can effectively confiscate foulants from the membrane surface (Qasim et al. 2018). Yet, applications of ultrasound are unable to yield a noteworthy impact on the blockage of pores inside the membranes. Furthermore, ultrasound can only remove external fouling of the membranes.

6.2.7

Protective Layering on Membrane

An expiatory shielding of polyelectrolytes layering on the surface of the membrane can significantly improve the lifecycle of membranes and decrease fouling tendency (Son et al. 2018). A cation polymer such as poly diallyl-dimethylammonium chloride and an anionic polymer such as poly sodium-4-styrene sulfonate generate a coat of shielding on the surface. Subsequently, as fouling arose, the polyelectrolyte layer would be detached following a rinse with a highly concentrated brine solution. This would, therefore, regenerate a new layer in situ.

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Membrane Recycling

Desalination produces enormous amounts of life cycle completed membranes which usually goes to the landfills. The solid wastes that generate from reverse osmosis plants are usually used membranes which are typically disposed of in the landfills, and this requires appropriate management (Coutinho de Paula and Santos Amaral 2018). Recycling and repurposing different types of discarded reverse osmosis membranes could reduce the disposal of membranes in the landfills and can gain economic efficiency. The recycling method is constructed on the chemical oxidation of thin-film-composite of reverse osmosis membranes (Coutinho de Paula and Santos Amaral 2018). This was applied by dipping the membranes into market available sodium hypochlorite (NaClO) solution. The membrane recycling had comparable efficiency and features to porous membranes. Replacing a fresh ultrafiltration membrane which has a normal life of 5 years with a membrane recycling (with a projected lifecycle of 2 years) resulted in financial savings of 98.9%. Moreover, discarded reverse osmosis membranes associated with membrane fouling can be utilized as sustenance for bacterial biofilm attachment in the membrane bioreactor (Moron-Lopez et al. 2019). Microcystin cyanobacteria, for instance, will grow on the recycled membranes with biofilm instead when attached to the membrane bioreactor. The practice can extend the life of reverse osmosis membrane life as well as reduce costs.

6.3

Water-Energy Nexus and Desalination Technologies

Numerous communities living in the inaccessible part of the world lacks dependable freshwater and electricity sources. For instance, in developing countries, arid and rural coastal regions are generally defined by a deficiency of water for potable and/or agricultural consumption. Subsequently, significant unwanted public health impacts were recorded from the arid, semi-arid, and coastal communities residing in those areas owing to the presence of saline water. The management of water resources, therefore, stimulates nearly all pieces of the economy, public health, agricultural and food safety, water supply and hygiene, energy and economic development, as well as environmental sustainability (Babu et al. 2018). Improved factors of various physical and social issues like water governance have focused on a comprehensive idea of water security. Water security is described as appropriately clean and potable water allocations that always support human beings and ecosystems mutually (Gunda et al. 2019). In addition, there is a significant energy impression related to supply, treatment, and conveyance of water. One of the most efficient opportunities on earth is to yield fresh water from solar energy and seawater: the two most ample resources (Gençer and Agrawal 2018). For illustration, desalination of saltwater or brackish water appears to be a prospective resolution for meeting the water demand–supply equilibrium in many coastal countries. Iran, for

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instance, is the nation surrounded by the Caspian Sea at the north, and the south side is bordered by the Persian Gulf and Gulf of Oman (Gorjian and Ghobadian 2015). Owing to the exhaustion of existing fossil fuels and amplified emissions of greenhouse gases, the desalination technologies run by fossil fuels are not viable anymore to resolve water and energy security for many countries like Iran. To conclude the gap in water–energy security, Iran has explored the exceptional solar energy potential with efficient connectivity to run desalination operations of about 15.3 kWh/m2/ day. Energy generation requires a considerable amount of water consumption and vice versa. Therefore, the term is coined as the water–energy nexus, applied to depict the relationship between these two crucial utilities (IRENA 2015). More precisely, the mutual interconnection within energy production and water supply is oftentimes recognized as the water–energy nexus that ultimately ensures food security (Fig. 6.3). This is especially necessary to reduce desalination energy consumption and increase the yield of desalinated water (Deshmukh et al. 2018). The inferences for the energy–water nexus are often uncertain because forecasting and planning choices on investment in electricity and water are conventionally taken without a comprehensive strategy (Arndt et al. 2018). In this respect, technological progress can be a precious chance for improving growth in developing nations while at the same time meeting worldwide and local environmental goals for the developed world meeting the water–energy nexus taking into account the context of food production (Fig. 6.3).

Fig. 6.3 Diagram of the water–energy nexus in the global context ensuring water, energy, and food security for the rural and urban settlement

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Global Desalination Market in the Nexus of Water– Energy and Environmental Impacts

At present, 150 nations are functioning over 19,500 desalination plants to yield 100 million cubic meters fresh water per day to meet the clean water demand for 300 million population throughout the world (Shahzad et al. 2019). The Middle East and Africa have been at the forefront of the global desalination market, followed by the Asia-Pacific regions. More than 41% of the worldwide freshwater demand was recorded in the Middle East regions in 2017. Globally 15,906 operative desalination plants generate approximately 95 million m3/day of desalinated water for human purposes, of which 48% are manufactured in North and Middle East Africa (Jones et al. 2019). There are 15,906 operational desalination plants around the globe generating about 95 million m3/day of desalinated water for human use, 48% of which are situated in the Middle East and North Africa region. The structural modifications in energy frameworks triggered worldwide energy reversals due to significant advances in various renewable energy technology. Developing nations are significantly aimed at the swift acceptance of variable renewable energy systems that focus on sustainable development, including the nexus of water–energy systems (Arndt et al. 2018). In the year 2017, the global water desalination market size was approximately 15.43 billion USD, which is predicted to rise double the size by the year 2025 (Adroit Market Research 2019). According to Pouyfaucon and García-Rodríguez (2018), three scenarios can be considered for the thermal-powered desalination technologies: (1) off-grid rural communities with inadequate freshwater needs, (2) communities with higher needs of both water and electricity, and (3) transitional needs of water and energy demands. The application of sunlight for water treatment is a likely sustainable answer to address the problems of water deficiencies and the elevated energy demands in off-grid rural areas. The financial research of desalination plants conventionally focuses entirely on price rate and personal profits (Aparicio et al. 2018). Yet, significant numbers of mega-scale seawater desalination plants have been assembled in recent times in the water-stressed countries to strengthen freshwater availability (Elimelech and Phillip 2011), and the building of novel and innovative desalination plants is anticipated to rise in imminent times in many coastal areas of the developing countries around the world. In this context, when we compare the thermal versus membrane desalination technologies, by the year 2020, membrane desalination will reach its global capacity of 4 million m3/day according to Almar Water solution (2016). Thermal desalination is predicted to remain the same. Yet two major barriers of desalination technologies are concentrated brine production and energy consumption. Today brine production around the globe is 141.5 million m3/day which is 50% higher than former estimations and 55% of the worldwide share is in Saudi Arabia, the UAE, Kuwait, and Qatar (Jones et al. 2019).

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Reducing the Carbon Footprint of Desalination

The carbon footprint of membrane-based desalination is extremely high. The decarbonization of membrane desalination is an outstanding trial to decrease carbon discharges to alleviate climate change (Fane 2018). As discussed in the earlier sections, the use of forward osmosis and high-pressure reverse osmosis, as well as membrane recycling, can significantly reduce the carbon footprint of desalination. Enhanced membranes could decrease energy claim by 15–20% (Fane 2018). Moreover, the current thermo-desalination plants which work through multistage flash or multi-effect distillation plants are typically positioned with power plants run by fossil energy and have high environmental effects (Abdelkareem et al. 2018) which are the driving forces in using renewable energy for desalination purposes (Ali et al. 2018) in countless remote communities (Freire-Gormaly and Bilton 2018). Henceforward, globally, the close association among water–electricity demand pushes the renewable energy-driven desalination (Alhaj and Al-Ghamdi 2019). Considering the environment and operational efficiency perspective, renewable energy can be utilized in energy and freshwater production mutually around the globe. Several sources such as solar, geothermal, wind, and tidal waves are the key sources of usual renewable energy of concern for desalination (Ali et al. 2018). Renewable energy can help decrease the environmental and ecological impacts of seawater desalination methods crucial for achieving global water security in off-grid rural communities (Alhaj and Al-Ghamdi 2019). Wind energy, as well as solar energy coupled with a desalination plant, can facilitate ever-increasing water requirements with less carbon footprint than standard methods in a sustainable manner (Aminfard et al. 2019). A hybrid power system is an efficient replacement for the delivery of clean water and energy to off-grid societies in distant regions at a minimum electrical charge. The optimized hybrid system can also reduce greenhouse gas emissions (Wu et al. 2018). Hybrid solar photovoltaics–electrolysis and micro-desalination have been developed in connection to water–energy nexus (Kim et al. 2018). As the membranebased desalination activities are less energy-intense and low carbon-emitting, the considerable concern is rising to integrate it with photovoltaics–reverse osmosis (Ali et al. 2018). However, the determination of most efficient type renewable energy form is yet to be configured which can maximize desalinated water yield while consuming a minimum volume of energy (Abdelkareem et al. 2018). The main reason for this is the existence of diverse methods of desalination and renewable energies. The selection of renewable energy as a source is affected by many parameters, for instance, desalination plant size, place, feeder pressure, feedwater features, and anticipated water price (Ali et al. 2018). Significant attempts have been made to assimilate traditional renewables (wind, solar, geothermal, tidal, and nuclear) and comparatively latest green energy sources with membrane-based desalination, primarily reverse osmosis and electrodialysis (Ali et al. 2018). However, the sporadic traits of the mentioned sources encompass severe obstacles to the techno-economic feasibility of the process. Renewable sources are typically utilized

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for small-scale desalination plants, but do not apply to power big plants because they are sporadic and unstable (Mito et al. 2019). In addition, the reverse osmosis plant’s immediate link to renewables may require swiftness and/or modular operation to complement the load with the available power. More solar intensity does not, however, deduce decreased carbon dioxide emissions from solar desalination plants (Alhaj and Al-Ghamdi 2019). Environmental factors such as seawater temperature and salinity gradients also limit different factors of desalination.

6.3.3

Integrated Hybrid Power Generation for Desalination

Research progress in membrane technology such as pressure-retarded osmosis and reverse electrodialysis enables a high saline solution to be mixed with a low saline solution to produce electricity. The seawater, brine water, wastewaters are examples of the high saline solution. Membrane distillation has demonstrated the ability to produce electricity and fresh water simultaneously by itself. Coupling membrane distillation with pressure-retarded osmosis or reverse electrodialysis boosts the process performances and offers a smooth and viable route to produce fresh water and sustainable electricity from a range of wastewaters, together with brines which would otherwise be considered as environmental obligations (Ali et al. 2018). Plenty of inexpensive renewables are an alternative way to power contemporary procedures of desalination (Fig. 6.4). Furthermore, the development of a scalable and economic “photothermal membrane” for reverse osmosis, ultrafiltration, membrane distillation, and solar steam generation has potential driving power for an energy-efficient membrane distillation method (Jun et al. 2019). Moradi et al. (2019)

Fig. 6.4 (a) Possible uses of geothermal, solar, and wind/wave renewable energies in the diverse desalination techniques in the form of heat, mechanical, and electrical energy

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reported about simultaneous energy generation and thermal waste retrieval scheme in a multi-effect desalination system. A heat recapture vapor generator is used for vapor production utilizing thermal energy from the outlet. The produced steam passes into a multistage desalination system for freshwater yield (Moradi et al. 2019).

6.3.4

Solar Desalination Technologies

Taking up the challenge of global water–energy security nexus, a systematic inspection on the contribution of renewable energies and different sources of water desalination techniques should be explored (Alhaj and Al-Ghamdi 2019). Use of solar photovoltaics can radically decline the desalination effects allied with energy. Desalination plants can be operated by solar energy directly or indirectly (Zhang et al. 2018). Solar energy is used directly to generate distilled water in the solar collector which is a direct solar desalination process. “Solar still” is the most significant example of the abovementioned process (Fig. 6.4). In contrast, solar energy is collected by “solar thermal collectors” or photovoltaic panels to harvest energy for thermal heat transport to the membrane desalination technology such as multi-effect desalination, multistage flash desalination, membrane distillation, or reverse osmosis. A serious question that arises by Elimelech and Phillip (2011), “is seawater (or brackish water, wastewater) desalination a justifiable to global water shortages?”. There are numerous apprehensions with desalination since there are likely adverse environmental impacts. The chief energy source for seawater reverse osmosis desalination unit is the thermoelectric power which outcomes particulate matter in the air, emission of greenhouse gas that moreover exacerbate in climate change. Use of reverse osmosis technology combined with solar photovoltaic systems is a financially feasible alternative to brackish water (and seawater) desalination and can overcome the current water and energy crises (Taha and Al-Sa’ed 2017). Independent photovoltaics–reverse osmosis wastewater treatment plants are promising to secure fresh water, especially in areas that require a usable electricity network (Abdelkareem et al. 2018). It appears that effective arrangement of photovoltaics cluster and addition of solar tracking, adjustment of tilting angle, photovoltaic array, cleaning technologies can effectively expand the effectiveness of independent photovoltaics–reverse osmosis yield. With feedwater for cooling the photovoltaic system, increased photovoltaics efficiency and freshwater output in the reverse osmosis unit raised cumulative productivity while decreasing total costs of function (Table 6.2) (Abdelkareem et al. 2018). Nevertheless, many of photovoltaics–reverse osmosis systems function spasmodically to cut the expenditures of renewable energy storage. Consequently, the influence of sporadic function influences the membrane life, develops fouling and scaling, as well as increases maintenance operational costs (Freire-Gormaly and Bilton 2018). A simple procedure of rinsing the membrane before shutdown can

Social and private costs of desalination plant of brackish groundwater: San Vicente del Raspeig, Spain

Photovoltaics–reverse osmosis desalination system; Marj Naajeh desalination unit, Jordan Valley, West Bank, Palestine Combined power generation system, coupled with multieffect desalination Comparative lifecycle analysis of solar-driven desalination system operated in Kuwait; Algeria; Abu Dhabi, UAE; Torrevieja, Spain; Carlsbad San Diego, CA; Sydney, Australia; Escondida, Chile

System studied Photovoltaics/diesel/battery/reverse osmosis desalination system; Khorasan, Iran

Not reported

Not reported Not reported

Not reported

Life cycle costs in $ $28,130

Not reported

Not reported

Not reported

CO2 emission: 4.32 kg CO2 eq./m3 of desalinated water. 47% lesser than thermal desalination Per unit price of desalinated water: 0.29 €/m3 Social benefits provided by the relaxation zone: 0.51 €/ m3

Ultimate costs of desalinated water 0.22 €/m3

Not reported

Levelized cost of energy 0.3975 $/kWh and 0.5975 $/kWh. Not reported

Not reported

Per unit supply costs Between 1.59 $/m3 and 2.39 $/m3 US$0.183/m3– 0.346/m3

Freshwater yield 143.64 kg/h

Breakdown costs Fuel, photovoltaics and battery systems input: 35%, 13%, and 12%, respectively Not reported

Table 6.2 Different types of solar-driven desalination and their reported costs and efficiency

Aparicio et al. (2018)

Taha and Al-Sa’ed (2017) Moradi, et al. (2019) Alhaj and Al-Ghamdi (2019)

References Wu et al. (2018)

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significantly improve the membrane life and preserve membrane permeability. Freire-Gormaly and Bilton (2018) further designated that using antiscalant and rinsing can significantly lower scale deposits on the membrane which leads to membrane fouling. Li et al. (2019) claimed that 20-year lifetime for photovoltaics-powered smallscale membrane systems for brackish water desalination is possible through the adaptation of three different systems (Li et al. 2019): 1. Nifty module-based functioning range, consistency, and lifespan choice 2. Operational design improved to enhance performance under changing circumstances 3. Accurate choice of membranes for the system process and preservation approach Revolutionary innovations in novel membrane nanomaterials and defect-free production methods are crucial to accomplish the goal of efficient solar desalination technologies (Alvarez et al. 2018). In the conceivable future, membrane-based separation and desalination techniques may stay critical instruments in water treatment. In addition, nanotechnology will assist to overcome working interferences, namely, foulants and poor selectivity in the separation of ions or molecules. In nanoscale grafting and doping, for instance, membrane efficiency and reliability will be improved, while selective membranes for the specific desalination tasks can be built using nanochannels. Next-generation desalination membranes will be antifoulant, extremely discerning, and chemically stable to oxidizing agents such as chlorine. Currently, most of the solar–water treatment processes are yet to be implemented for real-time applications. For example, Jun et al. (2019) examined uses of biofouling resistance in photothermal-induced inactivation of microorganisms in reverse osmosis and ultrafiltration procedures (Jun et al. 2019). Again, up-to-the-minute technologies often need complex systems with several parts that result in poor efficiency yet at an extremely high price (Yang et al. 2018). Consequently, economic affordability is a significant reason for scaling and marketing (Zhang et al. 2018). For instances, price of water from a small to medium-scale photovoltaics–reverse osmosis plants are in the range of US$0.2–22/m3 which was greater than traditional fossilfueled plants (Zhang et al. 2018). However, the evaluated low price of water (US $0.9–2.2/m3) for large-scale photovoltaics–reverse osmosis plants indicates alternative options will be available soon.

6.3.5

Nuclear Energy for Desalination

Nuclear energy proposes an achievable choice for simultaneous thermal energy generation and freshwater coproduction. A substantial quantity of heat is recovered during the process which can be utilized for freshwater reclamation. The recovered heat is employed to harvest vapor and make thermal electricity onsite to provide thermal power as well as facilitate membrane desalination. Different sizes of nuclear

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reactors can be used to power desalination. According to Al-Othman et al. (2019), like the photovoltaics–reverse osmosis desalination processes, nuclear energy can be coupled with multistage flash, multi-effect distillation, and reverse osmosis. The charge was calculated in the range of $0.4–1.8/m3 using nuclear power based on the reactor category and the procedures for desalination (Al-Othman et al. 2019). Small modular reactors appear to be the safest selection in nuclear desalination, as contrasting to medium/larger nuclear reactors particularly for “newcomer” nations in nuclear energy yet to have skills in nuclear power plant facilities.

6.3.6

Nano-Enabled Desalination

Nanotechnology provides novel innovative products for water treatment systems by regulating the material, size, morphology, and chemical composition. Nano-enabled water treatments possess excellent catalysts that increase the cost-effectiveness of water treatment with adsorption, optical, quantum, electrical, and/or antimicrobial properties (Table 6.1) (Alvarez et al. 2018). As an alternative to polyamides, many innovative nanoporous materials are being created (Stevens et al. 2017) which are exclusively used for reverse osmosis and nanofiltration membrane production. The development of distinct kinds of nanomaterials such as graphenes, graphene oxide, block copolymers, liquid crystals, aquaporins, and other biologically inspired molecular channels and thin-film composites and carbon nanotubes (Stevens et al. 2017; Yang et al. 2019) is an increasing field of research aimed at meeting the cumulative requirement of fresh water around the globe (Table 6.3). These carbon-based nanomaterials often have a bigger surface area and flexibility and have intrinsic adsorption and sieving ability to eliminate waterborne contaminants and minerals (Table 6.1). Because of their lightweight nature, these materials are readily portable for point of use devices (Alvarez et al. 2018) and would be ideal for both household and business-scale applications (Yang et al. 2018). Ali et al. (2019) revealed the bench-scale implementation of carbon nanotubes nanocomposite membranes which enhanced water penetration, high selectivity, and antifouling capacity (Ali et al. 2019). Nano-enabled treatment methods indicate that energy and chemical needs, as well as solid waste and wastewater residuals generation, the associated cost, and likely ecological effects, may be considerably reduced. For instances, nanophotonic-based technology can be combined with renewable energy sources to harvest fresh water in the distillation process. Also, to reduce interference and improve effectiveness and interruptions, nanotechnology can strengthen membranes and photocatalysts (Alvarez et al. 2018). However, the evolving upsurge of the invention can ultimately enable next-generation modular water treatment technologies to considerably enhance water security and resilience in the water supply (Alvarez et al. 2018). Besides, functional carbon nanotubes membranes further revealed superior antifoulant repellant because of the characteristics owing to elevated conductivity of electrical charge and the presence of adverse surface charge. By creating reactive

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Table 6.3 Different types of recently developed novel nanomaterials for RO desalination Material(s) 3D cross-linked honeycomb graphene Hybrid graphenenanomesh/single-walled carbon nanotube, ultrathin membrane Polyamide based thin film composite; singlewalled carbon nanotubes–polyether sulfone composite; metal–organic framework nanoparticle (ZIF-8) Thin-film composite with embedded polyester screen support Polyamide based thinfilm composite

Membrane type Nanofiltration

Nanofiltration

Efficiency/ performance The water production rate of 2.6 kg h1 m2 g1 High salt rejection

Application Sewage water, seawater

References Yang et al. (2018)

Reviewed all types of water

Yang et al. (2019)

Forward osmosis– nanofiltration

High permeance up to 53.5 l m2h1bar1 Salt rejection above 95% for Na2SO4

Brackish groundwater and wastewater reuse

Wang et al. (2018a, b)

Forward osmosis– nanofiltration Pressureassisted osmosis

Water permeability flux increased by 17% Feed pressure should be higher than the draw pressure Energy consumption: 4.00kWh/m3. Concentrate brine streams up to 125 g/L Salt flux exhibited a low value of 0.10/ 0.09 g/L)

Wastewater reclamation

Jamil et al. (2018)

Seawater and wastewater

Kim et al. (2018)

Recoveries of up to 72% from a 35 g/L saline feed at 48.3 bar Seawater Seawater/ wastewater

Peters and Hankins (2019)

Thin-film composite

Osmotically assisted reverse osmosis

Thin-film composite sulfonated polysulfone/ poly(vinyl chloride)

Forward osmosis

Zheng et al. (2018)

oxygen species, functional carbon nanotubes membranes impair microbes and repulse biofilm growth responsible for fouling (Ali et al. 2019). Graphene is another illustration of carbon-based substance that functions as solar–thermal converter. 3D cross-linked honeycomb graphene foam material captures and converts solar energy into heat. Subsequently, the heat distils water into steam and generates fresh water effectively even under low sun intensity (Yang et al. 2018).

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Desalination for Water Security in Water-Energy Nexus for Developing Country

Climate change is altering the composition of the natural environment and has the most severe impact on world water resources. As stated by Intergovernmental Panel on Climate Change, climate observation records and forecasts show that freshwater resources are susceptible to potentially heavily affected by climate change with broader implications for human communities and ecosystems (Bates et al. 2008). One of the crucial forecasts of climate change sea level rise is the invasion of saline water in soil, surface, and groundwater (Green et al. 2011; Taylor et al. 2013) especially in the coastal aquifers, being the primary interface between the oceanic and terrestrial system and subjected to overexploitation (Small and Nicholls 2003). South Asian coastal polders are enormously vibrant with complex socio-ecological organizations where numerous poverty-driven vulnerable communities dwell. According to the Bangladesh Demographic and Health Survey (BDHS 2014) data, all households in Bangladesh have access to an improved source of drinking water (Fig. 6.5) to meet the United Nations Sustainable Development Goals target of water security. However, the reality might be different. Lingering risks from salinity hazard, cyclones, and tidal flooding in coastal Bangladesh put enormous governance difficulties to guarantee secured potable water for communities with elevated and permanent rates of poverty (Hoque et al. 2019). People are at risk of developing hypertension leading to cardiovascular diseases due to consumption of saline

Fig. 6.5 Entrée to enhanced drinking water sources is almost universal in all over Bangladesh as well as in urban areas. In rural areas, however, access to better-quality sources of drinking water increased from 92% in 1994 to 97% in 2014. (Data source: Bangladesh Demographic and Health Survey (BDHS) 2014)

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tubewell water (Shammi et al. 2019; Vineis et al. 2011). Moreover, the concentration of sodium in the saline water is responsible for an elevated risk of high blood pressure leading to pre-eclampsia and eclampsia among the expectant females as well as a leading cause of newborn death in coastal districts of Bangladesh (Shammi et al. 2019). Moreover, Bangladesh surpassed the limit of water scarcity threshold at least three decades earlier defined by Falkenmark and his co-workers (Falkenmark et al. 1989). By 2025 Bangladesh will reach the threshold of absolute water scarcity and by 2050 well below the limit of complete water shortage. It is further important to note that the existing electricity-generation portfolio in a developing country like Bangladesh is profoundly thermal-based. Accordingly, it is important to remember that substandard technology implementation in many water-stressed countries ultimately causes a burden to the economy with higher energy prices per unit (Shahzad et al. 2019), and this, in the end, will not ensure water security as well leading to serious public health issues. Therefore, it is important to plan water–energy nexus strategically and carefully for a country like Bangladesh, in which coastal communities use higher sodium-containing saline groundwater for drinking and day to day purpose. Despite the elevated price and power, the potential of photovoltaics–reverse osmosis-based desalination remains on a pilot stage in many coastal communities of Bangladesh. Shammi et al. (2019) and Shamsuzzoha et al. (2018) reported the presence of pilot-scale reverse osmosis plants from Kalapara Upazila, Patuakhali District, and Patharghata Upazila, Barguna District. The photovoltaics–reverse osmosis system, in this regard, may have a positive influence on the disaster risk reduction-based program to supply the coastal household with fresh water and growing resilience in water security-related issues. Nevertheless, the environmental impacts associated with desalination plants should be considered with high emphasis. In fulfilling the cumulative needs of water security, developing countries, as well as rural communities of coastal and rural Bangladesh, can use solar energy and photothermal seawater membranes (Jun et al. 2019). This method of obtaining desalinated water will be cost-effective, fitting for an off-the-grid decentralized coastal location, easy-to-use with time-variant flexibility, appropriate for potable water solutions that can decrease standard power usage, as well as fulfilling the energy requirements. Besides, Bangladesh is implementing nuclear energy to accomplish its energy vision “electricity for all by 2021” as well as it aims to generate 2000 megawatts electricity from renewable energy by 2020. In this regard, a long-term strategy to implement desalination based on photovoltaics–reverse osmosis along with the nuclear desalination can be a strategic move to solve both water and energy crises simultaneously. As discussed in the earlier section, the use of small modular nuclear reactors can be the safest selection for a nuclear “newcomer” country like Bangladesh, who is yet to have skills in nuclear power plant facilities.

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Concluding Remarks

In the realms energy–water nexus, the desalination system should be planned carefully to overcome serious errors in the choice of the desalination method. To achieve water security, we need to shift our conventional water uses to unconventional water sources such as industrial greywater, wastewater, brackish groundwater, seawater, and stormwater. To achieve sustainable development goal, we need to consider energy sources less carbon-emitting as well as zero liquid discharge water to tackle climate change and industrial pollution. Desalination can be the answer to achieve water security and Sustainable Development Goal targets. However, we should always remember that desalination should be greener and environmentally viable. Discharging brine from the desalination plants may subsequently put a strain on the coastal and marine ecosystems as well as the vulnerable mangrove ecosystems if present. Moreover, among the numerous desalination methods, the thermal desalination techniques are most energy-intensive with large carbon footprint and not sustainable for water-energy nexus to achieve water security. For membrane-based desalination, although the industry is growing fast, membrane fouling still possesses a serious problem. Generation of end-of-life solid membrane wastes is another serious problem which usually goes to landfill as well as possessing massive carbon footprint. The major obstacles of desalination plants are vastly expensive in terms of water– energy nexus and certainly not justifiable strategically in many energy- and waterstarved worlds. Choosing the right type of desalination should be strategically planned, designed, and implemented to achieve water security. Although the desalination market is steadily growing around the globe based on membrane-based reverse osmosis desalination, integration of renewable energy such as photovoltaics– reverse osmosis, scheming nanomaterial for energy-efficient membranes, forward osmosis, and hybridization of different desalination processes can lessen the carbon footprint and enhance the cost-effectiveness as well as increase water security.

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Chapter 7

Nanotechnology for the Remediation of Heavy Metals and Metalloids in Contaminated Water Roop Singh Lodhi, Subhasis Das, Aiqin Zhang, and Paramita Das

Contents 7.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.2 Heavy Metals and Metalloids Contaminating Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.2.1 Types of Heavy Metals and Metalloid Pollutants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.2.2 Sources of Heavy Metals and Metalloids . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.3 Hazardous Effects of Heavy Metal- and Metalloid-Contaminated Water on Human Health, Plants, and Aquatic Environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.3.1 Hazardous Effects on Human Health . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.3.2 Hazardous Effects on Plants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.3.3 Hazardous Effects on Aquatic Environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.4 Nanotechnologies of Remediating Heavy Metals and Metalloids . . . . . . . . . . . . . . . . . . . . . . . 7.4.1 Conventional Nanotechnologies of Remediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.4.2 Membrane Filtration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.4.3 Electrochemical Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.4.4 Bioremediation Technologies of Heavy Metals and Metalloids . . . . . . . . . . . . . . . . . 7.4.5 Phytoremediation Technologies of Heavy Metals and Metalloids . . . . . . . . . . . . . . . 7.5 Nanomaterials for Heavy Metal and Metalloid Remediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.5.1 Metal–Organic Frameworks as Adsorbents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.5.2 Natural Clay Material as Adsorbent . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.5.3 Activated Carbon as Adsorbent . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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R. S. Lodhi · P. Das (*) Department of Chemical Engineering, Indian Institute of Science Education and Research Bhopal, Bhopal, Madhya Pradesh, India e-mail: [email protected]; [email protected] S. Das Environmental and Industrial Biotechnology Division, The Energy and Resources Institute, New Delhi, India e-mail: [email protected] A. Zhang State Laboratory of Surface and Interface Science and Technology, Henan Collaborative Innovation Center of Environmental Pollution Control and Ecological Restoration, Zhengzhou University of Light Industry, Zhengzhou, People’s Republic of China e-mail: [email protected] © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature Switzerland AG 2021 Inamuddin et al. (eds.), Water Pollution and Remediation: Heavy Metals, Environmental Chemistry for a Sustainable World 53, https://doi.org/10.1007/978-3-030-52421-0_7

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7.5.4 Low-Cost Adsorbents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.5.5 Nano-sized Carbonized Waste Biomass as Adsorbents . . . . . . . . . . . . . . . . . . . . . . . . . . 7.5.6 Cellulose-Based Materials as Adsorbents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.6 Conclusion and Future Trends . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract The environment has been seriously polluted by heavy metals and metalloids, and it has become one of the most severe problems today. This affects human health, plants, aquatics, air, and soil. Heavy metals are mainly naturally occurring compounds, but anthropogenic activities increase their concentration level in different environmental compartments. The remediation of heavy metals and metalloids is extremely needed as the high level of contamination, caused by the heavy metals, poses serious threats to the environment. In past years, various technologies for the remediation of heavy metals and metalloids in contaminated water have been extensively studied. In this book chapter, we have discussed about the heavy metals and metalloids which contaminate water; described their harmful effects on human health, plants, and aquatic environments; and presented several nanotechnologies as well as nanomaterials used in heavy metal and metalloid remediation of contaminated water. Keywords Functional nanomaterials · Heavy metals · Metalloids · Nanotechnologies · Remediation · Wastewater treatment

7.1

Introduction

Recently, environmental pollution is one of the most common problems that our society faces because of increased urbanization; industrialization; energy generation; poor industrial, agricultural, as well as domestic waste management; and other anthropogenic modes (Shawai et al. 2017). Pollutants can be either naturally occurring or foreign matter which, when present in the environment above their limiting concentration, cause harmful effects in the environment by bringing harmful changes in physical, chemical, and biological characteristics of air, water, and soil (Masindi and Muedi 2018). Pollutants can be broadly classified into three types— inorganic, organic, and biological pollutants. Industrial, agricultural, and domestic wastes mainly contribute to the inorganic wastes which are usually metals, salts, and minerals. Heavy metals present in nature are the elements having atomic weight higher than 40 and specific density of more than 5 g/cm3. Some metalloids, transition metals, basic metals, lanthanides, and actinides are considered as heavy metals. Metalloids are the type of chemical elements which show a combination of both metallic and nonmetallic properties (Srivastava and Goyal 2010). These inorganic pollutants can be introduced in the environment through various ways, such as mine

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drainage, smelting, and chemical, metallurgical, as well as natural processes. Heavy metals are highly soluble in water and can easily be absorbed by the living organisms which pose a threat for them due to heavy metal accumulation in the food chain above toxic level as well as their nonbiodegradability. If the heavy metals are not metabolized by the human body and get accumulated beyond the critical tolerance limit in soft tissues, they become toxic. Therefore, it is highly required to treat the heavy metals and metalloids in contaminated water before discharging them directly into the environment. They enter into the human body possibly through air, water, and food or absorption through skin contact in industrial, agricultural, residential, or other places (Masindi and Muedi 2018). Heavy metals and metalloids can be removed from the inorganic effluent by several remediation technologies, such as precipitation, adsorption, cementation, solvent–solvent extraction, ion exchange, membrane filtration, electrodialysis, electrochemical treatment, bioremediation, phytoremediation, etc. Nanotechnology has become a new promising research area because of the outstanding physical properties of materials at nanoscale, such as higher surface-to-volume ratio of nanomaterials which leads to higher reactivity, better selectivity, as well as efficiency when compared with their bulkier counterparts. Hence, remediation technologies based on nanomaterials and microorganisms are highly desired for the removal of contaminants from air, water, and soil. Several nanotechnologies and different types of nanomaterials used for the remediation of heavy metals and metalloids in contaminated water have been discussed in this book chapter.

7.2 7.2.1

Heavy Metals and Metalloids Contaminating Water Types of Heavy Metals and Metalloid Pollutants

Heavy metals which contaminate water are generally lead (Pb), chromium (Cr), mercury (Hg), nickel (Ni), zinc (Zn), cadmium (Cd), molybdenum (Mo), copper (Cu), cobalt (Co), and manganese (Mn) (Salem and El-Fouly 2000), whereas arsenic, selenium, and antimony are among the metalloids (Peterson et al. 1981). Among them some of the heavy metals like calcium, magnesium, potassium, and sodium are necessary metals for human health to make some bichemical functions in the body. Heavy metals like manganese, copper, molybdenum, iron, cobalt, zinc, and silver (relatively in lesser amount) are essential for living species as their metabolic nutrients. They act as catalysts for enzymatic activities at low concentrations. However, some heavy metals have strong affinity for sulfur when they bind via thiol (–SH) group. In animal body this thiol group hinders the functioning of enzymes due to which metabolic activities are affected (Hadia-e-Fatima 2018). Similarly, chromium, cadmium, mercury, lead, and arsenic have higher potential to cause hazardous effect to the environment when they accumulate in higher concentration (Ali and Khan 2019). Therefore, it is highly required to remove

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heavy metals and metalloids from industrial wastes before discharging them into waterways as well as prior to their disposal into ocean or land.

7.2.2

Sources of Heavy Metals and Metalloids

Heavy metals occur largely as sulfides, oxides, carbonates, and silicates in nature (Dean et al. 1972). All the heavy metals and metalloids are categorized into two types based on their sources of occurrence: natural sources and anthropogenic sources (see Fig. 7.1). Several natural processes like terrestrial and marine volcanic eruptions; forest fires; erosion; surface winds; meteoric, biogenic, rock weathering; etc. are the sources of the emission of heavy metals and metalloids into different environmental bodies (soil, water, and air). On the other hand, anthropogenic sources including mining, tailings, industrial wastes, agricultural runoff, occupational exposure, paints, treated timber (Maxwell 2007), lead–acid batteries, high traffic loads (Harvey et al. 2015), and microplastics floating in the oceans (Pizzolato et al. 2014; Brevik and Burgess 2012) can also lead to the release of pollutants to different environmental bodies. Among the anthropogenic sources, industries are the most primary sources of heavy metal and metalloid contamination due to emission of polluted air and discharge of wastewater.

Fig. 7.1 Sources of heavy metals and metalloids and their cycle in the environment. The category of heavy metals and metalloids based on their sources of occurrence and the contamination caused by them have been shown

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Table 7.1 Heavy metals and metalloids obtained from various sources Metals or metalloids Copper Cadmium

Lead

Mercury

Chromium Antimony

Arsenic

Natural sources Volcanic eruptions, windblown dust, and forest fires Respirable air-borne size particles from mines easily transferred by food chain, rain water and wind Dust of minings, soil by gasoline, rock weathering Volcanoes, forest fires, cinnabar (ore), and fossil fuels, such as coal and petroleum, atmospheric by rainfall Leaching from topsoil, volcanic eruption and rocks weathering Volcanic eruption, soil erosion, and volcanic eruption Marine sedimentary rocks, volcanic rocks, hydrothermal ore deposits and associated geothermal waters, ash, fossil fuels, coal

Anthropogenic sources Pesticides, fertilizers Paints and pigments, electroplating, phosphorus fertilizers Battery manufacture, combustion of leaded petrol, insecticides, herbicides Released from gold and antimony mining, combustion of coal, medical waste Tanneries, steel industries, and fly ash Coal combustion, smelting and mining processes, soil erosion, volcanic eruption Pesticides and wood preservatives

Fashola et al. (2016), Chibuike and Obiora (2014), Sawyer and Parkin (2003), Uzma and Chandio (2017), Blais et al. (2008), An and Kim (2009), Germ et al. (2007), Ayangbenro and Babalola (2017), Ali et al. (2013a), and Wang and Mulligan (2006)

Herein, air pollutants cause water pollution indirectly by releasing a number of heavy metals into the atmosphere, which, after dry or wet deposition, become potential sources of water pollution. Agricultural soil is polluted mostly by heavy metal-contaminated water coming from various phosphate-based fertilizers, pesticides, and decaying plant and animal residue. Therefore, agricultural runoff together with soil erosion is the potential source of water pollution. Heavy metals and metalloids obtained from various sources are shown in Table 7.1. Heavy metals and metalloids like lead and arsenic are present in children’s toys where lead is used as stabilizer, color enhancer, and anticorrosive agent, and arsenic is used with dyeing agent. Sometimes toys contain heavy metals in excess causing environmental pollution when dumped and disposed (Braxton et al. 2018). Lead is a common toxic heavy metal used in making pipes, drains, and soldering materials. Thus, lead can dissolve in drinking water and contaminate it if water pipes corrode (Uzma and Chandio 2017). As cadmium is found in various sources, such as in most rocks at a low concentration, coal, petroleum, geologic deposits, acidic waters, paint and pigments, corrugated galvanized pipe, plastic products as stabilizers, landfill leachates, emissions from fossil fuel, fertilizers, and disposed sewage sludge can contaminate groundwater largely. Cadmium is also obtained as a by-product during mining as well as smelting of lead and zinc and found to be used in nickel–cadmium batteries. Similarly, chromium comes from various sources like industrial, agricultural, paints, cement, paper, rubber, electroplating, chromic

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acid spray, and airborne chromium trioxide. The toxicity of chromium increases in the presence of zinc, iron, cobalt, and copper in drinking water. Arsenic, in the form of III and V oxidation state, is mostly found in water and is also considered as a very toxic heavy metal (Jain and Ali 2000). There are several industrial processes, such as smelting process of zinc, copper, and lead and manufacturing of chemicals and glasses, which are responsible for the release of arsenic into the environment. Arsenic can also be sourced from paints, fungicides, preservatives, and by-product of pesticide manufacturing. It is also found in rat poison which is used for grain storage and, thus, can contaminate soil and water via leaching (Goyer and Mahaffey 1972). Arsenic causes acute heavy metal poisoning in adults. Another heavy metal pollutant is mercury which is found in volcanic emissions and exists in elemental, organic, and inorganic form. Atmospheric mercury can be dispersed and returned back to earth during rainfall. Hence, it can accumulate in aquatic food chains and cause water contamination. Mercury is also used in thermometers, thermostats, and dental amalgam. Gold amalgam which was used in gilding leads to large number of casualties among the workers. It is found that certain bacterial conversion is responsible for mercury contamination as it transforms mercury into methylmercury which can be concentrated in the food chain causing malformations. Antimony and selenium are released from smelting, mining, coal combustion, volcanic eruption, and soil erosion (Blais et al. 2008; An and kim 2009; Germ et al. 2007). Similarly, beryllium is released from oil and coal combustion and also obtained through volcanic dust (Ayangbenro and Babalola 2017; Blais et al. 2008). Paint industries, printing, plating, mining process, and copper polishing are the sources of copper, whereas electroplating, porcelain enamelling, and non-ferrous metals lead to the release of nickel (Fashola et al. 2016; Chibuike and Obiora 2014; Salem et al. 2000). Battery, mining, photographic processing, and smelting are the main sources of silver, whereas brass manufacturing industries, oil refinery sector, mining, and plumbing are the sources of zinc (Prabhu and Poulose 2012; Qian et al. 2013; Chibuike and Obiora 2014; Gumpu et al. 2015).

7.3

Hazardous Effects of Heavy Metaland Metalloid-Contaminated Water on Human Health, Plants, and Aquatic Environment

In this section, we have discussed the hazardous effects of heavy metal- and metalloids-contaminated water on health, environment, and aquatics in detail.

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Hazardous Effects on Human Health

Hazardous effects on human health are observed when people are exposed to heavy metals though different ways, such as inhalation, ingestion, and/or skin contact. The harmful effects depend on several factors like the type, nature, and chemical form of contaminant, exposure time, and dose. Heavy metals and metalloids, when accumulated in higher concentration than permissible values in living organisms, exert detrimental effects (see Fig. 7.2). These compounds start to accumulate when intake becomes higher than excretion by metabolism, thus leading to interference in various enzymatic activities. Heavy metal toxicity can weaken nervous function, decrease the energy levels, and damage the liver, kidneys, lungs, and other vital organs. Longterm exposure to such toxic compounds leads to slow progress of some physical, muscular, and neurological degenerative processes which may finally cause muscular dystrophy, Alzheimer’s disease, Parkinson’s disease, and multiple sclerosis, while short-term exposure causes nausea; vomiting; cramping; sweating; headaches; breathing problem; impaired cognitive, motor, and language skills; etc. (Jaishankar et al. 2014a, b). Heavy metals and metalloids are used in industrial, agricultural, medical, and housing construction or domestic sectors. As a result, they are widely present

Fig. 7.2 The effects of heavy metals and metalloids on human body. The detrimental effects caused by heavy metals and metalloids, when accumulated in higher concentration than permissible values in living organisms, are shown

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Table 7.2 Maximum permissible limits of heavy metals and metalloids in drinking water according to various agencies (all concentrations in ppm) Metals/ Metalloids* Copper Cadmium Lead Mercury Chromium Antimony* Arsenic*

WHO 2.00a, c 0.003a, c 0.01a, c 0.001a, b 0.05a, c 0.005a 0.05b

IS 10500 0.05a 0.003a 0.01a 0.001a 0.05a ... 0.05d

EU standard 2.00a, c 0.005a, c 0.01a, c 0.001a, c 0.05a, c 0.005a, c 0.010c

CDW, Canada 1.00a, c 0.005a, c 0.01a, c 0.001a, c 0.05a, c 0.006a, c 0.01c, d

USEPA, USA 1.30b, c 0.005a, b, c 0.015a, c 0.002a, b, c 0.1a, b, c 0.006a, c 0.05b

NHMRC, Australia 2.00a 0.002a 0.01a 0.001a 0.05a 0.005a ...

MEP, China 1.00a 0.005a 0.01a 0.00a 0.05a 0.005a ...

*Indicates the metalloids Gu et al. (2019), bKumar and Puri (2012), cFernandez-Luqueno et al. (2013) and dAli et al. (2013b)

a

everywhere in our environment. Water pollution caused by heavy metals is very prominent in areas where mining, smelting, metal processing and refining, and paper and pulp processing facilities are located. The exponential rise of heavy metal and metalloid application creates an alarming concern among public due to increased possibility of human exposure to heavy metals along with several health risks. Chromium, mercury, cadmium, arsenic, and lead are the most toxic heavy metals and metalloids found in nature and are also nonbiodegradable and thus can affect the environment both directly and indirectly (Wang and Shi 2001). Chronic exposure of such toxic metals can cause severe damages in the brain, liver, lung, kidneys, and bones in the human body (Tchounwou et al. 2012). The effects of heavy metals and metalloids on human body are shown in Fig. 7.2. Lead is the most common heavy metal contaminant and is present excessively in gasoline in the form of tetraethyl lead. This form of lead contaminates water indirectly first by adsorption or deposition, and then it is transferred to soil followed by water runoff (Magda 1996). Diseases like renal failure may be caused due to lead and cadmium present in drinking water (Salem and El-Fouly 2000). Cumulatively, when lead concentration increases above permissible level, it may not only cause permanent damage to the central nervous system, brain, and/or kidneys but also lead to death (Jennings et al. 2002). Such damage results in memory and concentration problems, behavior and learning problems, (e.g., hyperactivity), high blood pressure, headache, hearing problems, slow growth, reproductive problems, digestive problems, and muscle and joint pain (Salem and El-Fouly 2000). Table 7.2 shows the maximum permissible limit of heavy metals and metalloids in drinking water according to various worldwide agencies like WHO which represents World Health Organization, IS 10500 represents the drinking water standards in India, EU Standard represents the drinking water standards in European Union, CDW represents the drinking water standards of Canada, USEPA represents the drinking water standards of United States Environmental Protection Agency, NHMRC represents National Health and Medical Research Council in Australia, MEP represents

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Ministry of environmental protection of China (Gu et al. (2019), Kumar and Puri (2012), Fernandez-Luqueno et al. (2013), Ali et al. (2013b)). Cadmium causes renal dysfunction, human hypertension, and cardiovascular diseases. It is reported that the effects of cadmium are mainly on the bone, brain, kidney, liver, and lung. At low concentration, cadmium causes cough, headache, and vomiting, whereas at large concentration, it leads to renal failure and bone disorders by replacing calcium. Antimony causes nasal ulceration, dermatitis, cancer, conjunctivitis, and liver, respiratory, and cardiovascular diseases (Blais et al. 2008; An and kim 2009). Beryllium causes heart diseases, allergic reactions, berylliosis, cancer, and lung diseases (Blais et al. 2008; Ayangbenro and Babalola 2017). Copper causes headache, abdominal pain, metabolic disorders, anemia, diarrhea, nausea, liver and kidney damage, and vomiting. Nickel causes dermatitis, shortness of breath, dizziness, cardiovascular diseases, lung and nasal cancer, chest pain, kidney diseases, nausea, dry cough, and headache (Fashola et al. 2016; Chibuike and Obiora 2014; Salem et al.’ 2000). Nickel and chromium show similar toxic properties. Chromium causes gastrointestinal hemorrhage, hemolysis, acute renal failure, and pulmonary fibrosis. Chromium in its hexavalent form is more toxic. Excessive dose can cause direct damage to the skin and lungs, whereas long-term exposure leads to kidney failure, liver damage, circulatory and nerve tissue (Hyodo et al. 1980). Selenium causes diseases like liver damage, gastrointestinal disturbances, loss of natural killer cell activity, and dysfunction of the endocrine system (Germ et al. 2007). Silver is responsible for mental fatigue, argyria and argyrosis, bronchitis, nose, cytopathological effects in fibroblast and keratinocytes, throat and chest irritation, emphysema, knotting of cartilage, and rheumatism (Prabhu and Poulose 2012; Qian et al. 2013). Zinc causes hematuria, impotence, lethargy, metal fume fever, icterus, kidney and liver failure, macular degeneration, seizures, ataxia, depression, gastrointestinal irritation, prostate cancer, and vomiting (Chibuike and Obiora 2014; Gumpu et al. 2015). Arsenic mostly affects the blood, kidneys, digestive and central nervous systems, and skin. It causes breathing problems, decreased intelligence, skin cancer, nausea, diarrhea, vomiting, and peripheral nervous system problems and even can lead to death if exposed to high levels. Due of arsenic-tainted beer, over 6000 people were poisoned in Manchester area of England in 1900, and as a result, more than 70 people were killed in this incident. Another report says that millions of people in Asia were poisoned due to arsenic-contaminated groundwater until 2014 (Fatoki et al. 2013). Mercury mostly affects the brain and kidneys and is mainly responsible for blindness, deafness, mental retardation, brain damage, kidney damage, and digestive problems. Presence of such heavy metals and metalloids can degrade water quality as seen from the recent example of Minamata disease in Japan which is caused due to mercury poisoning (Stankovic and Stankovic 2013; Amasawa et al. 2016).

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Hazardous Effects on Plants

Water containing heavy metals and metalloids first contaminates the soil and then indirectly affects the plants and their growth. Based on the accumulation efficiency and the accumulation level of different metals/metalloids in soil, plants are characterized into three categories: hyperaccumulators, non-hyperaccumulators, and transgenic plants (Mleczek et al. 2013). Arsenic does not affect plant growth at low concentration because it does not take part in metabolic activities (Hughes et al. 2011). The symptoms of arsenic toxicity in plants are growth reduction in roots and poor seed germination (Garg and Singla 2011). This may lead to rapid disruption of plasma membrane structure of plants (Smith et al. 2010). At higher concentrations, arsenic may affect the plants abruptly which may even cause death of plants (Tripathi et al. 2007). Antimony creates several hazardous effects on plants, such as slower growth, slower synthesis rate of metabolites, and suppression of chlorophyll synthesis and enzymatic activities (An and Kim 2009). While beryllium causes germination and chromosomal aberration (Blais et al. 2008; Gordon and Bowser 2003), copper causes oxidative stress, disruption of cellular function, and chlorosis (Chibuike and Obiora 2014; Salem et al. 2000), and nickel causes inhibition of enzyme activities, growth, decrease in chlorophyll content, reduction in nutrient uptake, and oxidative stress in plants (Fashola et al. 2016; Chibuike and Obiora 2014; Salem et al. 2000). Similarly, selenium is responsible for slow growth rate of plants by reducing its biomass and also alteration of protein properties (Germ et al. 2007). Silver also inhibits plant growth and cell transduction, decreases chlorophyll content, affects homeostasis, and causes plant cell lysis (Prabhu and Poulose 2012; Qian et al. 2013). Zinc affects photosynthesis by reducing chlorophyll content and decreases growth rate, germination rate, and plant biomass (Gumpu et al. 2015).

7.3.3

Hazardous Effects on Aquatic Environment

Heavy metals and metalloids have been found in various aquatic species beyond the permissible amount. Heavy metals like mercury have been detected in watercourses as well as in variety of fishes at a level above the standards established by the United States Public Health Service. Aquatic environmental pollution caused by lead, used as a gasoline additive, is also creating great environmental concern (Dean et al. 1972). Moreover, heavy metals like chromium are also found accumulating in aquatic species. This results into an increase in chromium level in fishes and other aquatic species, and hence, eating of them poses threat to human health. Therefore, understanding the risk of health hazards and the detrimental effects on the environment owing to metal and metalloid contamination, it is highly important to design and implement effective pollution control and prevention strategies (Ermolaeva et al. 2014). It is also very important to invent new technologies and design simple processes for the remediation of these toxic heavy metals and

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metalloids of contaminated water. In the next section, we will discuss various nanotechnologies for the remediation of heavy metals and metalloids.

7.4

Nanotechnologies of Remediating Heavy Metals and Metalloids

Remediation methods for water contamination based on various nanotechnologies are discussed in this section. The remediation methods include chemical precipitation, adsorption, cementation, electrodeposition, solvent extraction, ultrafiltration, ion exchange, membrane separation, electrocoagulation, electrodialysis, bioremediation, phytoremediation, etc. (see Fig. 7.3). Remediation of heavy metal- and metalloidcontaminated groundwater can be accomplished by (a) removal of the contaminants from the subsurface and (b) in situ treatment where the contaminant metals are left in the subsurface and reduce their mobility or concentration to the levels considered safe for human health as well as the environment.

Fig. 7.3 Heavy metal and metalloid remediation technologies. (Lewis 2017; Lee et al. 2004; Chen 2004; Kulshreshtha et al. 2014; Shawai et al. 2017). The schematic shows various nanotechnologies for the remediation of heavy metals and metalloids in contaminated water

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Conventional Nanotechnologies of Remediation

Chemical Precipitation Chemical precipitation is one of the most widely used remediation techniques used for the metal and metalloid removal from wastewater. It has the advantage like its effectiveness, ease of operation, and cost-effective nature in extracting heavy metals from wastewater (Lewis 2017). This method is used for removing metallic cations as well as anions like fluoride, cyanide, and phosphate, as well as organic molecules. In precipitation, soluble heavy metals are converted to insoluble metals which settle down and/or are filtered out of the water. Afterward a solid separation operation, such as coagulation and/or sedimentation or filtration is followed (Yadav et al. 2019). The conventional precipitation processes follow two types of precipitation: hydroxide precipitation and sulfide precipitation (Zueva 2018). Between these two types, hydroxide precipitation is more commonly followed as compared to sulfide precipitation because it is easy to operate and has low cost and control of pH. Each of the soluble metals has a definite pH value at which shows the optimum hydroxide precipitation. All the metal hydroxides have solubility within the pH range of 8–11. Several different hydroxides are used as precipitating agents in chemical precipitation. Lime is mostly used in the industry for wastewater treatment because of its low cost and easy handling. Chemical precipitation resulted in complete removal of most heavy metals like zinc, copper, manganese, iron, nickel, and cobalt as hydroxides, whereas it leads to incomplete removal of cadmium, lead, and mercury. In case of wastewater containing chromium, the solution should be first reduced by ferrous sulfate, sulfur dioxide, or metallic iron before treatment with lime for precipitation. For the removal of complex organic metallic compounds, breakpoint chlorination is required before chemical precipitation (Rodriguez et al. 2012). It is recently found that Ca(OH)2 and NaOH are used for the removal of Cu(II) and Cr(VI) ions from wastewater (Mirbagheri and Hosseini 2005). Here, first ferrous sulfate is used to reduce Cr(VI) to Cr(III), and then Ca(OH)2 was added resulting in the maximum precipitation of Cr(III) at pH level of 8.7. The result shows that the concentration of chromate was decreased from 30 to 0.01 mg L 1 due to this treatment. For Cu removal, aeration is used to reduce cuproammonia followed by the addition of Ca (OH)2 and NaOH which resulted in maximum precipitation at pH 12. Copper concentration was found to be reduced from 48.51 to 0.694 mg L 1. Lime precipitation was increased with the help of fly ash seed material (Chen et al. 2009). It is found that removal of heavy metal from wastewater is increased if coagulants like alum, iron salts, and organic polymers are also added (Charerntanyarak 1999). However, hydroxide precipitation has also some disadvantages like generation of large amount of low-density sludge, selection of an ideal pH for amphoteric metal hydroxides, and inhibition of hydroxide precipitation in case of presence of any complexing agents in the wastewater (Zueva 2018). Sulfide precipitation is also used effectively for the treatment of wastewater containing toxic heavy metal ions. It has several advantages, such as metal sulfide

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precipitates have lower solubilities as compared to hydroxide precipitates. Moreover, sulfide precipitates are not amphoteric; therefore, it is possible to remove large amount of metals over a wider range of pH as compared to hydroxide precipitation. Metal sulfide sludge produced during sulfide precipitation shows better dewatering and thickening characteristics in comparison to corresponding sludge of metal hydroxides (Köhler et al. 2007). Another sulfide precipitation process has been reported recently which use sulfate-reducing bacteria (SRB) to oxidize simple organic compounds under anaerobic conditions and transform sulfates into hydrogen sulfide (H2S). Afterward, H2S reacts with divalent metal ions to form insoluble metal sulfides (Köhler et al. 2007). Kousi et al. developed an upflow fixed-bed sodium dodecyl sulfate to monitor the treatment of wastewater contaminated with zinc. The reactor showed the capacity to completely reduce sulfates for initial concentrations up to 7200 mg L 1, remove soluble zinc for initial concentrations up to 400 mg L 1, and also remove completely the total organic carbon (TOC) from initial concentrations up to 1500 mg L 1 (Kousi et al. 2007). However, sulfide precipitation has serious disadvantages, such as it may cause evolution of toxic H2S fumes if present under acidic medium. Therefore, it is important to keep the condition neutral or basic while performing this process. Moreover, metal sulfide precipitation process may form colloidal precipitates which further hinder the separation process during their settling or filtration. Heavy metal chelating precipitation is used as an alternative treatment process by many companies in which heavy metals are precipitated from aqueous systems using chelating agents. Initially, three commercial heavy metal precipitants, trimercaptotriazine, potassium/sodium thiocarbonate, and sodium dimethyldithiocarbamate, were evaluated (Matlock et al. 2002a, b). However, due to the presence of insufficient binding sites and environmental risks during their use, other chelating agents were explored. Later, a new thiol-based compound, 1,3-benzenediamidoethanethiol dianion (BDET2 ), was developed which can effectively precipitate heavy metals from acid mine drainage, and mercury in the leachate solution was synthesized (Matlock et al. 2002a, b). Recently, an organic heavy metal chelator dipropyl dithiophosphate was synthesized which could reduce the concentration of lead, cadmium, copper, and mercury from 200 mg L 1 at pH 3–6 up to over 99.9% (Ying and Fang 2006). Potassium ethyl xanthate was also used to remove copper ions from wastewater over a broad range of copper concentration (50, 100, 500, and 1000 mg L 1) to as low as 3 mg L 1 (Chang et al. 2002).

Adsorption Adsorption is an effective and economic method for remediation of heavy metals from wastewater. In this process a substance is transferred from the liquid or gaseous phase to the surface of a solid and bounded by physical and/or chemical interactions. Adsorption method has many advantages like low cost, high efficiency, ease of operation, regenerative, metal selectivity, profitability, no toxic sludge generation, and effective (Keng et al. 2014). Adsorbents, such as activated carbon, carbon

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nanotubes, zeolite, chitosan, and clay, have been studied and developed to remove toxic heavy metals from wastewater and soil (Strelko et al. 2004; Kosobucki et al. 2008). The conventional adsorption processes mentioned in literature are removal of Pb(II), Cd(II), Ni(II), and Cu(II) using natural kaolinite clay (Jiang et al. 2010); removal of Ni(II), Pb(II), and Cu(II) using activated carbon (Kadirvelu et al. 2000); removal of hazardous heavy metals from aqueous environment using low-cost adsorption materials (Keng et al. 2014); and immobilization of selected heavy metals in sewage sludge by natural zeolites (Kosobucki et al. 2008). Adsorbent based on materials of natural origin like zeolites, clay, peat moss, and chitin is found to be an effective agent for removal of toxic heavy metals like Pb, Cd, Zn, Cu, Ni, Hg, Cr, etc. Chitosan has been used for treatment of Hg2+, Cu2+, Ni2+, Zn2+, Cr6+, Cd2+, and Pb2 + (Annadurai et al. 2003). These adsorbents showed great capability to remove toxic heavy metal substances from contaminated water (Tripathi and Ranjan 2015). Some other important natural adsorbents based on agricultural wastes (known as biosorption; we have discussed biosorption separately in bioremediation section) are rice husk, black gram, neem bark, Turkish coffee, waste tea, walnut shell, etc. Similarly, some industrial by-products are also used as low-cost adsorbents like coffee husks, blast furnace sludge, battery industry waste, iron(III) hydroxide waste slurry, fly ash, lignin, sea nodule residue, red mud, tea factory waste, sugar beet pulp, and grape stalk wastes.

Cementation Cementation is an effective and economic hydrometallurgical process used as a means to recover toxic and/or valuable metals from solutions as well as for the purification process of stream and wastewaters in industry (El-Batouti 2005). Cementation is a process in which a metal is precipitated from its salt solution to its elemental state by another electropositive metal via a spontaneous electrochemical reduction. This is associated with the consequent oxidation of a sacrificial metal to recover more expensive and noble metal present in aqueous solutions (Nassef and El-Taweel 2015). Cementation method has some advantages like low cost, low energy consumption, and recovery of metals in relatively pure metallic form, whereas the main disadvantage is that excess sacrificial metal is consumed in this process. In practice a considerable spread in the electromotive force between metals is necessary to ensure adequate cementation capability. Zinc dust, as a precipitant used for cementation process, has the potential to recover several metals like cadmium, mercury, and lead from industrial wastes even when present in small amounts (Dean et al. 1972).

Solvent–Solvent Extraction Solvent–solvent extraction is a suitable method for the removal of heavy metals from the wastewaters of chemical and electronic industries. In this treatment metals are

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selectively separated from aqueous solution using extractants which are mainly organic compounds with molecular mass of 200–450 gm/mole and almost insoluble in water (5–50 ppm). After that, the metals are converted into a form which becomes soluble in that organic compound when they are mixed together (Černá 1995). This method is generally used in chemical and metallurgical industries for extracting particular heavy metals. For example, kerosene was used as an extractant for the removal of Cd, Cu, Pb, and Zn from wastewater, and it was used as aid to the flotation of hydrophobic compounds in flotation process (Cauwenberg et al. 1998). First, kerosene is nicely mixed with wastewater, and then two phases were separated. Then the organic fraction was treated separately to increase the concentration of the heavy metal followed by its removal. This method is also successfully employed in uranium and copper industries, and process is conducted by using a number of different types of mixer-settlers, and counter current flow type contactors, where feed and solvent enter preferably at opposite ends of the system (Dermont et al. 2010).

Ion Exchange Ion-exchange process is widely used to remove heavy metals from wastewater due to several advantages, such as high treatment capacity, high removal efficiency, and fast kinetics (Fu and Wang 2011). In this process, reversible chemical reactions which remove dissolved ions from solution and replace them with other similarly charged ions are involved. Ion-exchange resin is a solid resin of either synthetic or natural type and can exchange its cations with the heavy metals present in the contaminated water. In water treatment, it is primarily used for softening where calcium and magnesium ions are removed from water; however, it is also used for the removal of other dissolved heavy metallic ions (Kapoor and Viraraghavan 1995; Tunsu and Wickman 2018; Chen 2004; Dabrowski et al. 2004). During this process, few parameters, such as pH, temperature, initial metal concentration, and contact time, need to be taken care of, and they affect the uptake of heavy metal ions by ion-exchange resins. The cation-exchange resins which are strongly acidic in nature contain sulfonic acid groups ( SO3H), and those which are weakly acidic contain carboxylic acid groups ( COOH); however, the hydrogen ions present in either sulfonic or carboxylic group of the resin also act as exchangeable ions with heavy metal cations (Fu and Wang 2011).

7.4.2

Membrane Filtration

Membrane filtration technology is very effective for heavy metal and metalloid removal from wastewater that contains higher concentration of pollutants. This technology has the advantages, such as high efficiency, simple operation, and less space constraints. Membrane filtration method consists of semipermeable

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membranes which act as molecular sieves permitting substances that are soluble and have sizes smaller than the pores of the membrane while retaining the larger-size molecules. This technique is a function of membrane pore size and molecular diameter of the substance, and its rate depends on the pressure applied on it (Lee et al. 2004). Membrane filtration includes ultrafiltration, reverse osmosis, nanofiltration, and electrodialysis techniques. Ultrafiltration is a membrane technique which uses hydrostatic pressure across the semipermeable membrane to force a liquid for the removal of dissolved and colloidal substances. The dissolved metal ions, which are present either in the form of hydrated ions or low molecular weight complexes, would easily pass through the ultrafiltration membranes as the pore sizes of the membrane are larger than the dissolved metal ions. Two types of ultrafiltration processes, micellar-enhanced ultrafiltration and polymer-enhanced ultrafiltration, were used to increase the removal efficiency of metal ions (Fu and Wang 2011). The efficiency of micellarenhanced ultrafiltration depends on the nature and concentrations of the metals, pH of the solutions, surfactants, ionic strength, and other operational parameters. Heavy metals like zinc have been removed from synthetic wastewater by micellar-enhanced ultrafiltration using sodium dodecyl sulfate (Landaburu-Aguirre et al. 2009). Similarly, Zn2+, Cu2+, Ni2+, Pb2+, and Cd2+ were also removed from synthetic wastewater using two types of anionic surfactants, such as sodium dodecyl sulfate and linear alkylbenzene sulfonate, in a lab-scale membrane system (Samper et al. 2009). Reverse osmosis (RO) is a process in which the fluid, which is being purified, passes through a semipermeable membrane in the direction opposite to that of natural osmosis when the hydrostatic pressure becomes greater than the osmotic pressure. Reverse osmosis has the capacity to remove larger amount of dissolved metal impurities. In the world, 20% of desalination is done by reverse osmosis process (Shahalam et al. 2002). Now these days it is a very popular and effective wastewater treatment scheme. Nanofiltration is the intermediation process between ultrafiltration and reverse osmosis. Nanofiltration has advantages like easy to operate, reliable, and energy efficient as well as efficient higher pollutant removal (Eriksson 1988). Nanofiltration is used for the remediation of heavy metal ions, such as nickel (Murthy and Chaudhari 2008), copper (Cséfalvay et al. 2009), chromium (Muthukrishnan and Guha 2008), and arsenic (Nguyen et al. 2009) from wastewater. Electrodialysis is another membrane process used for the separation of metal ions across charged membranes from one solution to another under an electric field acting as driving force (Fu and Wang 2011). In most electrodialysis processes, either cation-exchange or anion-exchange membranes are used. This is a continuous separation process which separates heavy metals like Zn2+, Cu2+, Pb2+, and Cd2+ from low concentration solutions by electropermutation combining ion-exchange resins and membranes. The separation efficiency of electrodialysis is based on ionic affinity (Smara et al. 2007).

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Electrochemical Treatment

Electrochemical methods are used for removal of metals and metalloids by electroplating (deposition of metal ions on electrode) on a cathode surface which is recovered again in their elemental metal state. This technology for wastewater treatment is relatively expensive and requires high electricity supply and thus not very widely applied (Fu and Wang 2011). Electrochemical treatments include four different technologies: electrodeposition, electrocoagulation, electroflotation, and electrooxidation. Out of these four processes, electrodeposition is used for metal recovery, electrocoagulation is employed for water treatment, electroflotation is used for organic pollutants recovery, while electrooxidation is used for wastewater treatment and degradation (Chen 2004). Removal of Cu2+, Cr6+, and Ni2+ is done with satisfactory results by electrochemical method (Hunsom et al. 2005).

7.4.4

Bioremediation Technologies of Heavy Metals and Metalloids

Bioremediation is a natural process which requires less energy as compared to other technologies. It is useful for the complete destruction of a wide variety of contaminants and can transform hazardous substances to harmless products. Moreover, it does not transfer contaminants from one environment medium to another, e.g., from land to water or air, but rather provides possible complete destruction of the target pollutants (Kulshreshtha et al. 2014). Biological methods for the removal and detoxification of heavy metals and metalloids from polluted water and sediments include several processes, such as volatilization, biosorption, leaching, precipitation, phytoremediation, oxidation/reduction, degradation, and bioaccumulation. In addition, recovery of heavy metals and metalloids is achieved via biosorption using microbial and plant biomass (Kikuchi and Tanaka 2012). Biosorption is followed by using nonliving biomass, their constituents, and metabolites to decontaminate heavy metals through physicochemical adsorption. According to the principle of biosorption, metal ions are interacted by various forces such as van der Waals forces, complexation, ion-exchange reaction, electrostatic interactions, surface precipitation, or their combinations, and these forces result in the adsorption process (Veglio and Beolchini 1997). Metal-binding capacity by different types of biomasses is reviewed for fungi (Ahluwalia and Goyal 2007; Volesky and Holan 1995; Miao et al. 2006), bacteria (Ahluwalia and Goyal 2007; Volesky and Holan 1995), algae (Davis et al. 2003; Mehta and Gaur 2005), and plant-derived materials (Babel and Kurniawan 2003). Bioaccumulation method uses living microorganisms to absorb and remove heavy metals and metalloids taped in intracellular space. In bioaccumulation, the heavy metals accumulated into the microbial cells are precipitated or bound onto intracellular structure (Malik 2004). Biomasses are sometimes pre-treated by some

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alkaline medium for the successful removal or decontamination of heavy metals. For example, living Aspergillus niger biomass is pre-treated with 0.1 M NaOH for removal of nickel. Similarly, marine green alga Chlorella sp. NKG16014 gives 48.7% of Cd removal efficiency. It was quantitatively evaluated by measuring the amount of cell adsorption and intracellular accumulation (Matsunaga et al. 1999). Oxidation/reduction reactions are used by microorganism for detoxification and removal of metal ions. For example, Fe(II) and Mn(II) are removed by Fe(II)- and Mn(II)-oxidizing bacteria following oxidation and precipitation processes by these bacteria. Mn(II)-oxidizing bacteria and fungi react with manganese and produce manganese oxides which have high desorption capacity for heavy metal cations, such as Cd(II), Cu(II), Co(II), Ni(II), Pb(II), and Zn(II). These bacterial and fungal generated manganese oxides are also capable of oxidizing various inorganic compounds like Cr(III) and As(III) (Miyata et al. 2007). Arsenic is also oxidized by this manganese oxide from its As(III) oxidation state to As(V) (Saito et al. 2004). Leaching technique involves biological processes and products for the removal of heavy metals and metalloids from sediments. Heavy metals in dredged contaminated sediments are removed via solubilization followed by microbial processes such as bioleaching and heterotrophic leaching (Gadd 2000). Washing and extraction of heavy metals in the sediments using biosurfactants are also done (Mulligan 2005). Bioleaching of mineral ores in bio-based industry is done by acidophilic iron- and sulfur-oxidizing bacteria, Acidithiobacillus thiooxidans and Acidithiobacillus ferrooxidans, respectively (Gadd 2004). By precipitation method, heavy metals and metalloids from aqueous phase are precipitated and removed using biological substances. Heavy metals from wastewater and leachates are removed by sulfide precipitation using sulfate-reducing bacteria in which precipitation is done by using alkalinity generated cyanobacteria and algae. Cyanobacteria can reduce sulfate into sulfide (Mehta and Gaur 2005). Leachate containing heavy metals was treated with mixed culture of sodium dodecyl sulfate using an anaerobic reactor containing dimethyl disulfide which is responsible for the precipitation of heavy metals (Essa et al. 2006). Similarly, volatilization and degradation methods also used several microorganisms for remediation of heavy metals and metalloids from wastewater. Selenium is removed by using Stenotrophomonas maltophilia which is a Gram-negative bacterium and isolated from seleniferous agricultural evaporation. This bacterium can actively produce volatile alkylselenides, such as dimethyl selenide, dimethyl selenenyl sulfide, and dimethyl diselenide in the presence of selenate (SeO42 ) and selenite (SeO32 ) (Dungan et al. 2003). Heavy metals may inhibit biodegradation of chlorinated organics by taking part in enzymatic metabolism. Metals may be present in many forms in environment. Recent advancement made use of metalresistant bacteria (cell and gene bioaugmentation), treatment amendments, clay minerals, and chelating agents to reduce bioavailable heavy metal concentrations. Phytoremediation is an alternative way of remediation of heavy metals as discussed in the next section in details (Olaniran et al. 2013).

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Phytoremediation Technologies of Heavy Metals and Metalloids

Phytoremediation is a form of bioremediation. It consists of the Greek word “phyto” which refers to plant and the Latin suffix “remedium” which means curing or restoring. In phytoremediation techniques, plants and soil microbes are used to decrease the concentration levels or the toxic effects of heavy metal and metalloid contamination in the environment (Shawai et al. 2017). Phytoremediation techniques are very effective and prosperous way to remediate heavy metals from water, soil, and air as compared to other conventional, physical, and chemical treatment methods. Suitable plant species are the key factor to achieve better remediation performance. Phytoremediation techniques seem to be the most efficient and eco-friendly, but it has some limitations (Ali et al. 2013a, b). There are two specific factors, bioadsorption factor and intracellular accumulation factor, which differentiate the mechanisms of the removal of metal ions by phytofiltration and phycoremediation.

Phytoextraction/Phytoaccumulation Phytoextraction is a very useful phytoremediation technique for removal of heavy metals and metalloids, from polluted oils, sediments, or water. In this process, plants uptake the heavy metals via metabolically active mechanisms and accumulate the heavy metals in their roots, leaf, fruits, and other plant body parts at intercellular level. Later, these heavy metals are converted to less toxic substances or another form than the original. The selection of plant species for phytoextraction is based on many factors, such as their potential to evapotranspiration groundwater, the degradative enzyme they produce, their growth rates and yield, the depth of their root zone, and their ability to bioaccumulate contaminant (Shawai et al. 2017). The hyperaccumulator plants, which produce comparatively less above ground biomass but accumulate significant extent of target heavy metals, are used. Apart from them, other plants, such as Brassica juncea (Indian mustard), are also useful for phytoextraction (Shawai et al. 2017). This process is novel, efficient, cost-effective, applicable in situ, solar energy driven, and environmentally and eco-friendly. Moreover, by this technique, the contaminants can be collected from the media and turned into an easily extractable form (plant tissues) (Prasad 2003).

Phytofiltration/Rhizofiltration Phytofiltration is used to remediate heavy metals present in low concentration in extracted groundwater, surface water, and wastewater. This technique involves either adsorption or precipitation of contaminants onto plant roots or absorption of contaminants surrounding the root zone. Phytofiltration can be used for the removal

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of Pb, Cd, Cu, Ni, Zn, and Cr, which are primarily retained within the roots (Shawai et al. 2017). In this process, both terrestrial and aquatic plants (submerged or floating) are used to remediate pollutants from solution mainly through their roots, and in some cases, fronds are also directly involved (Olguín and Sánchez-Galván 2012). Metabolic activities are involved for the accumulation and degradation of such contaminants by phytofiltration. Sunflower, Indian mustard, tobacco, spinach, and corn are generally used to remove lead from water, with sunflower having the greatest ability (Prasad 2003).

7.5

Nanomaterials for Heavy Metal and Metalloid Remediation

We have discussed about several nanotechnologies for remediation of heavy metals and metalloids from wastewater in this chapter. Although some treatment schemes give better accumulation and efficiency of remediation, these treatments may also lead to the generation of some toxic by-products besides their high cost of processing (Mleczek et al. 2013). Among these treatments, adsorption has been considered as the most effective, comparatively cost-effective, and commonly used method for remediation of heavy metals and metalloids from wastewater (Reddad et al. 2002). In this section, we are discussing the nanomaterials used as effective adsorbents for adsorption remediation technology of heavy metals and metalloids.

7.5.1

Metal–Organic Frameworks as Adsorbents

Metal–organic frameworks (denoted as MOFs) are defined as a type of threedimensional network porous materials and generated through organic ligands and metal ions to form a metal–ligand complex by self-assembly. Metal–organic frameworks which have been developed in recent years are considered as great adsorbents due to their unique physicochemical performance. The metal–organic framework particles commonly used as adsorbents are mentioned in Table 7.3. Metal–organic frameworks have larger porosity and higher specific surface area. The pore structure can further be efficiently designed, and the structural modifications can easily be done. In every step of the process, such as adsorption, separation, storage, and transportation, metal–organic frameworks show a better handling and operation performance than conventional adsorption materials such as active carbon (Fang et al. 2010). Adsorption of heavy metals ions on metal–organic frameworks take place via several mechanisms, such as electrostatic adsorption, π–π bonding interactions, hydrogen bonding, acid–base adsorption, and adsorption at open metal sites (Zhou and Kitagawa 2014). The removal efficiency for heavy metal ions by metal– organic frameworks can reach up to 99%. The maximum adsorption capacities of Cd2

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Table 7.3 Metal–organic frameworks particles used as adsorbents for heavy metal and metalloid remediation Metal–organic framework UIO-66 MOF-74-Zn ZIF-8@CA MIL-68 (W) CU-MOFs HKUST UIO-66

Preparation methods Solvothermal method Solvothermal method Solution method Microwave synthesis Solvothermal method Solvothermal method Liquid-phase impregnation post-synthetic modification (LP-PSM)

Contaminants Pb(II) Hg(II) Cr(IV) Cd(II) Pb(II) Cs(I) Co(II)

Adsorption capacity (mg/g) 205 63 41.8 0.139 219 153 256

Modified after Yin et al. (2018), Shi et al. (2018), Naeimi and Faghihian (2017), Yuan et al. (2017), and Bo et al. (2018) +

, Cr3+, Pb2+, and Hg2+ were reported at 49 mg/g, 117 mg/g, 232 mg/g, and 769 mg/g, respectively (Merrikhpour and Mahdavi 2017).

7.5.2

Natural Clay Material as Adsorbent

Natural clay is also used as an effective adsorbent for the removal of heavy metals from its solution. Clays as adsorbent categorized are into three groups: kaolinite, montmorillonite (smectite), and mica. Out of them, montmorillonite is most effective and cheapest as compared to activated carbon (Sharma and Forster 1993; Tripathi and Ranjan 2015). Removal of heavy metals via adsorption on clay material from El Hammam district (Alexandria) using high silica content ~53 wt% has been examined. The clay material has been tested and characterized by various techniques, such as crystalline structure by X-ray diffraction, constituents of the clay material by X-ray fluorescence, and thermal analysis for any change in structure causing from thermal treatment. To find out the adsorption efficiency of clay material, FeCl2, Ni (NO3)2.6H2O, and ZnCl2 solutions have been used as metal model compounds. Removal of heavy metals like cadmium, lead, nickel, chromium, iron, zinc, and copper by using clay materials is targeted due to their increasing contamination in wastewater streams (Missana and Garci 2007). In this process, the bentonite clay material is first grounded, sieved to mesh size Zn > Ni. Adsorption of heavy metals and metalloids by clay and clay composites is accomplished with series of complicated adsorption mechanisms, such as surface complexation, ion exchange, and metal binding of heavy

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metal cations directly on the clay surfaces, and montmorillonite clay is also used as an adsorbent for this process (Sdiri et al. 2012).

7.5.3

Activated Carbon as Adsorbent

Activated carbon can be formed from any carbonaceous materials of animal, plant, or mineral origin having high concentration of carbon using chemical or gas activation methods. Wood, charcoal, bituminous coals, lignite, peat, bone, fruit pits, nut shells, and synthetic polymers are used as raw materials for the manufacturing of activated carbon. Heavy metal removal by charcoal adsorption gives very effective and economic and industrially feasible solution. Nowadays, activated carbon is further modified by chemical treatments and incorporated into different nanocomposites (Nagajyoti et al. 2010). The modified activated charcoal gives excellent decontamination efficiency. Cadmium shows the highest removal via adsorption on activated charcoal (Jiang et al. 2010; Kadirvelu et al. 2000).

7.5.4

Low-Cost Adsorbents

Remediation of heavy metals like Pb, Cd, Zn, Cu, Ni, Hg, and Cr contamination is also done by using various low-cost adsorbents. The low-cost adsorbents, such as zeolites, clay, peat moss, and chitin, and various agricultural wastes like neem bark, rice husk, waste tea, black gram, Turkish coffee, and walnut shell have better potential to remediate heavy metals. Various industrial by-products used as low-cost adsorbents are fly ash, tea factory waste, waste slurry, lignin, blast furnace sludge, battery industry waste, iron(III) hydroxide, sea nodule residue, red mud, coffee husks, sugar beet pulp, areca waste, and grape stalk wastes (Babel and Kurniawan 2003; Siti et al. 2013). Montmorillonite clay is the most effective and cheapest low-cost adsorbent as compared to activated carbon. For better efficiency clay is modified as clay–polymer composites (Tripathi and Ranjan 2015), or improved by modifying them with peat mass which leads to larger surface area (200 m2/g) and high porosity and higher adsorption efficiency in removing Cu2+, Cd2+, Zn2+, and Ni2+ (Gosset et al. 1986). Chitosan is also a natural material obtained by alkaline N-deacetylation of chitin. Chitosan is very interesting as low-cost adsorbent because of its chelating property for heavy metal removal. Chitosan is used for treatment of heavy metals such as Hg2+, Cu2+, Ni2+, Zn2+, Cr6+, Cd2+, and Pb2+ (Ahmad et al. 2015).

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Nano-sized Carbonized Waste Biomass as Adsorbents

In this process, agricultural waste biomass is carbonized at nano-sized particles which are used as adsorbents in water remediation technology. The lignocellulose and nitrogenous agricultural waste biomass generally known as Dalbergia sissoo pods were used for the remediation of heavy metals like Cd(II), Pb(II), and Ni(II). Dalbergia sissoo pods contain high amount of proteins and crude fibers and have nitrogenous substances of various functional moieties for sequestering heavy metal ions in high concentration. Mechanisms involved for metal ion removal by these adsorbents include ionic adsorption by lignocellulosic biomass or reduction of metals by lignocellulosic biomass due to presence of functional groups, such as amines, phenolic, methoxyl, and hydroxyl. The maximum adsorption efficiency obtained by Dalbergia sissoo pods for Ni(II), Cd(II), and Pb(II) is reported as 95%, 95%, and 98% at different pH values, such as pH 6 for Ni(II) and pH 5 for Cd(II) and Pb(II) (Mahajan and Sud 2015).

7.5.6

Cellulose-Based Materials as Adsorbents

Cellulose-based materials are significantly used for remediation of heavy metals and metalloids in wastewater. Bagasse is a natural material containing 50% cellulose (Fu and Wang 2011) and used to adsorb Cu(III) and Cu(VI) ions and Hg both in native and immobilized forms. It can also be used to produce activated carbon for the adsorption of Pb(II) (Basu et al. 2015). Wood sawdust obtained from wood processing industry is used for adsorption of heavy metal ions like Cd(II) ions due to its cellulosic composition (45–50%). Other cellulose-based materials, such as eggshell, coconut tree, and sugarcane bagasse sawdust, are used for Cu, Pb, and Zn removal. Similarly, wood bark is used for arsenic, mercury, cadmium, copper, lead, chromium, zinc, iron, and nickel removal; modified banana peels can be used for Mn (II) removal (Ali 2017); peanut shell is used for cadmium, copper, lead, and zinc adsorption (Jaishankar et al. 2014a, b); rice hull and bran are used for Cu(II), Zn(II), Ni(II), C(II), and Cr(III) (Teixeira and Zezzi 2004). Cotton cellulose is used to remove boron in desalination plants. Cellulose is further chemically modified to form many of its useful derivatives which are used for wastewater remediation widely (Liu et al. 2002). Cellulose is commonly used as powder form, fine particles, or fiber form (Volesky and Holan 1995). Modification is done by etherification, esterification, oxidation, halogenations, and grafting and modified using mineral and organic bases and acids, oxidizing agents, and organic/inorganic compounds. Cellulose derivatives: Cellulose nanocrystals are a derivative of cellulose which can be derived from many cellulose-based sources, e.g., rice husk and straw, for the adsorption of Cd(II), Al(III), and Na(I) (Albernaz et al. 2015). Cellulose nanocrystals were also obtained from cotton linters after hydrolysis (Yu et al. 2013). These

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cellulose nanocrystals were further chemically modified to increase their efficiency for the adsorption of Pb(II) and Cd(II) from aqueous solutions. Another derivative of cellulose known as carboxymethyl cellulose is used for heavy metal remediation, and it remediates heavy metals via complexation with metal ions (Barakat and Schmidt 2010). Another derivative of cellulose is cellulose nanofibers which are used for the adsorption of Cd(II). Graft copolymers made from cellulose biopolymers are also used as adsorbents for the removal of Pb2+, Zn2+, Cu2+, and Cd2+ (Shawai et al. 2017). Cellulose-based hydrogels have the cross-linked three-dimensional network structure which are used for heavy metal ion adsorption widely nowadays. For example, carboxymethyl cellulose/chitosan hydrogels were used as renewable and biocompatible adsorbent for Fe(III) where the presence of chitosan increases the adsorption efficiency of carboxymethyl cellulose/chitosan hydrogel (Liu et al. 2012). Similarly, carboxymethyl cellulose cross-linked with epichlorohydrin can be converted to hydrogel beads, and these beads are used to remove Pb(II), Ni(II), and Cu(II) (Wang and Wang 2016). Cellulose-based composite materials find many applications in membrane technology for water purification (Thakur and Voicu 2016). For example, cellulose/Na– montmorillonite biocomposite is used to remove Cr(VI) (Kumar et al. 2011). Similarly, cellulose/chitosan biocomposite is used as an adsorbent for the removal of Cu(II), Zn(II), Cr(VI), Ni(II), and Pb(II) (Wu et al. 2010), and cellulose/TiO2 nanocomposite showed significant adsorption capacity for Pb(II). A composite of cellulose acetate/sulfonated polyetherimide (SPEI) showed the high adsorption capacity for heavy metal ions in the order of Cd(II) > Zn(II) > Ni(II) > Cu (II) (Nagendran et al. 2008; Barakat 2011).

7.6

Conclusion and Future Trends

Hazardous effects due to heavy metal- and metalloid-contaminated water is one of the growing environmental problems throughout the world. A variety of heavy metal and metalloid remediation technologies, such as adsorption, chemical precipitation, ion exchange, membrane filtration, electrochemical treatment, and bioremediation, have been developed to remove such hazardous substances from wastewater in order to meet the environmental regulations. Recent data show the adverse health effects caused due to exposure of various heavy metal and metalloid contamination in water, such as renal tubular damage, osteomalacia, hypertension due to exposure to cadmium, anemia, encephalopathy, nephropathy caused due to exposure to lead, high risk of lung and skin cancer, diarrhea due to exposure to arsenic in water, tendency to face encephalopathy, arrhythmia, diabetes due to exposure to nickel, etc. In this chapter we have discussed the heavy metals and metalloids which contaminate water; their adverse effects on human health, plants, and aquatic environments; and several technologies based on nanomaterials for the remediation of such heavy metals and metalloids from contaminated water. The literature survey shows that ion

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exchange, adsorption, and membrane filtration are the most frequently used technologies for the treatment of heavy metal-contaminated wastewater. Though ion exchange, solvent–solvent extraction, and membrane filtration processes have been widely used to remove metals and metalloids from wastewater, adsorption technologies by low-cost adsorbents and biosorbents are found to be the most effective and economic method for removing low concentration heavy metals from wastewater. Bioremediation, being a natural and energy-efficient process as compared to other technologies, is useful for the complete destruction of a wide variety of contaminants, and it can transform hazardous substances to harmless products. Bioremediation is not only an effective process for the complete degradation of pollutants from air, soil, water, and raw materials from industrial waste, but it does not transfer contaminants from one environment medium to another. Therefore, bio-based nanomaterials, bioadsorbents, and microorganism-based nanotechnologies play an important role for cost-effective, energy-efficient, environmentally friendly remediation of heavy metals and metalloids in contaminated water. Hence, for sustainable future applications, there is a considerable scope of research in this area. Acknowledgments Paramita Das gratefully acknowledges the infrastructural support from Indian Institute of Science Education and Research (IISER), Bhopal.

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Chapter 8

Hybrid Treatment Technologies for the Treatment of Industrial Wastewater Vikas S. Hakke, Murali Mohan Seepana, Shirish H. Sonawane, Anand Kishore Kola, and Ramsagar Vooradi

Contents 8.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.2 Different Types of Pollutants from Pharmaceutical and Textile Industries . . . . . . . . . . . . . . 8.2.1 Pollutants from Pharmaceutical Industries . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.2.2 Pollutants from Textile Industries . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.3 Conventional Methodologies for the Treatment of Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . 8.4 Application of Nanomaterials as Photocatalysts for Wastewater Treatment . . . . . . . . . . . . . 8.5 Hybrid Wastewater Treatment Techniques . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.5.1 Effect of Wastewater Properties . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.5.2 Hybrid Wastewater Treatment Techniques Using Hydrodynamic Cavitation . . . . . 8.5.3 Ultrasound Cavitation and Applications in Wastewater Treatment . . . . . . . . . . . . . . . . 8.5.4 Hydrodynamic Cavitation and Photocatalysis with Photo-Fenton Process . . . . . . . . 8.5.5 Hybrid Treatment Techniques Using Photocatalysis and Nanofiltration . . . . . . . . . . 8.5.6 Hybrid Hydrogel-Based System for Removal of the Dyes Using Hydrodynamic Cavitation Process . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.6 Future Prospectus . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.7 Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract The twentieth century has witnessed a rapid industrialization of chemical processes. It has facilitated in improvising living style of humans but other side effluents from the industries adversely polluting the water. The industrial wastewater has to be treated to avoid the water pollution and water scarcity around the world. A single treatment technology that leads to zero water pollutant discharge from industry appears to be unrealistic; thus the combination of different technologies will be the new approach to deal efficiently with the present condition of wastewater treatment. In this chapter, such hybrid technologies are discussed with the different

V. S. Hakke · M. M. Seepana (*) · S. H. Sonawane · A. K. Kola · R. Vooradi Chemical Engineering Department, National Institute of Technology, Warangal, Telangana, India e-mail: [email protected]; [email protected]; [email protected]; [email protected] © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature Switzerland AG 2021 Inamuddin et al. (eds.), Water Pollution and Remediation: Heavy Metals, Environmental Chemistry for a Sustainable World 53, https://doi.org/10.1007/978-3-030-52421-0_8

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case studies. Effective use of hydrodynamic cavitation techniques in combination with the micro-/ultrafiltration, use of photocatalysis combined with ultrafiltration and use of novel adsorbent materials like hydrogels, shows improvement in the removal of pollutants from the industrial wastewater. This article also endorses the inclusion of energy and water reuse plan within the treatment scheme and accordingly proposes a conceptual industrial wastewater treatment system. Keywords Industrial wastewater · Hydrodynamic cavitation · Fenton process · Nanofiltration · Hybrid treatment technologies

8.1

Introduction

Increasing population, urbanization, and industrialization demand more and more fresh water, from which only 3% maximum is utilized for drinking purpose and the rest are all used in process. Starting with small applications like domestic cleaning to being used in process industries for different processes, water is needed everywhere. All utilized water thrown out of the system as wastewater need treatment for reuse; this is because of limited source of drinking water. Treatment of wastewater within the economical constraint is one of the latest challenges in the world. Scarcity of drinking water makes water more valuable day by day. 70% of world water is hard water giving addition to constraint of limited water sources. Wastewater released from the industry or domestic system is associated with many harmful and toxic chemicals and impurities. Cosmetic effluents coming from the domestic house units nowadays increase more and lead to harmful effect on the environment as well as human health. Conventional methods for the treatment of such new-generation pollutants are inefficient; thus we need some of the hydride technologies that can effectively work on separation of pollutants and disposal of pollutants in the safer way. Separation and disposal of pollutants, which consist of inorganic, organic, and bioactive entities via upgraded and new wastewater treatments, are the present challenges. Domestic wastewater consists of heavy metals (Moriyama et al. 1988), citronellol, hexyl cinnamic aldehyde, and menthol as well as some preservatives like citric acid and triclosan; xenobiotic organic compounds were reported (Eriksson et al. 2003). On the other hand, industrial effluents consist of many complex pollutants, metallic, dyes and toxic chemicals and all (Kumar et al. 2018). Expensive wastewater treatment with its time-consuming nature makes him several times ineffective and less considerable that causes harm to the environment because of direct release. Hybrid techniques such as advanced oxidation processes (Bhanvase et al. 2017; Shende et al. 2018), hydrodynamic cavitation (Gogate and Pandit 2005; Saharan et al. 2011), Fenton, and photo-Fenton (Lucas and Peres 2006; Papić et al. 2009) are reported for the wastewater treatment widely. Depending upon pollutant size, type, and source, these hybrid techniques are implemented; the study shows that advanced

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oxidation processes, hydrodynamic cavitation process, and chemical oxidation process are very much useful for the inorganic micropollutants like dyes coming out from the textile industry. Oxidation of pollutants will be achieved by synthesis of radicals in the process stream that can be possible with hydrodynamic cavitation. In recent decade many researcher are working on the combined effect of advance adsorption material and hydrodynamic cavitation (Bethi et al. 2016). Hydrodynamic cavitation is a new emerging technology utilized for the wastewater treatment in recent years. Cavities are synthesized with the mechanical devices like venture or orifice plate; sudden reduction in the flow area causes the pressure difference across the plate. The created pressure difference, when it crosses the vapor pressure of the liquid, partial vaporization of liquid started which synthesized vapor bubbles in the stream. These bubbles collapse when pressure drop tends to normalized slightly high pressure and temperature zones are generated. In hydrodynamic cavitation synthesis of theses, cavities and their collision are used predominantly. Cavities are produced by throttling of liquid through orifice and subsequent collision if cavities are achieved within a fraction of time. These collapses of cavities will generate high pressure about 1000 bar and temperature 1500 K locally. High temperature and pressure for very small instant of time are sufficient for the synthesis of radicals, which reacts with impurity, and reduction takes place (Sivakumar and Pandit 2002). New generation of micropollutants and bioactive compounds associated with polluted water needs sustainable systems with low energy consumption. Energy consumption associated with the process need to be considered, as only separation of pollutants will not solve the problem. The real-time control systems associated with the hybrid system is need to be designed for the quick response to the detected pollutants, and implemented for the wastewater treatment process to optimize the resources. Large energy consumption associated with wastewater treatment will add to the cost and may affect other counterparts of the environment such as non-renewable energy sources. In the present book chapter, we are going to overlook on the present convectional methodologies for the wastewater treatment, different hydride techniques like photocatalysis and Fenton process, hydrodynamic cavitation, ultrasound cavitation, ceramic and nano-ceramic membrane, and their application for the wastewater treatments. The need of these novel hybrid systems to present wastewater treatment is necessary. Nano-adsorbents are attractive due to their high surface area with small volume; in recent years they have proven promising solutions for the separation of heavy metals pollutants from wastewater. Bhatia et al. reported the various nanoadsorbents that include alumina, anatase, carbon nanotubes, chitosan, copper, iron and zinc oxide, magnetite, nanoclay, and zirconium nanoparticles. They gave overview on the synthesis of nano-adsorbents and their effective optimal conditions of operations. Different pH conditions and concentration of the wastewater, contact time of pollutants and nano-adsorbents, and amount of nano-adsorbents all these factors have prime effect on efficient use of nano-adsorbents (Bhatia et al. 2017).

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Different Types of Pollutants from Pharmaceutical and Textile Industries Pollutants from Pharmaceutical Industries

Pharmaceutical products are essential to cure the diseases of human as well animals. But, presence of the same in the environment including water resources poses serious problems. These are named as pollutants from pharmaceutical industry. The foremost resources of these pharmaceutical pollutants include hospital waste, treated and raw effluent of pharmaceutical industries, excretion of livestock which has been treated with antibiotics, wastewater treatment plants, and agricultural field’s runoffs. The country-wise, major pharmaceutical pollutants discharged and found in nearby water bodies are summarized in Table 8.1 (Larsson 2014).

8.2.2

Pollutants from Textile Industries

The textile applications are part of the modern world. The vast amount of wastewater is produced without proper treatment. This wastewater effluent is causing the serious threat to human health. Textile industry is the most chemical intensive and the largest user of water, hence the main polluter of potable water. The main chemical pollutants are produced from the dye usage in the textile industries. Textile effluents consist of biochemical oxygen demand and a high organic load, strong color, low dissolved oxygen, and low biodegradability. The average value of pH for textile effluent is 8.5. The major textile pollutants in the water bodies are summarized in Table 8.2 (Ghaly et al. 2014; Jaganathan et al. 2014). Table 8.1 Major pharmaceutical pollutants found in waterbodies S. No. 1

Country China

2

India

3 4 5 6 7 8 9 10

Denmark Germany Switzerland Norway Korea Israel Spain United States Pakistan Taiwan

11 12

Major pharmaceutical pollutant Oxytetracycline antibiotics, estrogenic sex steroids, penicillin G and its metabolites and many Fluoroquinolone antibiotics, salicylic acid anti-inflammatory drugs and many Sulfonamide antibiotics and intermediates/metabolites Phenazone and metabolites Venlafaxine—antidepressant Bacitracin—antibiotic Lincomycin—antibiotic Venlafaxine and metabolites, carbamazepine, and venlafaxine Venlafaxine Narcotic opioids Several antibiotics Sulfonamides, nonsteroidal anti-inflammatory drugs, and other drugs

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Table 8.2 The major water pollutants from textile industry S. No. 1 2

Chemical process in textile industry Sizing Kiering

3

Bleaching

4 5 6 7

Mercerizing Dyeing Printing Finishing

Possible discharged pollutants Benzene, resins, fats, waxes, starch, and glucose Caustic soda, waxes, soda ash, sodium silicate, and fragments of cloth Cyanide, wax, grease, soda ash, sodium silicate, hydrogen peroxide, acids Caustic soda Dye stuff, sulfides, acetic acid, mordant Dye, nitrate, phosphate, starch, gums, mordant, acids Lead, starch, tallow, salts, finishing agents

In the recent, research is more focused on removal of listed major pollutants from pharma industry and textile industry. Different methods have been explored for removing the pharmaceutical pollutants such as adsorption using activated carbon, graphene oxide and carbon nanotubes, biological treatment processes, advanced oxidation processes, and hybrid waste water treatment processes. Alone method is not efficient in removal of pollutants. Based on the nature of the pollutant, the combinations of treatment methods are adopted. These are known as hybrid treatment method and well explained in the preceding sections.

8.3

Conventional Methodologies for the Treatment of Wastewater

Water and air are the fundamental components for the existence of life. Yet being aware of its significance, it is poorly managed in the most parts of the world. Since the beginning of the industrial revolution, a wide variety of potentially harmful pollutants is being directly liberated into the natural water sources, thus making it unfit for consumption. These effluents contain high contents of heavy metals, harmful organic compounds, and biohazards that may bring serious implications on human health and ecosystem. In the year 2000, around 4000 million cases of diarrhea have been reported due to unsafe water. Nearly, 12% of birds, 24% of mammals, and around one-third of amphibians that are dependent on the inland waters are now on the verge of extinction due to water pollution (Earnhart 2013). In the current scenario of rapid industrial growth, the discharge of wastewater becomes inevitable. Therefore, much focus needs to be toward reducing the volume and strength of these pollutants to the maximum possible extent. This can be done by taking various “in plant” measures such as reducing and reusing water, substituting the chemicals that cause contamination in the process and good housekeeping practices, etc. A number of techniques are also available that have been developed keeping in view of

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the “end of pipe” treatment to minimize the waste and to decontaminate the wastewater. In general, these techniques in modern wastewater treatment can be classified into three categories depending upon the operational mechanism, i.e., physical, chemical, and biological (Kulkarni et al. 2000). A combination of these techniques is employed in a logical sequence that depends upon the type of organic and inorganic pollutants present in the wastewater for carrying out the treatment efficiently. Today the wastewater treatment for industries has become a very complex issue. This is because of the wide un-uniform quality of wastewater produced by different industries at different scale. In such a scenario, the concept of common effluent treatment plant is a promising solution for treating these industrial effluents at low cost. This is also widely accepted among small- and medium-scale industries where an individual wastewater treatment plant is not affordable because of space and economical constraints. A conventional design of the industrial wastewater treatment plant involves various unit operations and processes (Patwardhan 2017). These are as discussed in the following sections. Screening This operation is usually employed in industries such as textile, leather, pulp, and paper mills where good housekeeping techniques are not practiced properly. In this process, the large floating and suspended solids are removed by using screens of different sizes. It also protects pipes, valve, and pumps from clogging and further damages. It is important that the organic solids would pass to the next unit operations and not get obstructed during screening operation. Also in order to prevent screen blockages and pungent smell, the captures solids are removed in regular periods of interval (Ng 2006). The solids screened out from the effluent needs adequate arrangement for proper disposal and treatment at the end of the process. Sedimentation and Flotation The screening process is done for separating large suspended solids, while the sedimentation and flotation operation is carried out to remove the colloidal solids. This is also called “negative sedimentation” as the impurities take prolonged time periods to get settle. In this process, the waste water is allowed to retain in a sufficiently large tank for a considerable period of time to separate particle through gravity separation (Peterson 1980). This operation is involved in industries that produce substantial amount of suspended solids (both organic and inorganic) such as dairies, pulp, and paper mills, slaughterhouses, oil refineries, etc. This operation not only reduces the solids load for further treatment process but also recovers the useful material that can be reused by the industries. Filtration This operation is often used as a pretreatment process to reduce biological oxygen demand loads. The pretreated water was later discharged to reverse osmosis unit or activated carbon column for further treatment. The operation can be used where water is used reused for irrigating land, lawn sprinkling, groundwater injection, etc.

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Mixing Mixing is an operation which is used for a number of applications in wastewater treatment plant such as blending of disinfectant in treated water, flocculation of colloids with coagulants, mixing neutralizing chemicals to adjust the pH level of wastewater streams, skimming of floating oil and grease, etc. It is also used for maintaining fresh condition of water and for equalization where the quality and flow rates are highly fluctuating. Drying and Incineration This operation is practiced to handle the wastewater sludge and to remove toxic substances that would have a negative effect when comes in contact with environmental medium such as soil. Drying is employed by spreading the layers (thickness 20–30 cm) of sludge on a dry sand bed. A fraction of moisture percolates through the sand drying bed and gets filtered while the remaining evaporates into the atmosphere. The filtered water is recycled back to the plant, while the dried sludge is removed from the drying beds. The dried sludge is further incinerated if it contains toxic contents. Adequate measures need to be adopted while in incineration to avoid the risk of air pollution. Freezing In this operation, the industrial effluent is cooled initially by using water and then frozen by refrigerants (such as butane). During this process, the ice crystals formed is fundamentally pure water. The operation is carried out until half of the water is frozen to ice. The unfrozen liquid is separated by draining, while the ice is melted back to form pure water (end product). It has been found out a 100% separation can be achieved in most of the cases (Lorain et al. 2001). Foaming This operation is useful for the industries that produce surface active pollutants such as detergents. In this, the pollutant is removed from the wastewater by creating air bubbles (foam). The pollutant-enriched foam is separated from the wastewater at the liquid–gas interface. This foam is further collapsed to create the solute-rich liquid product. Apart from separating the pollutants, foaming can also provide a considerable insulation to the activated sludge tanks in winter season (Corbala-Robles et al. 2016). Dialysis and Osmosis In this operation, mineralized water is passed through a series of membranes where anions and cations electrodes are aligned alternated. The ions present in the water migrate toward these electrodes (either cation or anion) and pure demineralized water was collected at the end of this operation. A major limitation with this process is the ability to remove only ionized contaminants present in the water. In case of two different concentrated solutions present in the water, a semipermeable membrane is used that works on the principle of reverse osmosis when applied to pressure. Generally, this operation is applied to the cotton textile industry, for the recovery of caustic soda spent during mercerizing of fabrics.

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Adsorption This operation is applied for the removal of particularly those organic contaminants that are present in the dissolved state and remains resistive to biodegradation. In this process, granular activated carbon is used to adsorb the colloidal and dissolved organic matter. The non-pathogenic bacteria that also grow on this media consume the difficult-to-degrade organic matter and ultimately get desorbed and returned to the aeration tank. This operation usually employed in the textile and pharmaceutical industries that produce organic waste such as dyes, phenols, etc. Gas Transfer Gas transfer is an operation which is usually employed in aerobic processes such as trickling filtration, activated sludge process, etc. It is useful for the industries that produce volatile solvents which can be largely removed by aeration and thus helps to reduce the chemical oxygen demand of the wastewater. It is also used to remove nitrogen from water by converting it to ammonia gas and for the transfer of chlorine from gaseous to liquid phase. This operation has become a very essential element from the environmental point of view, and subsequently a number of technologies have been developed over the years with an aim to enhance the gas transfer process such as aeration diffusers, membrane stripping, and packed tower air stripping (Castro and Zander 1995). Elutriation Elutriation is basically used to reduce the alkalinity of the digested sludge by washing it with low alkaline water. This operation helps to reduce the consumption of conditioning chemicals used for anaerobic digestion of sludge. In this process, the digested sludge and elutriate is mixed in the ratio of three to five times (on volume basis). The soluble degradation products is carried away by the elutriate, while the sludge is allowed to settle and then further forwarded for the next wastewater treatment operations (Baskerville et al. 1981). A major drawback, however, with this operation is the washing away of finely divided solids during the process that cannot be reentered back into the wastewater treatment facility. pH Correction pH correction is an important process that is employed universally for all industrial wastewater treatment plants. This treatment is vital before it is sent to biological treatment, disinfection with chlorine, coagulation, nitrification, denitrification, etc. Chemicals such as sulfuric acid, nitric acid, sodium hydroxide, ammonium hydroxide, lime, etc. are commonly used for this process. It is recommended to carry out a laboratory-scale study to choose the right neutralizing chemical and optimum dosage. Coagulation Coagulation is used to remove the suspended solids through sedimentation in the wastewater treatment plant. In this operation, coagulants is used that adsorbs the suspended organic impurities and form large aggregates. These aggregates are further allowed to settle through sedimentation process and subsequently removed (Jiang 2015). Most commonly chemicals used in this process as coagulants are lime,

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alum, ferrous sulfate, and ferric chloride. Adequate laboratory trial test is needed for the selection of appropriate coagulant before introducing them in the wastewater treatment plants. Oxidation and Reduction This operation is mainly used to reduce biological oxidation demand and odor control. It is used in industries that contain pollutants such as heavy metals, phenols, cyanide-bearing waste, oil refineries waste, food processing waste, etc. Most commonly, chlorine is used as an oxidizing agent in this process. It is well-known the quality of industrial wastewater varies with time. Therefore, it is advisable to conduct a proper laboratory test to identify the right amount of oxidizing agents. Aerobic and Anaerobic Processes This process is aimed to purify the wastewater from biodegradable organic matter with the help of microorganisms that may be inherently present or else introduced in the wastewater. While in the aerobic process, the degradation of organic matter takes place in oxygen-rich environmental conditions, the same is not needed for anaerobic digestion. Appropriate environmental conditions, food, and oxygen requirement play a critical role for the proper growth and functioning of these microorganisms and thus to enable smooth operation of the treatment plant.

8.4

Application of Nanomaterials as Photocatalysts for Wastewater Treatment

The catalysts which absorbs the light to accelerate the kinetics of reaction and achieve the required conversion are generally termed as photocatalysts. Photocatalytic activity of catalyst depends upon its unsaturated stable structure, its ability of production of free radicals, and presences of or creation of unpaired electron, which are responsible for a secondary reaction like water splitting and reduction of pollutants. In recent years many catalysts were synthesized and utilized as photocatalysts. Titanium dioxide and zinc oxide are highly physicochemical stable compounds utilized as effective photocatalysts (Lv et al. 2017; Pirhashemi et al. 2018). It has been reported that many researchers worked on the improvement of zinc oxide photocatalytic activity by coupling it with different materials like graphene or manganese doping and all (Li et al. 2017; Liu et al. 2017). Doping action in zinc oxide will enhance its light absorption activity, coupling with graphene-like materials which will reduce the recombination of synthesized electron again. It was found that a high combination rate of photo-induced electrons is present with the manganese-doped zinc oxide pair. Due to which one can achieve the highest degradation efficiency among all other doped samples under visible light. It was reported that the photocatalytic efficiency of undoped and doped zinc oxide could be significantly improved by hybridizing it with graphene oxide (Labhane et al. 2016; Labhane et al. 2018a, b). P.K. Labhane et al. significantly used the hydrothermal

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method for the synthesis of manganese-doped zinc oxide nanoparticles which is imbedded on reduced graphene oxide sheets. They have found the synthesized photocatalyst shows quite better performance for the degradation of Rhodamine B and Congo red dye pollutants. The synthesized photocatalyst of manganese-doped zinc oxide embedded on reduced graphene oxide, utilized for the absorption of solar light which will avail the presence of unpaired electron induced by photolysis. Production of these unpaired electrons under sunlight is much higher which leads to the synthesis of radicals as shown in Fig. 8.1 Congo red and Rhodamine B dye from polluted wastewater when interacting with manganese–zinc oxide/reduced graphene oxide; presence of manganese2+ ions plays an important role in degradation mechanism. Recombination of synthesized electron restricted by the presence of Manganese2+ ions and similar charge repulsive forces. This leads tapering of released electrons in the zinc oxide band. It was found that the accumulation of unpaired electrons in the catalyst reduced graphene oxide nanosheets, due to which synthesis of hydroxyl radicals is more efficient. Produced hydroxyl radicals react with dye molecules and braking action takes place. These dye molecules react with hydroxyl radicals and breakdown into its less harmful colorless forms. A probable mechanism of degradation of the dye molecule is shown in Fig. 8.2. Many researchers are reported the use of titanium dioxide as the photocatalyst during the advance oxidation processes under the exposure of light with the combination of oxidants like hydrogen peroxide, ozone to synthesize hydroxyl radical. Produced hydroxyl group reacts with organic pollutant and degradation or breakdown of pollutant achieved. Industries like textile, the effluents contains a high concentration of organic pollutants in the form of color and its pigments. These

Fig. 8.1 Synthesized photocatalyst of manganese (Mn)-doped zinc oxide (ZnO) embedded on reduced graphene oxide (RGO)

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Fig. 8.2 Synthesized photocatalyst of manganese (Mn)-doped zinc oxide (ZnO) embedded on reduced graphene oxide (RGO). (Reprinted with permission of Chemical Physics Letters, Elsevier from P.K. Labhane et al. 2018a, b)

effluents are harmful to aquatic as well as human lives. The toxic color pigments increase the toxicity, as well as when these micro color pollutants adsorb on the skin of aquatic bodies which disturb the food chain of aquatic life. Titanium dioxide is semiconductor as it covers a large range of band gap; thus the photocatalytic applications of titanium dioxide are more popular. The basic difference between other semiconductor photocatalyst and titanium dioxide is that it can operate on the ultraviolet range. Titanium dioxide produces electron pairs and holes in the conduction even it is exposed by ultraviolet irradiation, as its surface gets excited. Due to the synthesis of electron pair, a positive hole will be generated which attracts the water molecules and reacts react to form the hydroxyl group. Series of reactions will start on the surface of titanium dioxide, and the oxidation process will be continued. The presence of photocatalyst and light source is required for the advance oxidation process, so any one of them can play the limiting element that terminates the process. For the cost-effectiveness of the process, sunlight is used as a source of light that needs to develop the catalyst working at sunlight band gaps. Some of the sunlight excited photocatalyst reported for the advance oxidation process are given in Table 8.3.

8.5

Hybrid Wastewater Treatment Techniques

New-generation pollutants mainly in micro and nanosize are hazardous to the environment; to treat these pollutants, hybrid wastewater treatments are essential. Cavities generated during the flowing flow can be used for the synthesis of high energy zone in hydrodynamic cavitation, but only high energy zone impact will not

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Table 8.3 Sunlight-induced nanomaterials as photocatalysts in wastewater treatment Photocatalyst composition Titanium dioxide/cadmium oxide—zinc oxide (SuarezParra et al. 2003) Sulphur-doped titanium dioxide(Yu et al. 2005)

Zinc oxide/aluminum film (Peng et al. 2006)

Titanium dioxide (Khataee et al. 2009) Non-metal-doped titanium dioxide nanoparticles (Pelaez et al. 2009)

Er3+:YAlO3/ zinc oxide composite (Wang et al. 2009) Graphene decorated cadmium sulfide particles L-Glutathione doped with tin sulfide ultrathin nanosheets

Application Solar energy activation Degradation of blue azo dye A visible light used for activation Antibacterial activity increased due to doping Micrococcus lylae bacteria removed Visible light excitation Removal of phenol from the waste water achieved High decolorization activity found Visible light used Antibacterial activity increased due to doping A visible light used for activation Microcystin-LR removed Prepared by ultrasonic dispersion Removal of dye acid red B Removal of dye Rhodamine B Removal of methyl orange and chromium

be sufficient to treat the pollutants. Synthesized cavities generated the pressure impact that can sometimes break the pollutants, but cannot completely degrade in their constituent elements or separate them. This situation needs additional unit operation combination with cavitation; in general, advance oxidation, adsorption processes, and filtration process can be used. Nanofiltration process is itself more complex due to the clogging action of nanopores on filter papers that makes the filtration process more complex and costlier. The advance adsorption process is one of the advisable processes with a synthesis of more advance adsorbents. The cavities that are produced due to the throttling effect of orifice tend to collapse with highpressure impact which disassociates the water in the unstable zone. A combination of hydrogen peroxide will help to enhance this synthesis of radicals of water (Doğan et al. 2006). The concentration of hydrogen peroxide plays an important role in the combined effect of cavitation and hydrogen peroxide addition, because after a certain concentration of hydrogen peroxide, the maximum limit of radical formation is attained after which the concentration of hydrogen peroxide will not be effective. The reversible reaction of water dissociation as the effect of hydrogen peroxide is as follows (Shirsath et al. 2011): OH
þ H2 O2 ! H2 O þ HO
2 Only the synthesis of free radicals will not effective. The utilization of these free radicals in reaction with the pollutants is essential; contact time and patterns by which radicals react with pollutants are very much important. Some operating

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conditions of combine systems also affect the intensity if treatment like turbulence generated in the flowing system, operation pressure, pH, and type of micro-/ nanopollutants will decide the success of the hybrid technology. Different properties of wastewater also effect on the operating conditions, and requirements of cavitation are discussed below.

8.5.1

Effect of Wastewater Properties

The magnitude of effect due to different wastewater properties on the cavitation process is much high as compared to other parameters in cavitation. For example, it was found that maintaining of required pressure difference is very difficult for development of cavities and reach up to the required effect as the density of the wastewater increases due to high concentration of pollutants. The inception of the bubble generated won’t be that effective to generate a high-energy zone. Due to low energy dissociation, less number of free radicals will be generated, and cavitation is not that effective. Thus, there is a requirement of high-pressure drop generation across the flow disturbance used to synthesized cavitation (orifice). This is only one property of wastewater which affects seriously likewise; the following are the parameters related to the wastewater which effect on the efficiency of cavitation hybrid process: 1. 2. 3. 4. 5.

Wastewater average vapor pressure Wastewater viscosity Wastewater surface tension Wastewater bulk temperature Presence of dissolved solids, gasses in wastewater

Wastewater consists of various gasses, as well as high vapor pressure components, which have a wide effect on the synthesis of cavities in flowing systems. Increased gas content in the wastewater may directly decrease the intensity of the inception of the bubble in wastewater. This phenomenon may be because of less pressure drop due to gas content. Cavities are produced in the wastewater when desired vapor pressure variation is achieved, and bubble growth is depending upon the cohesive forces of liquid phase interaction with the cohesive force of the gaseous phase inside the bubble. Increasing viscosity of wastewater shows an increment in cohesive forces that ultimately affect a minimum requirement of cohesive forces of the gaseous phase inside the bubble. It was found that cohesive forces requirement increase with increment in viscosity of wastewater. All these properties of wastewater interact with the hybrid system and accordingly efficiency of the combination of hybrid systems will change. The effectiveness of a hybrid wastewater treatment system is decided by threshold values of the parameters like the amount of surfactant added or advanced oxidation agent used, orifice design, turbulence synthesized and pressure drop develops across the cavitator (obstacle) all has threshold values that ultimately decide the effectiveness of a combination of unit operations.

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Hybrid Wastewater Treatment Techniques Using Hydrodynamic Cavitation

Conventional methods are inappropriate or sometimes insufficient to treat many hazardous pollutants that include micropollutants, bioactive compounds, and non-degradable dissolved solids and all. The large quantity of water used by industries continued the contamination of water day by day with new contaminations, and the need for robust, sustainable, and low energy required process of water treatment is required. The hybrid wastewater treatment process with the mentioned needs is much important because only one treatment operation will not resolve the problem but also we need to understand the requirement for the development of low energy consumption solutions on wastewater treatment, as they are future needs. One of the energy conservative hybrid systems combines convectional processes and develops a more efficient robust and sustainable process. The combination of appropriate conventional water treatment processes with each other will develop an optimal energy requirement process. The hybrid system developed by a combination of hydrodynamic cavitation and adsorption has significantly more effective than that of individual operations of both conventional processes. The development of a hybrid system needs to understand the merits and demerits of individual convectional processes for the significant development of the system. Cavitation techniques that are used in combination with the adsorption process may work effectively, as advanced oxidation process and separation of pollutants using the adsorption process. Orifice plate is used in the cavitation in which due to sudden reduction in the area of flow, the liquid is exposed to high-pressure zone; this change of pressure leads to small vapor droplet generation. Progressively development of these liquid–vapor droplets through the liquid stream and sudden collision at the normalized pressure develop a high-energy zone in flowing liquid. This high-energy zone is utilized in hydrodynamic cavitation techniques. The collision of droplets develops the high energy zone, which develops the hydroxyl radicals in the water. Oxidation and breakdown of long-chain organic polymers are achieved that makes easy to separate the micropollutants with the adsorption method (Lü et al. 2014). The pollutants like methyl group used in the textile industry are harmful to the environment, as they present in combination with tri-phenyl in the form of the long-chain polymer. Such micropollutants are much stable and high energy required for their break down. Because of which treatment of textile industrial wastewater is much costlier. On the other hand, hydrodynamic cavitation is the most effective and energy-efficient way to breakdown the tri-phenyl methyl group also the achievement is the most energy-efficient way. The tri-phenyl methane dye is conventionally removed from the wastewater by adsorption process but the poisoning of adsorbent make process ineffective cost-wise. The desorbing of the tri-phenyl methane group is also a major issue in post wastewater treatment process (Wei et al. 2014). It has been reported in the literature the efficiency of the adsorption process can be increased by using novel polymers having improved properties of adsorbent. Researchers noted that polymer hydrogel is used for the wastewater treatment

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Orifice Plate

Absorption Chamber Bypass Line

Storage Tank

Centrifugal Pump

Fig. 8.3 Schematic diagram of combined hydrodynamic cavitation and hydrogel adsorption column setup

effectively (Brannon-Peppas and Peppas 1991). Multi-dimensional cross-linked polymers of hydrophilic monomers are generally used as hydrogels. The hydrogels are synthesized by step-growth process to give the netlike structure, which helps to develop more hydrogen bonding with water and impurities associated with the water (Özcan et al. 2007; Sivudu and Rhee 2009). Due to their ability to swell 200% times of their original volume, they respond with a large quantity of water and purification is achieved. The effective use of hydrogels is controlled by pH of wastewater, temperature, electric conductance, magnetic conductance, total organic carbon, and sometimes photolytic activity (light exposure) of the effluent (Doğan et al. 2006; El Qada et al. 2006). The hybrid system develops after a combination of hydrodynamic cavitation and the adsorption process is represented in Fig. 8.3. The treatment of wastewater contaminated with tri-phenyl methane group (dye wastewater) has divided into two sections: the first section will comprise of the orifice plate, and the second section is the adsorption chamber as shown in Fig. 8.3. The hybrid system developed with the combination of these has improved results in all energy-optimized way. In the first section of the hybrid system, the bubble behavior plays an important role. The high energy zone of the system is affected by the nucleation of the bubble, prolong stabilization, growth of bubble, a collision of the bubble, and the distance at which collision occurs. Bubble behavior mostly depends upon the design of cavitator and the operating conditions. The cavitator is nothing but the simply designed orifice plate, which has an opening orifice of diameter at least equal to 1/100 of a diameter of the pipe, utilized. Here, the operating parameters are pressure difference generated across the cavitator, viscosity of the wastewater and the concentration of micropollutants. Cavities generated collapse at a certain fixed distance from the mechanical cavitator; this high energy zone sometimes observed operates at a pressure of 1000 bar and temperature about 1500 K (Gogate and Pandit 2005). Hydrodynamics of bubble is dependent upon cavitation number, the velocity of flow, and intensity of mixing or aggressiveness of fluid.

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Fig. 8.4 Geometrical dimensions of the orifice (cavitation device)

1mm 125mm

Synthesis of bubbles at a high rate with required growth in a certain range and their collapse all these lead to production of hydroxyl group radicals by dissociation of water into their two basic functional groups. These reactive groups when synthesized immediately start to react with the other molecules present in surroundings. The components like tri-phenol-methyl groups that need high energy then get in contact with the hydroxyl group and get oxidized which ultimately break the long chain of the polymer into parts. The typical orifice plate used for the cavitation is shown in Fig. 8.4, which is a single opening at the center. Depending upon the requirement, these opening on orifice plate may increase. In the second section of hybrid process convectional adsorption carried out with improved adsorbents. Conventionally used adsorbents have the problem of poisoning due to which the process becomes excessively costlier; thus the need for the new generation of adsorbents is needed with cost-effective production and effective utilization over the large quantity of wastewater. Nanocomposite clay-based co-polymers are one of the solutions for this. Polylactic acid, polyacrylic acid, and poly-acrylamide are used to modify the clay. To initiate polymerization process, monomers have gone through ultrasonic irradiation for the synthesis of hydrogels. Uniform distribution of clay across the hydrogel and free radicals generation are the two most effective benefits of ultrasound irradiations. Hydrogels utilized for the adsorption process are studied and reported by many researchers. They also found that with the nano or micro-materials like kaolin, aluminum phyllosilicate clay, and phyllosilicate (minerals), such inorganic materials have a wide surface area and show the good capacity of adsorption, when combined with hydrogels. Incorporation of such nanomaterials is easy with ultrasound irradiation. Due to the swelling property of the hydrogel in contact with water, the nanocomposite clay surface is exposed to the impurity associated with the wastewater. The water along with impurity will penetrate more in intermolecular layers of hydrogel and clay combination due to their nanostructure. These high penetrations in intermolecular space will increase the adsorption of impurity from wastewater. Enhancement of adsorption efficiency as well in the mechanical strength of the polymer hydrogels can be achieved by a combination of clay. Hydrodynamic cavitation combined with a packed bed hydrogel process has been proposed with the aim of the treatment of a bulk quantity of wastewater. The clay

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naturally posesses the high chemical stability, chemical resistance, and mechanical strength which are most desirable and abundant availability makes them suitable for use. Synthesis of polymer hydrogels in the presence of ultrasound facilitated the uniform dispersion of kaolin clay in the hydrogel matrix that helps in increasing the strength of hydrogel as well as enhancing the adsorption capacity of the hydrogels.

8.5.3

Ultrasound Cavitation and Applications in Wastewater Treatment

Sound energy utilized and transfer in the form of vibrational energy, through sound waves. Ultrasound waves with high energy will be generated with a sonicator. These high-energy sound waves produce cavities in the liquid phase for a very small instant of time. Like in the cavitation process, these cavities (bubbles) have a very short life span, collision/inception of these bubbles generates high-energy zone in liquid. Though the fundamental principle of treatment through the cavities is the same, both cavitation process and ultrasound techniques are different due to their mode of synthesis of cavities. The addition of catalytic reactions or advance adsorption materials (process) will help to treat the wastewater. It was reported that many advance hybrid materials are synthesized through ultrasound treatment. The following is the case study for the removal of phenol using the combined effect of ultrasound sonication and nanoclay.

Case Study on the Combined Effect of Ultrasound and Nanoclay on Adsorption of Phenol The common aromatic organic compound phenol is found in most of the industrial wastewater and sometimes in domestic wastewater also. Due to its two functional group hydroxyl group and phenolic functional group, it shows more reactivity to each chemical including water. The presence of a phenolic group makes him carcinogenic compound to human beings so it comes in the hazardous category of chemicals. The phenolic group based chemicals are commonly found in wastewater which comes from pesticides, paints and petroleum-based industries. Two functional group contents of phenol make it difficult to treat with conventional chemical treatment; as well, it is also not suitable for biological treatments. Chemical degradation with oxidation method or physicochemical separation is a very high-energy consuming process used for wastewater treatment, which has low efficiency. Ultrasound cavitation comes up with an efficient tool for the degradation process of the phenol in wastewater (Shirsath et al. 2011). Surfactant utilized plays an important role in the degradation of phenol in a highly intensified ultrasound device reported by Sonawane and co-workers (Shirsath et al. 2013). They studied the performance of bentonite nanoclay for adsorption of phenol assisted with ultrasound.

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Organically modified clay shows improved adsorption of phenol under sonication bath assistance. Natural bentonite clay modified with the different surfactants like tetrabutylammonium chloride, cetyltrimethylammonium bromide, and hexadecyltrimethylammonium bromide to study the performance in the adsorption of phenol is investigated in a bath sonicator. The Sonawane research group (Shirsath et al. 2013) studied that sonicated modified nanoclay gives much higher adsorption performance (near to 50%) compared to that of unmodified nanoclay (~35%). It was also reported that the adsorption of phenol with modified nanoclay is time-dependent process. Adsorption isotherm achievement with modified nanoclay achieved faster with the high amount of phenol adsorption. Depending on the results it is found that modified nanoclay with tetrabutylammonium chloride surfactant is the best suitable for adsorption of phenol.

8.5.4

Hydrodynamic Cavitation and Photocatalysis with Photo-Fenton Process

The degradation of methyl orange and azo dye was carried out using the hydrodynamic cavitation. In this work, the Fenton process was also used along with the venture. It is found that multi-fold degradation of the cavitation is increased if the hydrogen peroxide is added. They attempted iron as advanced hybrid additives into the process (Innocenzi et al. 2019). In one of the attempts of degradation of the Sulfadiazine which is pharmaceutical intermediate. In this work, hydrodynamic cavitation along with nanosize iron oxide particles was used for the degradation of the sulfadiazine. The degradation of the pharmaceutical pollutant is higher as hydrodynamic cavitation brings the renewal surface of the iron oxide nanoparticles hence there is higher degradation with this kind of hybrid system (Roy and Moholkar 2019). Ultrasound-assisted photo-Fenton degradation of the sodium alginate was carried out. In this work, titanium oxide was used as photocatalyst. It is found that the combination of the photocatalytic with ultrasound and the Fenton process able to broke the ester bonds of effluent due to strong oxidation. It is possible to convert the sodium alginate into the lower molecular weight compounds (Zhou et al. 2017). Salesh and Taufik (2019) attempted to synthesize the combination of iron oxide and zinc oxide catalyst loading onto the graphene using hydrothermal methods. In this work, they used zinc oxide as photocatalyst and iron oxide as a Fenton catalyst. The loading was carried out onto graphene as there will not be any kind of leaching of the catalyst. They found that the complete degradation of the dye will take place in 1 h. They carried out some experiments to reuse the catalyst. Rajoriya et al. (2019) prepared the nitrogen-doped titanium dioxide photocatalyst for photocatalytic degradation for acetamidophenol using the combination using hydrodynamic cavitation and acoustic cavitation. Titanium dioxide is synthesized using the sol–gel process. The degradation is ninefold more than the sono-photocatalytic system. The oxidation

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components were identified using mass spectroscopy. Josih and Gogate (2019) reported industrial wastewater treatment carried out using hydrodynamic cavitation and advanced oxidation process. The combination of the oxygen/hydrodynamic cavitation/Fenton with hydrogen peroxide loading. It is reported to the best hybrid treatment for industrial wastewater treatment. Thanekar and co-workers (2018) used the advanced oxidation process using hydrodynamic cavitation for carbamazepine that is a pharmaceutical complex. The combined process gives a synergistic effect. The highest degradation of the dye is achieved by combining the hydrodynamic cavitation with oxidation and ozonation process at optimized process condition. This process has potential to scale up at industrial scale due to higher cavitational yield and comparatively less cost. Suresh et al. attempted to synthesize the combination of the bismuth-doped titanium dioxide photocatalyst using the sol–gel process (Kumar et al. 2017). An attempt was made to degrade methylene blue using hydrodynamic cavitation and photocatalyst. They found that degradation of methylene blue using hydrodynamic cavitation and hydrogen peroxide gives better degradation than the photocatalytic process which is higher more than 94%. Suresh Kumar et al. studied the degradation of the three different dyes combined present in wastewater by using hydrodynamic cavitation and advanced oxidation process (Kumar et al. 2018). In this case the number of attempts was made in combination with hydrodynamic cavitation using hydrogen peroxide and photocatalyst, Fenton process. The combination brings first-order kinetics. The complete decolorization brings within 40 min. Gagol et al. reported the review on the treatment of wastewater treatment using advanced oxidation using the cavitation process. Gagol et al. also reported phenol, nitrophenol pharmaceutical components, and dyes degradation. The number of parameters such as pH, temperature, and concentration of pollutants was optimized (Gągol et al. 2018). Economic analysis was carried out for the industrial wastewater treatment process. The advantages and limitations were discussed for the individual components.

Case Study: Degradation of Dye Wastewater Using Advanced Oxidation Process Along with Hydrodynamic Cavitation An attempt was made to synthesize the titanium oxide and bismuth using the sol–gel process using the cavitation technique. Then the hydrodynamic cavitation and photocatalyst and hydrogen peroxide were used for the degradation of the methylene blue. The photocatalyst was characterized using the transmission electron microscopy and X-ray diffraction analysis. It is found that the X-ray diffraction shows the anatase phase with bismuth components. From transmission electron microscopy image, it is found that the particle size is less than 35 nm. The addition of bismuth components was used to hybrid photocatalyst. The optimization study of hydrodynamic cavitation was carried out for different pressures from 1 to 10 bar inlet pressure. The optimized pressure was 5 bar. Optimization for different pH was

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%decolorization of MB dye

70 HC+Photocatalytic

60 50 40

HC

30

Photocatalytic

20 10 0

3.41

2.49

8.62

Reaction rate constant(k)×103 min-1 Fig. 8.5 Plot of the rate constant against percentage decolorization of methylene blue (MB) using hydrodynamic cavitation (HC), photocatalytic, HC + photocatalytic. (Reprinted with permission of Chemical Engineering and Processing, Elsevier from Kumar et al. 2017)

also carried out from the pH at 2 to 10. It is found that the decolorization is highest at 2 pH for 50 PPM solution of the dye at 5 bar inlet pressure. As shown in Fig. 8.5, it is found that hydrodynamic cavitation will give degradation around 30%, while less than 30% degradation is given by the photocatalytic system alone, while a combination of the hydrodynamic cavitation and photocatalytic system gives 60% more degradation (Kumar et al. 2017). As shown in Fig. 8.6, there is the comparison for the energy efficiency for different types of individual system such as hydrodynamic cavitation, photocatalytic system, and hydrogen peroxide alone. The combinations of energy efficiency of hydrodynamic cavitation (HC) + photocatalytic and hydrodynamic cavitation plus hydrogen peroxide (H2O2) were also reported. It is found that using the hydrodynamic cavitation plus hydrogen peroxide shows the highest energy efficiency mg of total organic carbon (TOC) reduced /J. There is an optimum concentration of hydrogen peroxide is 1:20 at which gives the maximum decolorization, while the hydrodynamic cavitation and visible photocatalytic system give more synergy; however, it is energy-efficient hydrogen peroxide addition.

8.5.5

Hybrid Treatment Techniques Using Photocatalysis and Nanofiltration

Photocatalytic oxidation has been studied for many decades as it is an advanced oxidation technology for reducing water pollution. Photocatalysis oxidizes and degrades various recalcitrant organic and inorganic pollutants present in aqueous media. Photocatalysis is classified into homogeneous and heterogeneous wherein the

Energy efficiency (mg of TOC reduced / J)

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0.000007 0.000006 0.000005 0.000004 0.000003 0.000002 0.000001 0

only HC

H2O2 alone

Photocatalyc

HC+H2O2

HC + Photocatalysis

Fig. 8.6 Comparison of energy efficiency of different process, the variation of total organic carbon (TOC), with different processes hydrodynamic cavitation (HC) and other combinations photocatalytic. (Reprinted with permission of Chemical Engineering and Processing, Elsevier from Kumar et al. 2017)

catalyst is in a homogeneous state in the former case; hence the recovery of catalyst is difficult after completion of the process. Whereas, it is easy to separate and reuse the catalyst in case of later due to the suspended form of catalyst. The present section explores the efficiency of the combined operation of photocatalytic oxidation and nanofiltration for the decolorization and degradation of methylene blue and crystal violet dyes. This work made an important contribution to the combined process of photocatalysis and nanofiltration. The individual effect of photocatalysis and nanofiltration processes and their combined effect were assessed for zinc oxide and titanium dioxide catalysts for crystal violet dye degradation. To the best of our knowledge, no such study is reported so far in the literature for the decolorization and degradation of dyes from aqueous media. The schematics of the integrated system exercised in this work for decolorization and degradation of dye from aqueous media is shown in Fig. 8.7. A feed tank of 10 L volume is mounted with a 125 W ultraviolet light from the top. The control valves, rotameter, and pressure gauges were arranged at desired locations between the reactor and nanofiltration unit to adjust and support the required flow rates and pressures. For the higher efficiency and enhancement in the degradation and decolorization of dye, the retentate from the nanofiltration unit was recycled to the feed tank. The permeate was collected in test tubes at every 15 min of the time interval. During the degradation process, the permeate and retentate samples were regularly collected every 15 min and analyzed for dye concentration in both permeate and retentate. The permeate volume flux (J) is determined using the following Eq. (8.1):

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Fig. 8.7 Experimental setup of hybrid system, (1) ceramic nano porous membrane, (2) photocatalytic reactor, (3) 0.87 horse power centrifugal pump, (4) pressure gauge, and (5) manual control valves

J¼

V 2 3 m =m day t:A

ð8:1Þ

where: V ¼ permeate volume (m3) t ¼ time (day) A ¼ effective membrane area (m2) The percentage rejection of dye by the nanofiltration unit was determined using Eq. (8.2). Dye rejection ð%Þ ¼

Cr  Cp
100 Cr

ð8:2Þ

where Cr ¼ concentration of dye (g/m3) in retentate Cp ¼ concentration of dye (g/m3) in permeate From the results, it is observed that the permeate flux of membrane was increased when the inlet pressure was increased from 2 to 4 bars (Fig. 8.8). However, with the progress of the operation, the flux through the nanofiltration was reduced from 1.66 to 1.43, 1.86 to 1.54, and 2.08 to 1.7 l/m2.min at 2, 3, and 4 bars, respectively, over 90 min of operation. The reduction in flux across the membrane could be because of concentration polarization or fouling of organic pollutants on the membrane surface. From the obtained results, it is identified that 4 bar inlet pressure has shown higher permeate flux across the membrane among different inlet pressures.

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Fig. 8.8 Inlet pressure and its effect on permeate flux in the hybrid system

The effect of catalyst loading on the dismissal of dye was studied by varying the dosage of zinc oxide catalyst in the photocatalytic reactor. In the typical experiments, the photocatalytic reactor was charged with 5 L of dye wastewater containing 500 mg/L of concentration. Dye removal studies were carried out using a hybrid system at various catalyst dosages such as 0.1,0.2, 0.5, and 1.0 g/L. The operating temperature of the hybrid system as kept constant at 25  C. From the results, it was identified that the 0.2 g/L of zinc oxide catalyst has shown maximum photocatalytic activity toward mineralization and decolorization of dye in aqueous media. Around 80% of the decolorization of dye has been achieved by the integrated process of photocatalysis with nanofiltration. The effect of the dosage of zinc oxide catalyst for decolorization and the percentage reduction of dye is shown in Fig. 8.9a, b, respectively. From the achieved results, it is observed that 0.1, 0.2, 0.5, and 1.0 g/L of the catalyst dosage has resulted in 76, 80, 28, and 14% of dye decolorization, respectively, in the combined system. Catalyst dosage beyond 0.2 g/L has shown lower dye decolorization. This effect might be due to aggregation of catalyst particles at excess loading which shows that the adsorption of dye molecules decreases due to the depletion of total available active sites on the catalyst surface. Moreover, the higher loading of catalyst resulted in high turbidity of wastewater that minimizes the transmittance of light across wastewater during irradiation. This ultimately deactivates the active catalyst molecules by colliding with ground-state molecules (Neppolian et al. 2002). This study concludes that 0.2 g/L zinc oxide dosage is sufficient to attain better decolorization of dye using the hybrid system. Therefore, 0.2 g/L of zinc oxide catalyst is observed as an optimal catalyst dosage for dye removal, and further studies were carried out at this optimized catalyst dosage. Different initial pHs ranging from pH 1.0 to pH 10.0 was studied for decolorization of dye and removal of total organic carbon using the hybrid system. The pH of wastewater was adjusted by using diluted sodium hydroxide and sulfuric acid. The

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a Dye Concentration (mg/L)

600

ZnO 0.1 g/L ZnO 0.5 g/L

ZnO 0.2 g/L ZnO 1 g/L

30

60

500 400 300 200 100 0 0

b

15

45

75

90

Time (Min) 90 80

76

80

% Decolourization

70 60 50 40 28

30 20

14

10 0 0.1

0.2

0.5

1

ZnO Catalyst Loading (g/L)

Fig. 8.9 (a) Effect of zinc oxide(ZnO) catalyst loading on dye removal in hybrid system [Initial concentration ¼ 500 mg/L]. (b) Percentage decolorization of dye in hybrid system at different zinc oxide(ZnO) catalyst loadings [Initial concentration ¼ 500 mg/L]

initial dye concentration and zinc oxide catalyst loading were fixed as 500 mg/L and 0.2 g/L, respectively, for all the pH experiments. Figure 8.10a depicts the effect of different solution pH on dye decolorization in the integrated process. It has been observed that decolorization of dye has been achieved up to 69, 85.98, 94, 87.5, and 75.6% at pH 1, 3, 5, 7, and 10, respectively, in the combined process of photocatalysis and nanofiltration. From Fig. 8.10b, it is observed that with an increase in pH to 5, the decolorization of dye has been increased gradually, whereas decolorization and mineralization rate of organic carbon has been decreased after pH 5. From the obtained results, it has been found that pH 5.0 is an optimum value to achieve desired results. This could be due to the

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a

pH=1

pH=3

pH=5

pH=10

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pH=7

600 500 400 300 200 100 0 0

b

15

30

45

60

75

90

Time (min) 100 90

% Removal of MB dye

80 70 60 50 40 30 20 10 0 0

3

6

9

12

pH

Fig. 8.10 (a) Effect of initial pH of dye solution on removal of dye in hybrid system [Initial concentration ¼ 500 mg/L, Catalyst loading ¼ 0.2 g/L]. (b) Percentage decolorization of dye in hybrid system at various pH values [Initial concentration ¼ 500 mg/L, Catalyst loading ¼ 0.2 g/L]

adsorption of dye on the catalyst surface which is dependent on its surface charge. At pH ¼ 5, the zinc oxide surface might be negatively charged due to the generation of massive hydroxyl radicals, which facilitates the adsorption of dye as well as its selfsensitized degradation (Lu et al. 2009). Moreover, the solution pH >5 could have imparted negative charge to zinc oxide surface, consequently inhibiting dye degradation by lower adsorption of dye molecules onto zinc oxide surface. It is reported that the production of hydroxyl radicals is favored at high pH (Pardeshi and Patil 2009). However, the hydroxyl ions can be also generated by scavenging the

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photogenerated holes on the surface of zinc oxide particles. These hydroxyl ions compete with direct oxidative degradation of dye by holes in the valence band of zinc oxide. The performance of individual systems, nanofiltration, photocatalysis, and their combination were successfully investigated. From the results, it is observed that the preliminary treatment of dye wastewater with photocatalysis and secondary treatment with nanofiltration was effective. At 0.2 g/L of zinc oxide catalyst loading and pH 5.26, the maximum dye removal was achieved. These are considered as the optimum operating conditions. Also, it is observed that on increasing the initial dye concentration, the decolorization was decreased. And in the combined system, 70% of total organic carbon was removed in 94.23 min of operational time.

8.5.6

Hybrid Hydrogel-Based System for Removal of the Dyes Using Hydrodynamic Cavitation Process

Aman et al. reported an innovative attempt for the removal of the dye using hydrogel and hydrodynamic cavitation-based processes (Raj et al. 2018). In this case study, the authors have synthesized the acrylamide gel using ultrasound-assisted in situ cavitation-based process. The hybrid gel was prepared using the addition of titanium dioxide nanoparticles during the polymerization process. The packed bed was prepared using nanocomposite hydrogel. Initially, swelling studies of the hydrogel were carried out. It is found that the swelling of the hydrogel can reach up to 800%. A single 1 mm orifice plate was used for the hydrodynamic cavitation process. The barriers of inlet pressure were carried out in the range of 1–10 bar pressure. The optimum pressure was 3 bar pressure for the degradation of the crystal violet dye. Effect of loading of titanium dioxide into the hydrogel was also carried out; it is found that loading of 0.5% titanium dioxide gives the maximum removal of the dye using nanocomposite hydrogel. Figure 8.11 shows the comparison of the individual system performance with a hybrid hydrogel system. The total organic carbon reduction for the individual system is 19% for hydrodynamic cavitation, and around 26% reduction total organic carbon is obtained using the hydrogel. Due to hydrodynamic cavitation, there is the fragmentation of the organic molecule due to high impingement during the cavitation process. While using hydrogel, the small fragmented molecules will be adsorbed into the polymer hydrogel. Using the combination of the hydrodynamic cavitation and hydrogel gives the removal around 70%. The hybrid system takes less than 90 min.

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350

TOC (mg/L)

300 250 200 150 100

CV+HC (3 bar)+PAM(TiO2 0.5%) CV+PAM+TiO2 0.5%) cv+HC (3 bar)

50 0 0

15

30

45 Time (min)

60

75

90

Fig. 8.11 Performance of hybrid system as compared to individual effect of hydrodynamic cavitation (HC) and poly-acrylamide (PAM)-titanium dioxide(TiO2)

8.6

Future Prospectus

In the investigation, it is found that individual hydrodynamic cavitation will not be going to degrade the reclaimed pollutant. Complete mineralization is possible by using the hydrodynamic cavitation and other additional techniques such as photocatalysis or the Fenton process. However, using photocatalysis it needs to develop the visible light catalysis to cover the complete spectrum of the sun. Recovery of the photocatalysis and photoreactors is the next need for scaling up to the industrial scale. The cost of the hybrid system is one of the important barriers to scale up; hence it needs to have the development of a hybrid system that can give complete mineralization with a smaller fixed cost and operating cost. In place of use of the polymeric membrane for the nanofiltration system, it needs to develop the ceramic membrane which can be utilized to reduce the total dissolved solids of the water along with chemical oxygen demand of wastewater. There is large scope to make treatment of the wastewater and reuse the same water for utilities; this can be achieved using the biofiltration system combine with hydrodynamic cavitation and ceramic membrane.

8.7

Conclusion

The bismuth loaded with titanium oxide photocatalytic system is the best system which can give maximum degradation and covers the better solar spectrum. Optimization of the inlet pressure of the hydrodynamic cavitation varies from 3 to 10 bars, which also depends on the presence of the pollutant type. The degradation of the dye’s pollutants also depends on the pH. Most of the dyes degrade at the acidic pH. There is also a need to add the additional hydrogen peroxide to generate the

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radicals, which helps to degrade the dye components; optimization is also needed to carry out in case of the hydrogen peroxide addition. At the higher level of the hydrogen, peroxide addition acts as the radical scavengers. The support of graphene helps in non-leaching of the catalyst in case of photocatalytic process.

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Sivudu KS, Rhee KY (2009) Preparation and characterization of pH-responsive hydrogel magnetite nanocomposite. Colloids Surf A: Physicochem Eng Asp (Elsevier) 349(1–3):29–34. https://doi. org/10.1016/j.colsurfa.2009.07.048 Suarez-Parra R et al (2003) Visible light-induced degradation of blue textile azo dye on TiO2/CdOZnO coupled nanoporous films. Sol Energy Material & Sol Cell (Elsevier) 76(2):189–199. https://doi.org/10.1016/S0927-0248(02)00346-X Thanekar P, Panda M, Gogate PR (2018) Degradation of carbamazepine using hydrodynamic cavitation combined with advanced oxidation processes’, Ultrasonics sonochemistry, vol 40. Elsevier, pp 567–576. https://doi.org/10.1016/j.ultsonch.2017.08.001 Wang J et al (2009) Photocatalytic degradation of organic dyes with Er3+: YAlO3/ZnO composite under solar light. Sol Energy Material & Sol Cell (Elsevier) 93(3):355–361. https://doi.org/10. 1016/j.solmat.2008.11.017 Wei R et al (2014) Glutatione modified ultrathin SnS2 nanosheets with highly photocatalytic activity for wastewater treatment. Mater Res Express (IOP Publishing) 1(2):25018. https://doi. org/10.1088/2053-1591/1/2/025018 Yu JC et al (2005) Efficient visible-light-induced photocatalytic disinfection on sulfur-doped nanocrystalline titania. Environ Sci & Technol (ACS Publications) 39(4):1175–1179. https:// doi.org/10.1021/es035374h Zhou Q et al (2017) Degradation kinetics of sodium alginate via sono-Fenton, photo-Fenton and sono-photo-Fenton methods in the presence of TiO2 nanoparticles. Polym Degrad & Stab (Elsevier) 135:111–120. https://doi.org/10.1016/j.polymdegradstab.2016.11.012

Chapter 9

Removal of Heavy Metals in Biofiltration Systems Andreas Aditya Hermawan

, Amin Talei

, and Babak Salamatinia

Contents 9.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.2 Biofiltration Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.3 Filter Media of Biofiltration Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.3.1 Sandy Loam . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.3.2 Loamy Sand . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.3.3 Fly Ash . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.3.4 Zeolite . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.3.5 Halloysite Nanotubes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.3.6 Biochar . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.3.7 Other Materials Used as Either Filter Media or an Additive . . . . . . . . . . . . . . . . . . . . . 9.4 Impact of Infiltration Rate on Heavy Metals Removal Efficiency in Biofiltration System . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.5 Accumulation of Heavy Metals in Biofiltration Media . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.6 Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

244 246 247 248 248 249 250 251 252 253 254 254 255 256

Abstract Stormwater generated in urbanized areas is generally contaminated with different pollutants such as heavy metals, nutrients, and suspended solids. One of the successful infrastructures in managing stormwater quality in cities is biofiltration system. Biofilters which consist of a layered filter media and vegetation are found successful in removing different classes of pollutants. Filter media of such systems is found to be the main contributor in removing heavy metals; however, its efficiency is

A. A. Hermawan · A. Talei (*) Discipline of Civil Engineering, School of Engineering, Monash University Malaysia, Subang Jaya, Selangor, Malaysia e-mail: [email protected]; [email protected] B. Salamatinia Discipline of Chemical Engineering, School of Engineering, Monash University Malaysia, Subang Jaya, Selangor, Malaysia e-mail: [email protected] © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature Switzerland AG 2021 Inamuddin et al. (eds.), Water Pollution and Remediation: Heavy Metals, Environmental Chemistry for a Sustainable World 53, https://doi.org/10.1007/978-3-030-52421-0_9

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highly dependent on material composition and size distribution. This study is aimed to review and discuss different types of filter media and their performances in removing different heavy metals and also the practical limitations of using such material in biofiltration systems. Moreover, other influencing factors in selecting the filter media including infiltration rate, metals accumulation in soil, and the shortterm and long-term functionality of the system are discussed. Keywords Biofiltration system · Stormwater management · Heavy metal removal · Filter media

9.1

Introduction

Stormwater is the surface runoff generated by rainfall which enters natural drainage systems like streams and rivers (Adams and Dove 1984). Stormwater contains pollutant substances from many sources such as roads, parking lots, and commercial areas. Heavy metals, as one of the major pollutant groups, are sourced from automobile fluid leaks and tires, abrasion process of metal components of engines, paints in buildings and other infrastructures, as well as atmospheric deposition (Davis et al. 2003). Some of the most common heavy metals traced in stormwater are zinc (Zn), lead (Pb), chromium (Cr), manganese (Mn), copper (Cu), arsenic (As), nickel (Ni), iron (Fe), and cadmium (Cd). Table 9.1 shows the average concentration of heavy metals in surface water at different urban densities all over the world (Duncan 1999). In general, it can be observed that highly urbanized areas produce higher pollutant concentration (except for Cd and Fe) compared to the less urbanized ones due to the intensified human activities such as construction and transportation. In addition, Table 9.1 also provides the surface water standards from the United States (Environmental Protection Agency 2017), Malaysia (DOE 2006), and China (China Water Risk 1996). As it can be seen, the overall stormwater quality in selected urbanized sites of this study does not comply with the presented standards and may harm the environment when joining water bodies such as rivers, ponds, lakes, or oceans. Therefore, stormwater quality management becomes a necessity to control the water quality in receiving water bodies. In the developed and developing countries, stormwater management has been one of the major research topics. Conventional stormwater management such as implementation of gutter–pipe system has been used for many years to discharge stormwater runoff to the nearest receiving water (Houle et al. 2013). In such conventional practice, the aim is to remove the stormwater from a site as fast as possible to mitigate on-site flooding. However, some researchers have proven that such practice is devastating for the downstream water bodies due to the increased frequency and magnitude of discharge, alteration of stream channel morphology, and more groundwater pollution (Jennings and Jarnagin 2002). Additionally, these

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Table 9.1 Heavy metals concentration in world average surface water based on urban density in comparison with surface water standards from the United States, Malaysia, and China

Heavy metals Pba Zna

High urban (μg/L) 146.00 240.00

Medium urban (μg/L) 104.00 75.00

Low urban (μg/L) 44.00 197.00

Standards United States (μg/L)
8.5
86.0
3.7

Cua

50.00

Cda Cra Nia Fea

4.00 25.00 31.00 2735.00

10.00 19.00 – 3908.00

10.00 19.00 25.00 5546.00


8.8
50.0
8.3
300.0

Mna Hga

236.00 0.20

160.00 0.03

100.00 0.05

–
0.025

32.00

37.00

Malaysia (μg/L) –
5000.0
20.0
10.0
50.0
50.0
1000.0
100.0
1.0

China (μg/L)
50.0
1000.0
1000.0
5.0
50.0 –
500.0
100.0
0.1

a

Zn zinc, Pb lead, Cd cadmium, Cr chromium, Mn manganese, Cu copper, Fe iron, Hg mercury, Ni nickel

methods do not cater for the stormwater quality and may result in pollutants buildup in the long run. In a study conducted by McCarthy (2008), authors found that the efficient method for managing stormwater quantity is to focus on reducing the amount of generated stormwater. This can be achieved by working on the on-site stormwater management techniques. Due to the increased rate of urbanization, especially in developing countries, conventional stormwater management is no longer sustainable. To date, stormwater treatment has gone through several improvements. By changing the perception of managing stormwater runoff as nuisance that must be gotten rid of to a utilizable resource, several solutions have been developed to capture stormwater and improve its quality while returning it to the water resources. Water-sensitive urban design as one of the design philosophies in urban water design aims to minimize the hydrological impacts of urban development on the surrounding environment. Water-sensitive urban design has been practiced in many countries including Australia since the 1990s (Lloyd et al. 2002; Ahammed 2017). Similar practices were developed in different countries such as low-impact development in the United States and sustainable drainage systems in the United Kingdom. There are several technologies that have been comprehensively studied to address the problems in urban drainage systems. The examples of such solutions are wetlands, detention ponds, green roofs, and biofiltration systems. Biofiltration systems mainly consist of a filter media and vegetation which are built on the ground and target stormwater management for both quality and quantity aspects (Davis et al. 2009). This review study is focused on the heavy metals removal from stormwater runoff by the filter media of biofiltration systems. Several types of filter media are

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compared and discussed based on their characteristics such as composition, particle size, and removal capacity. In addition, the effect of infiltration rate of filter media in biofiltration system on their heavy metals removal performance is also presented.

9.2

Biofiltration Systems

Biofiltration systems are one of the technologies that can be classified under the best management practices. These systems are generally constructed as the landscape depressions for an on-site treatment of stormwater. Biofiltration system is potentially able to offer improvement in managing stormwater by reducing the runoff volume and enhancing the water quality (Hatt et al. 2009; Hermawan et al. 2019). Biofiltration systems comprise two major components: filter media and vegetation as shown in Fig. 9.1. The filter media can trap heavy metals by ion-exchange mechanism. In addition, it can work as a sediment trap and remove the suspended solids from infiltrated water. On the other hand, vegetation mainly contributes to nutrient removal as well as prolonging the lifespan of biofiltration system by mitigating clogging potential. Numerous lab-scale and field-scale studies have been carried out in the past few decades to evaluate the efficiency of biofiltration systems using various types of filter media, vegetation, and climate (Davis et al. 2003; Dietz and Clausen 2005; Hsieh and Davis 2005; Hatt et al. 2008). Section 9.2 of this chapter reviews the studies conducted on the filter media of biofiltration systems with a focus on heavy metals removal from stormwater runoff.

Fig. 9.1 Cross section view of a typical biofiltration system consisting of three layers (drainage, transition, filter), ponding zone, and vegetation on top. A perforated pipe is installed in the drainage layer to allow outflow of the filtered water. (Image is modified after FAWB 2009)

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9.3

247

Filter Media of Biofiltration Systems

Many studies on biofiltration systems have been focused on the optimization of the layered filter media structure. The common practice is a three-layer media which starts from a drainage layer at the bottom, followed by transition layer in the middle and a filter media layer on top, in which the roots of the plants penetrate as shown in Fig. 9.1. Drainage layer consists of big aggregates such as gravel and crushed rocks, whereas the filter media consists of sand with different particle sizes ranging from very fine to very coarse (Hsieh and Davis 2005). In order to prevent the sand from being washed into the drainage layer, an additional layer called transition layer is placed in between. Transition layer has a particle size between drainage and filter media and is working as a bridge between the two layers (FAWB 2009). Payne et al. (2015) suggested that the design of the layered media should follow the general filter criteria as shown in Eqs. (9.1) and (9.2). Considering these criteria, the segregation of fine particles from filter layer to drainage layer is expected to be minimized, thus reducing the clogging potential in the filter media. D15 ðTransitionÞ
D85 ðFilterÞ

ð9:1Þ

D15 ðDrainageÞ
D85 ðTransitionÞ

ð9:2Þ

*

D15 means that 15% of the materials contained have particle diameter less than this value ** D85 means that 85% of the materials contained have particle diameter less than this value In order to meet the required water quality criteria, various types of filter media have been investigated in the structure of biofiltration systems. For enhancing the heavy metals’ treatment efficiency in stormwater, fine-grained materials can be added to the filter media. Although activated carbon, as a conventional adsorbent material, has been known for its ability to remove heavy metals, it is costly and in need of a long chemical process in its production. Due to the demand of moving toward sustainable environment, natural adsorbents have gained more popularity in this application over the last few years (Asubiojo and Ajelabi 2010). Natural adsorbents are generally considered as low cost, yet providing similar effectiveness in removing heavy metals compared with activated carbon. The examples of such natural adsorbents in biofiltration systems are sand, fly ash, zeolite, biochar, and halloysite. In most applications, sand is the main component, while other materials are added to enhance the metal ion removal process. In the following subsections, some of the practiced filter compositions are reviewed and discussed.

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Sandy Loam

Sandy loam is a soil mixture that contains 7% clay or less, 50% silt or less, and 43–52% sand (Cosby et al. 1984). It can be categorized further depending on the coarseness of the materials. In biofiltration system application, sandy loam is one of the most popular filter media due to the balance between silt and sand materials. Sandy loam is commonly used in biofiltration system that does not treat high volume of water since the high percentage of clay and silt material in this mix would hinder the water infiltration rate in it. However, lower infiltration rate of this mix is an advantage in removing heavy metal ions due to the longer retention time. Davis et al. (2003) conducted a study on heavy metal removal capability of sandy loam used in a biofiltration system. In this study, the sandy loam filter media was topped with mulch and was vegetated with North America native plant Juniperus horizontalis. The authors reported that the biofiltration system can remove Cu(II), Pb(II), and Zn (II) ions up to 99% (% of removal rate). Wang et al. (2017) designed a biofiltration system column study with sandy loam media mixed with 10% of lignin. The system could remove 95–100% of heavy metals including Cu(II), Pb(II), Cd(II), and Zn(II). Overall, application of sandy loam as the filter media of biofiltration system has been found effective in removing heavy metals.

9.3.2

Loamy Sand

Loamy sand is commonly found with composition of more than 85% sand and less than 15% clay and silt. However, there are few types of loamy sand with sand content of 70–85% and clay–silt content of maximum 30%. Due to the higher sand content in comparison with sandy loam, biofilters with loamy sand media are supposed to be more capable of handling higher volume of water due to their higher hydraulic conductivity. In fact, few of the design guidelines for biofiltration systems such as Adoption Guidelines for Stormwater Biofiltration Systems (FAWB 2009) and Stormwater Management Manual for Malaysia (MSMA 2012) have recommended the use of loamy sand as the filter media. The recommended particle size distribution for loamy sand media according to the aforementioned guidelines is shown in Table 9.2. Lim et al. (2015) studied the performance of loamy sand as the biofiltration filter media. The loamy sand used in the study was consisted of 75.3% sand, 18.3% silt, and 2.9% clay. The results showed that reduction of Zn(II), Cd(II), and Pb(II) ions in stormwater can be up to 98%, while for Cu(II) the removal rate was in the range of 82–98%. Overall, the authors recommended the loamy sand as a proper filter media in biofiltration systems. Most recently, Hermawan et al. (2018) conducted a soilcolumn study to customize biofiltration systems for tropical conditions. In this study, the percentage of fine material reduced to 7% (all in silt category with no clay), while the main component of the filter media (93%) was consisted of sand. It was found

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Table 9.2 Recommended particle size distribution of loamy sand for biofiltration system filter media Percentage (%)

Material Clay and silt Very fine sand Fine sand Medium to coarse sand Coarse sand Fine gravel Organic materials

Size (mm) < 0.05 0.05–0.15

Facility for Advancing Water Biofiltration (FAWB 2009) 420 nm)

500 W Xe (λ > 420 nm)

250 W Xe (λ > 420 nm)

300 W Xe (λ > 420 nm) 300 W Xe (λ > 420 nm)

250 W Xe (λ > 420 nm)

Fluorescent 15 W

Lamp 300 W Xe (λ > 420 nm)

Zhang et al. (2018) Qu et al. (2017)

Zhang et al. (2011) Qin et al. (2017) Zhang et al. (2016) Zhang et al. (2012a, b) Liu et al. (2015)

Liu et al. (2014)

References Tu et al. (2016)

Table 11.4 SnS2 nanoparticles and heterostructures synthesized for Cr(VI) reduction and the main photocatalytic parameters. SPNH–MOSF@SnS2 spirobenzopyran–macroporous ordered siliceous foam@tin(IV) sulfide

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Fig. 11.12 SnS2/reduced graphene oxide (figure a) and spirobenzopyran–macroporous ordered siliceous foams SPNH–MOSF@SnS2 (figure b) heterostructures and their corresponding photocatalytic plots. Figure 11.12a: reprinted with permission of Springer Nature, Journal of Materials Science, from (Liu et al. 2015). Figure 11.12b: reprinted with permission of Elsevier, Applied Catalysis B: Environmental, from (Qu et al. 2017)

For the mentioned heterostructure, the mechanism can be described as follows. After excitation of SnS2 electrons, they are transferred to the reduced graphene oxide that acts as an electron harvesting center. The charge carrier is rapidly separated because the transfer is rapid, and they migrate to the surface to participate in the reduction/oxidation reactions. An illustration of the mechanism is depicted in Fig. 11.13 (Liu et al. 2015). A greener alternative uses carbonized moss decorated with SnS2 nanoparticles. The C-moss has the particularity of having a mountain shape structure, increasing its surface area ratio and enabling higher adsorption of heavy metals. Xia Zhang et al. reported on a SnS2@carbonized moss heterostructure with high capacity to reduce Cr(VI) (Zhang et al. 2018). In 70 min SnS2@C-moss reduces 98.5% of the total Cr (VI), while tin(IV) sulfide reduces 96.4% in 120 min. The carbonized moss contributes to the heterostructure with higher surface area and conductivity characteristics, inhibiting the recombination of electrons and holes. This contribution improves the photocatalytic activity. This novel heterostructure was tested in three complete photocatalytic cycles, remaining almost unchanged.

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Fig. 11.13 Schematic illustration of the process involved in the photocatalytic reduction of Cr (VI) using SnS2/rGO heterostructures. CB conduction band, VB valence band, Eg energy band gap, hν Planck constant*light frequency. (Modified after Zhang et al. 2012a, b)

A multifunctional heterostructure of SnS2 and macroporous ordered siliceous foams (MOSF) was developed by Jiafu Qu et al. (2017). The design of this composite included the functionalization with spirobenzopyran derivative (SPNH) that quelates the Cr(III) product of Cr(VI) reduction and the use of MOSF as adsorbent and catalysts support. It was proved that complete reduction of Cr (VI) of a 50 mg/L solution was achieved in about 85 min under visible light irradiation. The quelating properties of spirobenzopyran are activated by ultraviolet light, when a phenoxy group is generated by a ring opening. An image of the composite, together with the photocatalytic reduction efficiency of 50 mg/L spirobenzopyran–macroporous ordered siliceous foams@SnS2 at different initial concentration of Cr (K2Cr2O7), can be observed in Fig. 11.12b. A scheme of complete process (reduction and removal) of spirobenzopyran–macroporous ordered siliceous foams@SnS2 photocatalysis is shown in Fig. 11.14. Independently of the SnS2 nanoparticle morphology, all materials presented excellent photocatalytic activity. In all the mentioned reports, a minimum of 90% Cr(VI) reduction rate was reached in less than 150 min. Also, material recyclability was an essential factor that was successfully proved. There is a tendency of heterostructures to be more efficient in shorter times. Moreover, a combination of the reduction agent (heterostructure) together with chelating molecules seems to be a very promising alternative for the removal of Cr from water resources.

11.4.3 Tin-Based Compounds as Antimicrobial Agents Bacterial infections due to the consumption of contaminated water are one of the leading causes of disease in humans. The inhibition of microbes is necessary to revert this situation and to guarantee the quality of water. This goal can be achieved

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Fig. 11.14 Schematic illustration of the reduction and removal of Cr(VI) process of SPNH– MOSF@SnS2 (spirobenzopyran–macroporous ordered siliceous foams@SnS2). (Reprinted with permission of Elsevier Applied Catalysis B: Environmental from Qu et al. 2017)

with inorganic or organic antimicrobial agents, being the inorganic reagents more stable at high temperatures and pressure. Tin compounds have been widely studied for their antimicrobial properties, since they generally have favorable qualities for this purpose, such as low toxicity, high stability, eco-friendly, and straightforward synthesis, among others (Fakhri et al. 2015; Vasanthi et al. 2013; Prabha et al. 2017). As was described, tin-based compounds are excellent photocatalysts for metal and dye water remediation. Taking advantage of these two properties (antimicrobial and photocatalysts), tin compounds are very promising for the effective treatment of contaminated water. For this reason, in the following paragraphs, we describe the advances found in literature about antimicrobial tests of tin-based materials, mainly of SnO2 compounds. Designing metal-based agents to kill pathogens requires understanding the antimicrobial action and the complexity of metal biochemistry (Lemire et al. 2013). As the mechanism may vary according to the chemical environment of tin, we classify the compounds into three groups to describe the antimicrobial mechanism: nanostructures of tin(IV) oxide (SnO2) and tin(IV) sulfide (SnS2), nanostructures of metal-doped SnO2 and SnS2, and heterostructures of SnO2.

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SnS2 and SnO2 nanoparticles can be obtained by different methods. Physical methods include ion sputtering and thermal evaporation, while chemical methods include sol–gel, surfactants mediated synthesis, reverse cell, and hydrothermal. The most common and cost-effective is the hydrothermal (Fakhri et al. 2015). The possibility of obtaining these compounds in the nanometer scale is advantageous concerning the antimicrobial activity. Smaller particles than the cell and cellular organelles can quickly enter the cell (Ahamed et al. 2018). In the literature, two are the proposed mechanisms for the microbial activity of SnS2 and SnO2: (i) reactive oxygen species generation and (ii) internalization of the nanoparticles in microorganisms’ cell membranes in contact with them (Prabha et al. 2017). The generation of reactive oxygen species and oxidative stress are the leading causes of nanoparticles toxicity. The formation of species, as O2•- (superoxide anion) and HO• (hydroxyl radicals) act in the cell, causing inflammation, cell cycle arrest, mutation, genetic damage, and apoptosis. Also, the reactive oxygen species can reduce the mitochondrial membrane potential (Ahamed et al. 2018). The visible and ultraviolet light irradiation of the nanoparticles increases the efficiency of ROS production. For this reason, in the dark, the antibacterial activities are weaker than with irradiation of light (Fakhri et al. 2015). Figure 11.15 shows a representation of the reactive oxygen species generation under light irradiation and in darkness. Then, it is expected that the antimicrobial activity under light irradiation is potentiated. The studies carried out with SnO2 suggested that the accumulation of the nanoparticles on the surface of the bacterial cell membrane, due to chemical or physical interactions, induced the formation of “pits” on the bacteriological cell membranes and walls. They proved that the presence of the nanoparticles in the periplasm of the cell is capable of damaging the deoxyribonucleic acid (DNA) imitation ability, resulting in the incapacity of the expression of proteins and

Fig. 11.15 Diagram of mechanisms proposed for antimicrobial activity of tin chalcogenides nanoparticles. ROS reactive oxygen species

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enzymes vital for the production of adenosine triphosphate (ATP) (Khan et al. 2018) (Fig. 11.15). On the other hand, the primary mechanisms for the tin(IV) oxide and sulfide antimicrobial activity are the generation of reactive oxygen species and the damage of these species in the macromolecules of the microbial cells. Alternative materials were tested to enhance antibacterial activity. Metal doping of SnO2 nanoparticles could be an efficient way to modulate the antibacterial activity. For example, Sn1-xSbxO2 nanoparticles (0  x  0.1) were developed for this purpose. In this case, the enhanced antibacterial activity was attributed to the larger surface area (a smaller particle size) and the higher rate of generated ROS compared to SnO2 (Mazloom et al. 2017). Similar effects were found for other examples such as Zr-doped SnO2 nanopowders (Manjula and Selvan 2017), Co-doped SnO2 nanoparticles (Nasir et al. 2017), and Ag–SnO2 nanocomposites (Kumar Nair et al. 2018). However, in other materials, such as Cu/SnO2 nanocomposite bilayer coatings, the lethal effect of the Cu is more predominant than that of the ROS action (Fukumura et al. 2018). Other tin-based heterostructures were developed, such as graphene oxide, carbon nanotubes (Pandiyan et al. 2019), and calcium phosphate (Cui et al. 2019) doped with SnO2. In summary, the antimicrobial activity of tin compounds is associated to the increased reactive oxygen species generation for undoped and metal-doped tin chalcogenides and is shown in Fig. 11.16. The antimicrobial activity is usually tested in bacteria and fungi. The most common used pathogens and the tin compound used as an antibacterial agent are listed in Table 11.5. The assays showed that bacteria that have a better surround by complex and multilayer wall are more resistant to antibacterial agents. Gram-negative bacteria have a thin layer of peptidoglycan that is surrounded by an outer membrane of lipopolysaccharides. Cell wall of Gram-positive bacteria is composed of peptidoglycan layers. Thus, Gram-negative bacteria are more susceptible to the antibacterial activity mediated by reactive oxygen species generation (Mazloom et al. 2017).

11.5

Perspectives of Tin-Based Compounds for Water Remediation

Tin compounds have been widely studied and proved to be efficient for water remediation, including organic dyes, metals, and pathogens. Although they are not still in a commercial stage, they could be an exciting alternative to systems that are more complicated and expensive, due to the simplicity of preparation, low cost, versatility, and absorption in the visible region.

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Fig. 11.16 Diagram that summarize the mechanisms of antimicrobial action of tin compounds. ROS reactive oxygen species

Further research must be performed before we see tin-based compounds as part of water remediation equipment, but researchers are getting close to it. Current efforts are focused on the rational design of heterostructures, since this approach leads to a noticeably increased activity. The enhanced properties of these new materials envisage the possibility of including them in commercial water remediation systems in the near future.

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Table 11.5 Summary of tin-based materials with proven antimicrobial activity and the microorganisms used in the tests. Except Bacillus subtilis and Puccinia graminis, all the microorganisms listed are human pathogens Microorganism Bacillus subtilis (G+)

Staphylococcus aureus (G+)

Tin material SnS2 NPs Green synthesized undoped SnO2 and Co-doped SnO2 NPs Co-doped SnO2 NPs Zr-doped SnS2 nanopowders SnS2 NPs SnO2, SnS2, SnO2/SnS2 nanostructure SnO2 NPs SnO2 nanorods Heterostructure of SnO2–ZnO

Escherichia coli (G
)

Sn1-xSbxO2 NPs SnS2 NPs SnO2, SnS2, SnO2/SnS2 nanostructure SnO2 nanorods Green synthesized undoped SnO2 and Co-doped SnO2 NPs ZnS quantum dots decorated SnO2 nanosheets Ag–SnO2 nanocomposites Heterostructure of SnO2–ZnO

Pseudomonas aeruginosa (G
) Proteus vulgaris (G
) Klebsiella pneumoniae (G
) Puccinia graminis Candida albicans

Sn1-xSbxO2 nanoparticles SnO2 doped GO and CNT SnO2-doped calcium phosphate coating Cu/SnO2 nanocomposite bilayer coatings SnO2 nanoparticles Sn1-xSbxO2 nanoparticles SnO2 nanoparticles Sr-doped SnS2 nanopowders Zr-doped SnO2 nanopowders SnO2 doped GO and CNT SnO2, SnS2, SnO2/SnS2 nanostructure

References Khimani et al. (2019) Khan et al. (2018) Nasir et al. (2017) Srivind et al. (2017) Khimani et al. (2019) Fakhri et al. (2015) Kumar et al. (2017) and Phukan et al. (2017) Díez-Pascual and DíezVicente (2017) Sudhaparimala and Vaishnavi (2016) Mazloom et al. (2017) Khimani et al. (2019) Fakhri et al. (2015) Díez-Pascual and DíezVicente (2017) Khan et al. (2018) Hosseini et al. (2018) Kumar Nair et al. (2018) Sudhaparimala and Vaishnavi (2016) Mazloom et al. (2017) Pandiyan et al. (2019) Cui et al. (2019) Fukumura et al. (2018) Kumar et al. (2017) Mazloom et al. (2017) Phukan et al. (2017) Prabha et al. (2017) Manjula et al. (2017) Pandiyan et al. (2019) Fakhri et al. (2015)

NPs nanoparticles, GO reduced graphene, CNT carbon nanotubes, G+ positive Gram, G
negative Gram

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Chapter 12

Some Effective Methods for Treatment of Wastewater from Cu Production Vesna Krstić

Contents 12.1 12.2

12.3

12.4

12.5 12.6

12.7

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.1.1 Primary and Secondary Copper Production . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Metal Toxicity, Its Biological Activity, and Factors Affecting the Removal of Toxic Metals from Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.2.1 Behavior of Heavy Metals in the Aquatic Environment . . . . . . . . . . . . . . . . . . . . . . 12.2.2 Metal Properties and Factors That Influence Their Removal from the Water Methods for Purifying Wastewater from Toxic Metals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.3.1 Adsorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.3.2 Biosorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.3.3 Low-Cost Adsorbents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.3.4 Cu and Modified Mesoporous Carbon and X-Ray Photoelectron Spectroscopy (XPS) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Different Types of Nanomaterials for Wastewater Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . 12.4.1 Nano-adsorbents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.4.2 Nanomaterials as Catalysts and Electrocatalysts . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.4.3 Nanomembranes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.4.4 Combination of Biological Nanotechnological Processes . . . . . . . . . . . . . . . . . . . . . 12.4.5 Discussion and Future Perspectives . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Photocatalysis as a Method for Wastewater Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . The Removal of Heavy Metal from Wastewater Using Nano Zero-Valent Iron (nZVI) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.6.1 Structure and Synthesis of Nano Zero-Valent Iron (nZVI) . . . . . . . . . . . . . . . . . . . . 12.6.2 In Situ Application of Nanomaterials for Decontamination of Groundwater . 12.6.3 “Green” Synthesis of Nano Zero-Valent Iron (nZVI) . . . . . . . . . . . . . . . . . . . . . . . . . Electrolysis with Some Dimensionally Stable Anodes (DSA) Based on Ti . . . . . . . . . . . . 12.7.1 General Characteristics of Titanium (Ti) and Dimensionally Stable Anode (DSA) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.7.2 Catalyst Surface Characterization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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V. Krstić (*) Mining and Metallurgy Institute Bor, Bor, Republic of Serbia Technical Faculty, University of Belgrade, Bor, Republic of Serbia e-mail: [email protected] © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature Switzerland AG 2021 Inamuddin et al. (eds.), Water Pollution and Remediation: Heavy Metals, Environmental Chemistry for a Sustainable World 53, https://doi.org/10.1007/978-3-030-52421-0_12

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Application of Different Catalytic Activated Titanium Mixed Metal Oxide (MMO) Anodes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.7.4 Discussion and Further Perspectives . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.8 Wastewater Treatment Using Phytoremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.8.1 Rhizofiltration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.8.2 Aquatic Ecosystems for Removal of Pollutants from Wastewater . . . . . . . . . . . . 12.9 Conclusions and Future Perspectives . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract The total global production of refined copper in 2017 was approximately 19 million tons, with an annual growth rate of 3.4%. During the copper production process, a large proportion of the accompanying toxic metals end up in the environment. For this reason, there is a significant need for advanced wastewater treatment methods and technologies in order to ensure optimal water quality, eliminate heavy metals and other pollutants from water, and suggest appropriate industrial technology for the treatment of wastewater. Although various techniques for treatment of wastewater contaminated with heavy metals are being applied today, the choice of the most suitable wastewater treatment process depends on some basic commonly accepted parameters which will be discussed in this paper. The methods and techniques such as adsorption on the new sorbents (biosorbents, agricultural and industrial wastes (lignocellulosic materials) as an ecological adsorbent; nano-adsorbents, activated carbon, carbon nanotubes, graphene, MgO, MnO, ZnO, TiO2, Fe3O4, etc.), nanotechnology, photocatalysis, nano zero-valent iron (nZVI), the use of dimensionally stable anodes in electrolysis, and phytoremediation have proved to be adequate in the treatment of wastewater from the support in particular toxic metals such as copper (Cu), lead (Pb), cadmium (Cd), nickel (Ni), chromium (Cr), arsenic (As), zinc (Zn), and mercury (Hg), from primary and secondary copper production. Sorbents can be regenerated or concentrated by combustion and electrolysis using dimensionally stable anode; metals can be selectively separated and can be returned to the production process. Working principles and the advantages and disadvantages of the mentioned materials and methods for water remediation will be discussed in this paper. Due to their importance of the impact on the living world and on the environment, the toxicity of each of these polluting metals will also be demonstrated. The results show that water is generally polluted and that in the near future, we will have to take the most serious approach to addressing this problem. Great efforts are already being made to come up with the most efficient and inexpensive methods for wastewater treatment. This generally requires combining multiple methods for quality problem-solving, in accordance with the type and concentration of the pollution identified. In addition to engaging experts from the natural sciences, it is also necessary to include a management system and link up ministries of ecology at the state level and international level, in order to approach this problem more efficiently and to preserve rivers that flow through multiple lands and carry with them substances harmful to human health and to the environment and rivers which then flow with these substances into lakes, seas, and oceans.

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Keywords Toxicity and heavy metals · Wastewater treatment · Adsorption on new adsorbents · Nanotechnology · Photocatalysis · Nano zero-valent iron · Dimensionally stable anode · Phytoremediation

Abbreviations ANFIS BDST BE BET CNMs CS/SCNTs DSA DSE DLVO EDTA EDX ENMs EU FRAP FTIR GO HRAES IET MC MCL MMO MSC MWCNTs NA NF NMR NPs nZVI NT OEP OER PBI PES PET ppb PGMs PPTA PS

Adaptive neural fuzzy inference system Bed depth service time Binding energy Brunauer–Emmett–Teller Carbon nanomaterials Chitosan/silicon-coated carbon nanotubes Dimensionally stable anode Dimensionally stable electrodes Derjaguin, Landau, Verwey, and Overbeek theory Ethylenediaminetetraacetic acid Energy-dispersive X-ray Electrospun nanofiber membranes European Union Ferric reducing antioxidant power Fourier transform infrared spectroscopy Graphene oxide High-resolution Auger electron spectroscopy Isoelectric point Mesoporous carbon Maximum contaminant level Mixed metal oxide Modified mesoporous silica–carbon Multiwalled carbon nanotubes Not available Nanofiltration Nuclear magnetic resonance Nanoparticles Nano zero-valent iron Nanotube Oxygen evolution potential Oxygen evolution reaction Polybenzimidazole Polyethersulfone Polyethylene terephthalate Parts-per-billion Platinum group metals Para-phenylene terephthamide Polystyrene

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PSF PVC PVDF PVP RO SEM SHE SWCNTs TEM TGA UF US EPA UV WHO XPS XRD

12.1

Polysulfone Polyvinyl chloride Polyvinylidene fluoride Polyvinylphenol Reverse osmosis Scanning electron microscopy Standard hydrogen electrode Single-walled carbon nanotubes Transmission electron microscopy Thermogravimetric analysis Ultrafiltration United States Environmental Protection Agency Ultraviolet World Health Organization X-ray photoelectron spectroscopy X-ray diffraction

Introduction

Water on earth is one of the most important resources in the world, but only about 1% of this natural resource is used for human consumption. It is estimated that more than 1.1 billion people lack the supply of satisfying drinking water (Anjum et al. 2016; Williams et al. 1991), due to an increase in the cost of drinking water treatment, a higher population with access to supplies, and various climate and environmental changes (Adeleye et al. 2016; WHO 2017a, b). With rapid scientific and technological developments in all sectors of the economy, the emission of toxic microelements has increased, which in turn become significant environmental pollutants. Pollution of heavy metals and organic pollutants poses a serious threat to human health due to hematotoxicity (Lima et al. 2019; Barakat 2011). With the increasing emissions of heavy metals, their concentrations in sediments, soil, and water are increasing. Adequate treatment of wastewater and drinking water can reduce these problems; however traditional methods for their processing are not effective enough because they cannot completely eliminate the resulting impurities and meet stringent water quality standards (de Wet and Odume 2019; Ferroudj et al. 2013a). The removal of toxic metal ions from wastewater can be achieved by different physicochemical methods (precipitation, membrane processes, adsorption, electrochemical processes, ion exchange). These processes have significant disadvantages, for example, the incomplete removal of contaminants, a high energy consumption, and the production of contaminated sludge. Conventional technology does not give adequate results for wastewater treatment, because water purification is not complete

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Atmospheric water

Surface water

Effective precipitation

Water supply Soil water

Underground water

Virtual water transfer

Transferred water Water utilization system

Virtual water consumption system

Water utilization

Products trade

Reclamation and reuse Evaporation & Consumption

317

Products consumption Recyling

Virtual water export

Products in circulation

Visible Water

Virtual Water Embed

Fig. 12.1 Relationship between visible and virtual water for regional water resource management. (Reprinted with permission of Omega from Li et al. (2019))

and operating costs are high (Acar and Eren 2006; Gundersen and Stiennes 2003). New methods developed for removing ions of heavy metals from wastewater using green technology give better results (Gao et al. 2017; Ali et al. 2016; Guiza 2017; Wang et al. 2016; Ramteke and Gogate 2016; Lončar et al. 2019). More recently, a number of procedures have been examined for the development of cheaper and more efficient technologies, ways of reducing the amount of wastewater and improving the quality of treated effluents. Adsorption on new adsorbents is one of the alternative wastewater treatments, which in recent years has led to an interest in cheap and natural adsorbents that are not harmful to the environment (Lakhdhar et al. 2016; Ramteke and Gogate 2016). Adsorptive materials may be minerals, of biological origin, and zeolites, industrial by-products, agricultural waste, biomass, and polymeric materials (Kurniawan et al. 2005). All these are of interest. Membrane separations are increasingly used in the treatment of inorganic effluents for ease of use. There are various types of membrane filtration such as ultrafiltration (UF), nanofiltration (NF), and reverse osmosis (RO) (Cordier et al. 2019; Kurniawan et al. 2006). Electrical treatments such as electrolysis with dimensionally stable electrodes (DSE) (Krstić and Pešovski 2019) and electrodialysis (Vanoppen et al. 2019) have also contributed to the treatment of wastewater and environmental protection. Photocatalytic methods are a promising method in the reduction of environmental pollutants (Zhang et al. 2019; Ferroudj et al. 2013a). Based on the nature of the nanomaterials, there are nano-adsorbents, nanocatalysts, and nanomembranes, and also the integration of these nanotechnologies with biological methods is also mentioned (Kyzas and Matis 2015; Gupta et al. 2015a, b). Li et al. (2019) based on the example of the Northwestern District of China, using their own methodology, give practical suggestions for the coordination between real and virtual waters as one of the ideas for regional water resource management (Fig. 12.1). Due to rapid industrialization, wastewater treatment has become one of the “burning” issues of today, which leads to the need to engage a large number of researchers in seeking solutions. Wastewater today should discharge into waterways

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purified to the level of legislation. Although many of the techniques used for the purification of wastewaters are being developed, the removal of heavy metals should allow not only technologically suitable for treatment to the particular working conditions but also the capacity to satisfy the maximum contaminant level (MCL) standards. The wastewater discharge criterion depends on the technological processes of a particular industry and on the allowed emissions. Significant financial resources are needed to solve this problem. Since the number of inhabitants on the planet is increasing rapidly, the need for water is increasing, while pollution is constantly generated, and in the near future, it is expected that humanity will face a serious lack of water for human use.

12.1.1 Primary and Secondary Copper Production The importance of using copper in industry today leads to a greater need for activities in this field, which leads to significant environmental problems. Today, copper can be obtained in two ways in the world, primarily by mining exploitation and secondly by recycling of industrial waste. Jingjing et al. (2019) analyzed these two processes from the aspect of energy consumption and pollution emissions, in order to better understand the overall effect on environmental protection. The results of their analysis show the most serious impacts of the polluting production of refined copper, which affects the pollution of the environment and is toxic to man and the environment globally. In the world today, two technologies for the production of copper from copper ore, pyrometallurgy and hydrometallurgy, are used. Pyrometallurgy is used for sulfide copper ores and hydrometallurgy for low-quality oxidized copper ores. China is today the biggest copper producer in the world and using pyrometallurgical technology receives more than 98% refined copper from sulfide ores (Wang et al. 2015a, b); however Chile is the world’s largest producer of mined copper. Figure 12.2 shows the chain of primary and secondary copper production (Jingjing et al. 2019). The environment is the most endangered by primary recovery of copper from mining activities and the melting of primary copper by pyrometallurgy. The main factors for obtaining copper by the secondary process (recycling of waste) are refining and electrolysis, which also pollute the environment. For this reason, particular attention should be paid to these two processes, and work should be focused on improving their technology, since copper production would thus be cleaner and pollution of the environment less harmful (Krstić and Pešovski 2019). Table 12.1 Jingjing et al. (2019) show the type of pollution that is most prevalent for the primary and for the secondary production of copper. According to their analysis results, the primary production of copper has about eight times more influence on the environment pollution than the secondary production of copper. Also, the analysis showed that electricity is the sensitive factor for both directions of copper production (Miah et al. 2017; Yu et al. 2014a, b). Krstić and Pešovski (2019) proposed the optimization of the use of energy in the production of copper, related to

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Fig. 12.2 System boundary. The chain of primary and secondary copper production. (Reprinted with permission of the Resources, Conservation and Recycling from Jingjing et al. (2019)) Table 12.1 Main pollutants in the copper production process Types Air pollutant Liquid pollutant

Solid pollutant

Primary copper Sulfur dioxide, particulate matter, flue gas containing arsenic (As), cadmium (Cd), sulfuric acid mist, etc. Electrolyte, contaminated acid containing fluorine, chlorine, arsenic, acidic wastewater containing heavy metals, washing water, etc. Dust, smelting water-quenched slag, beneficiation tailings, leaching slag, sewage treatment slag, etc.

Secondary copper Dioxins, sulfur dioxide, nitrogen oxides, etc. Electrolyte, contaminated liquid containing heavy metals such as copper (Cu), lead (Pb), tin (Sn), nickel (Ni), etc. Nonmetallic scrap, smelting slag, anode mud, dust, etc.

Reprinted with permission of Resources, Conservation and Recycling from Jingjing et al. (2019)

the use of dimensionally stable anode (based on titanium) in copper electrolysis, which can produce pure cathode copper and reduce the impact on environmental pollution. The same dimensionally stable anode can be used after a solvent extraction process for the purification of wastewater containing copper ions (Cu2+), which would be returned to the copper production process, and the wastewater would be released from copper ions (Cu2+), (Krstić and Pešovski 2017). Since the quality of life improves with time, people are increasingly concerned with the impact on environmental and human health. The results analyzed in Table 12.2 show that increasing the share of secondary copper plays an important role in protecting the environment and human health. In Table 12.2 for 2015, according to Jingjing et al. (2019), the share of copper obtained by secondary

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Table 12.2 Environmental impacts of different scenarios Scenario FUP FUS BS S1 S2

Primary copper ratio 100% 0 72.40% 65% 60%

Secondary copper ratio 0.00 100.00% 27.60% 35.00% 40.00%

Environmental impact 7.05E-09 8.76E-10 2.42E-10 4.89E-09 4.58E-09

Reprinted with permission of Resources, Conservation and Recycling from Jingjing et al. (2019) Symbols: FU, the functional unit; FUP, the primary scenario; FUS, the secondary scenario; BS, refined copper scenario; S1, S2, several combinations or scenarios

means was 27.60% of the total refined copper, while the impact of the functional unit (FU) of refined copper on the environment (BS scenario) should be calculated by 2.42E-10. In order to examine the environmental impact of recycling the secondary process of obtaining copper from the waste, several combinations or scenarios (S1, S2) were set up. Table 12.2 shows that when the share of secondary copper increases to 35% in scenario S1, the environmental impact will be reduced to 8.55% of the BS scenario. When in scenario S2 the share reaches 40%, the environmental impact will be reduced to 14.32% in relation to the BS scenario. These results show that increasing the share of secondary copper production has a positive effect on reducing the impact on the environment. The global pollution potential of the primary scenario (FUP) was ten times higher than the secondary scenario (FUS). According to the results analyzed, the recycling of copper waste (obtaining copper by a secondary process) may provide a better solution to obtaining copper than from the mine (by processing the ore) from the ecological aspect.

12.2

Metal Toxicity, Its Biological Activity, and Factors Affecting the Removal of Toxic Metals from Wastewater

12.2.1 Behavior of Heavy Metals in the Aquatic Environment It is important to ascertain the global cycle of heavy metals in nature, with the intention of determining and taking note of the results obtained. Solubility in the soil depends on the pH of the medium, redox potential and then on its primary composition and mineral composition, the concentration of inorganic compounds, the amount and type of organic compounds in soils and soil solutions, temperature, pressure, moisture content, and microbiological activity (Nzediegwu et al. 2019; Alkorta et al. 2004). Transport of heavy metals in the soil–water system occurs mainly in dissolved or suspended form, diffusion, or mass transfer, and processes that remove them from soil solutions are precipitation, coprecipitation, adsorption, and incorporation into biological systems (Wang et al. 2019; Tang et al. 2019).

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The chemical diversity of heavy metals has resulted in a variety of shapes in which they occur in the aquatic environment, as well as their participation in various interactions that affect solubility, mobility, and toxicity. The explanation for this phenomenon can be found in the fact that heavy metals interact with both dissolved substances in water, as well as with solid (rock, soil, sediment). The most striking example is iron, which can be seen as a potential substrate of heavy metals. It is precisely the iron substrate that often determines the further fate of heavy metals in the aqueous environment through the process of sorption and coprecipitation in the oxidation environment and remobilization of the microelements in the reductive dissolution of iron hydroxide (Qian et al. 2010; Alkorta et al. 2004; Amonette and Rai 1990). Also, manganese in the oxidized form (such as hydroxide) is a significant substrate of heavy metals, but also their potential source in decreasing redox potential (Shaw et al. 1990). Therefore, a large number of heavy metals in water are often found and transported in the form of fine colloids adsorbed and coprecipitated on colloidal hydroxides of iron and manganese (Efecan et al. 2009). Heavy metals can be transported through water through sorbents and on organic substrates. Also, mineral clays, due to their widespread distribution in the environment, represent unavoidable links in the heavy metal cycle because finely dispersed clay particles represent important metal substrates due to their pronounced adsorptive and ion-exchange properties (Mazur et al. 2018; Blanco Delgado et al. 2011). Carbonates and sulfides can also be heavy metal substrates. In assessing the behavior of heavy metals in the aquatic environment, one should take into account the forms in which metals occur, possible immobilization mechanisms and the mobilization mechanism.

Heavy Metals in Wastewater Bearing in mind the stricter legal regulations of the last few years regarding the content of primarily heavy metals in wastewaters, before their discharge into watercourses, the removal of heavy metals as the most important polluting substances in the environment is one of the most serious environmental problems of today (Wang et al. 2019). Removal of heavy metals is especially important because of their biological persistence. Heavy metals in contact with aquatic ecosystems are subject to biochemical transformations and turn into much more toxic forms. Most of the flora and fauna in the world accumulate heavy metals, so the concentration in these can be ten times higher than in water. The rapid development of all kinds of industry, and above all mining and metallurgy, the artificial fertilizer industry, the battery industry, and the electronics industry, has led to polluted water that contains heavy metals. Heavy metals are elements having an atomic mass between 63.5 and 200.6 g/mol and a density greater than 5.0 g/cm3. A large number of elements fall into this category, but those listed in Table 12.3 are of particular importance for the protection of health and the environment (Joseph et al. 2019).

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Table 12.3 Characteristics of common heavy metals

Heavy metal Copper (Cu) Lead (Pb)

Cadmium (Cd) Chromium (Cr) Arsenic (As) Mercury (Hg)

Human health effects Gastrointestinal issues Liver or kidney damage Kidney damage Reduced neural development Kidney damage Carcinogenic Allergic dermatitis Diarrhea, nausea, vomiting Skin damage Circulatory system issues Kidney damage Nervous system damage

Common sources Naturally occurring Household plumbing systems

Max. contam. level US EPAa WHOb
1 mgL mgL
1 1.3 2.0

Lead-based products Household plumbing systems

0.0

0.01

Naturally occurring Various chemical industries Naturally occurring Steel manufacturing

0.005

0.003

0.1

0.05

Naturally occurring Electronics production

0.010

0.010

Fossil fuel combustion electronics industries

0.002

0.006

Reprinted with permission of Chemosphere from Joseph et al. (2019) a Values established by the United States Environmental Protection Agency (USEPA 2019) b Values established by the World Health Organization (WHO 2017a, b)

Organic pollution is biodegradable and as such does not represent an unsolvable problem for the living world and the environment. However, the problem with heavy metals is that they are accumulated in living organisms and cause serious damage and diseases that can be carcinogenic and lead to death. The mechanism of how metals become toxic can be different. Metals can modify the active form of biomolecules, to block different biological functions of certain groups of biomolecules (e.g., proteins or enzymes) or to replace essential metal ions in biomolecules. The most commonly available electron donors for binding with metals are the carboxyl, amino, and sulfide groups, the active sites of the most important enzymes involved in the transport of oxygen and cell energy. Similarities in the toxicity of metals can be understood by the classification according to the reactivity of the metal as shown in Fig. 12.3. Metals from class A have a tendency for oxygen, while metals from class B have a tendency for nitrogen or sulfur. According to toxicity, the most toxic are ions of class B metals and then interclass ions, and the least toxic ones are class A ions. From the ecotoxicological aspect, the most interesting is lead, Pb and cadmium, Cd. They are not essential, they do not have any known metabolic role, their presence in the organism is due solely to contamination, they have a high ratio

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Fig. 12.3 Division of metals against toxicity. Similarities in the toxicity of metals can be understood by the classification according to the reactivity of the metal. (Reprinted with permission of Chemical Engineering Science from Krstić et al. (2018))

of anthropogenic and natural intake to the body, and they are endocrine disruptors and immunosuppressors. The most toxic class B has different toxicity mechanisms. Pollution of water caused by cadmium, chromium, copper, lead, mercury, nickel, and arsenic is especially dangerous for the living world. Some of these can be found in several oxidation states (Table 12.4). Cadmium and lead are most commonly found in the form of bivalent ions, so they can enter the food chain and cause serious degenerative changes and death of living organisms because they block functional groups of essential compounds and thus prevent the binding of essential cations (manganese, Mg; calcium, Ca; zinc, Zn; iron, Fe). In this way, they prevent metabolic and other life processes and replace other ions of metals in the compounds by modifying the activity of biologically important molecules (Oyar et al. 2007; Zhang et al. 2010). Arsenic appears in drinking water mainly in the form of arsenite, As3+, or arsenate, As5+, where arsenite, As3+, is significantly more toxic than arsenate, As5 + , and several hundred times more toxic than methylated arsenic compounds (Mohan and Pittman 2007a, b; Ghosh et al. 2019). Heavy metals are present in groundwater due to the dissolution of minerals from the earth’s crust or human activities (processing of ore). Large concentrations of arsenic in groundwater occur, for example, in wells and springs (Ghosh et al. 2019; Tubić et al. 2010a). It should be

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Table 12.4 Chemical properties of heavy metals

Heavy metal 1 Copper 2 Lead 3 Cadmium 4 Chromium

Molecular weight (g mol
1) 63.5 207.2 112.4 52.0

5 6 7 8

58.7 74.9 65.4 200.6

Nickel Arsenic Zinc Mercury

Oxidation state(s)a +1,+2 +2, +4 +2 0, +2,+3, +6 0,+2, +3
3,+3,+5 +2 +1,+2

van der Waals radius (10
12 m) 140 202 158 200

Electronegativity (Pauling scale) 1.90 2.33 1.69 1.66

Log KOW NA 4.02  0.28b 3.86  0.36b NA

163 119 139 155

1.91 2.18 1.65 2.00

NA NA NA 0.62c

Reprinted with permission of Chemosphere from Joseph et al. (2019) NA, not available a Bold values represent the most common oxidation state(s) for the heavy metal b Values determined experimentally by Sakultantimetha et al. (2009) c Values provided by Michigan Department of Environmental Quality

borne in mind that heavy metals (in particular copper), in certain chemical compositions and doses, perform important biological functions in the body. Toxicity occurs when the doses are inadequate and the chemical form of the metal is changed.

12.2.2 Metal Properties and Factors That Influence Their Removal from the Water Copper Copper can often be found in natural and waste mining waters in copper mine exploitations. In addition they can be found because they are widely used in agriculture, in machinery, in semiconductor, and in the electrical industry. Extensive studies have shown that ions of copper in water cause serious disturbances in plants and aquatic organisms. An excessive copper intake, according to epidemiological studies, has harmful effects on human health, leading to cirrhosis of the liver, kidney damage, hemolysis, vomiting, and cramps (Karabelli et al. 2011; Oyar et al. 2007; Xiao et al. 2011). Copper in an aqueous medium can be found in three basic forms, such as suspended, colloidal, and dissociable (as free copper ions and complexes with organic and inorganic ligands). Copper forms complexes with strong bases such as carbonates, nitrates, sulfates, chlorides, ammonia, and hydroxide and neutral ligands, such as ethylenediamine and pyridine. The intense sorption of copper is the reason for its high content in sediments. The degree of sorption of copper, as well as other metals, depends on the presence of clay particles, ligands, humic and fulvic acids, iron oxide and manganese, and pH and the presence of other cations. At higher

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concentrations of chloride in the aqueous system, the solubility of the copper increases, and it reduces the degree of its sorption due to the formation of copper chloride complexes. An increased copper content higher than 1000 mg/kg of dry mass in sediments is often associated with wastewater from the mine. Copper concentrations and bioavailability for aquatic organisms depend on pH, hardness, ionic strength, redox potentials, complex ligands, alkalinity, suspended matter, and water and sediment interactions (Oyar et al. 2007; Šćiban et al. 2015). Copper does not represent a major ecotoxicological problem, but due to its widespread use, copper affects the ecosystem. In small quantities, copper is essential for most organisms because it is necessary for the activity of various enzymes and for the utilization of iron. However, with the increase in concentration, copper can be toxic to both plants and animals. Aquatic organisms can accumulate copper in dissolved form by direct absorption over the body surface and in particular by ingestion of contaminated food. Abnormalities have been noted in both fish and birds as a result of the toxic effects of copper. Mammals are significantly less sensitive to the toxic effects of copper in aquatic organisms, as their absorption from the digestive tract is well controlled (Pešovski et al. 2017; Mâşu et al. 2011).

Factors That Influence the Removal of Copper Cu2+ Ions from Wastewater The effect of pH was investigated in the range of 3 to 11 with different nanomaterials (Xiao et al. 2011; Pešovski et al. 2017; Karabelli et al. 2011), and in all studies, it was noticed that the change in pH has an almost negligible effect on the degree of removal of copper Cu2+ ions, a fact which can be explained by the reduction mechanism of its removal. When Xiao et al. (2011) were examining this reduction mechanism, a blur of copper ions solution at pH > 5.5 was observed, which could indicate the existence of a hydrolysis process in an aqueous solution. On the other hand, the theoretical analysis of the copper solution using the appropriate software showed that copper Cu2+ ions are the dominant chemical form of copper in the solution up to pH 7. Figure 12.4 shows the Eh–pH diagram of the Cu–O–H system. At higher pH values, hydroxide forms are dominant, such as Cu(OH)+, Cu2(OH)22+, Cu(OH)2, Cu(OH)3 ¯, and Cu3(OH)42+. The experimental results of this study indicate an increase in the degree of removal of copper ions with an increase in pH, whereby this result can be explained by the strong influence of pH on the shelling of the nanoparticle shell. At a pH lower than its isoelectric point (IET), the nanomaterial is positively isolated, and when the pH is greater than isoelectric point, it becomes negatively charged and attracts cationic forms of copper, increasing the chance of interaction of the adsorbate and surface. With the increase in the concentration of copper Cu2+ ions, a decrease in the percentage of copper removal was observed (Karabelli et al. 2008a, b, 2011; Üzüm et al. 2009). A gradual decrease in sorption becomes noticeable only at high concentrations greater than 100 mg/L (Karabelli et al. 2008a, b, 2011). Concentrations greater than 200 mg/ L lead to a decrease in the sorption capacity, which can be caused by the formation of

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Fig. 12.4 A Pourbaix (the potential–pH) diagram of the system Cu–O–H (copper–oxygen–hydrogen), at 25  C, 105 Pa, ƩCu ¼ 10
10 molkg
1. (Reprinted with permission of the National Institute of Advanced Industrial Science and Technology (AIST) Japan from Naoto Takeno (2005))

a copper layer on the surface of the nanomaterial, which prevents access to active sites. Figure 12.5 shows the distribution of copper in pure metal solution and pH [Cu] ¼ 9.5 mg/L at 25  C in wastewater from India. In order to clarify the influence of other inorganic pollutants present in wastewater, Mazur et al. (2018) obtained a chemical equilibrium diagram by modelling the elaborated system MINEQL+ (the chemical equilibrium modelling system), using the physicochemical characterization of wastewater in different countries.

Lead Lead can act on all bodily organs and systems in the human body, both in adults and in children. Exposure to lead can cause a slight increase in blood pressure and anemia. Lead poisoning is most often due to the consumption of water or food contaminated with lead, but can also occur after accidentally ingesting contaminated dust or from paint (Zhang et al. 2010; Oyar et al. 2007). Long-term exposure to lead compounds, especially strong oxidants such as PbO2, can cause nephropathy. In

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100

100

Cu2+

327

Cu(OH)2 (S)

Cu(OH)2 (S) Cu2+ 80 Cu(OH)42–

Molar Fraction (%)

Molar Fraction (%)

80

60

40 Cu(OH)–3

20

Cu(OH)42– 60

40

20

+ CuCl– CuOH

Cu2(OH)2+2

CuOH+ 0

Cu(OH)–3

CuSO4

Cu2(OH)2+2

0 0

2

4

6

8 pH

10

12

14

0

2

4

6

8

10

12

14

pH

Fig. 12.5 Distribution of copper in (a) a pure metal solution and (b) a sample of wastewater from India containing metal as a function of the solution pH (ionic strength calculated by MINEQL+: a software environment for chemical equilibrium modelling). (Reprinted with permission of the Journal of Environmental Management from Mazur et al. (2018))

lower concentrations, it behaves like calcium and interferes with the conductivity of the ions in the nervous channels. Acute lead poisoning is treated with disodium calcium EDTA (ethylenediaminetetraacetic acid). This chelating agent has a higher affinity for calcium than for lead, which leads to the changes, and then is excreted through the urine leaving harmless calcium in the organism. In high contamination, lead is accumulated in the liver, kidneys, and bones.

Factors That Influence the Removal of Lead Pb2+ Ions from Wastewater Zhang et al. (2010) examined the various effects on the efficiency of removal of lead Pb2+ ions nano zeros with valence iron stabilized kaolinite. In the interval investigated of pH 1 to 6, it was found that there was a significant increase in the degree of removal of lead ions from 0% at pH 1 to about 90% at a pH of 4 to 6. An increase in the percentage of removal is observed at pH 2, but at pH 4–6, it does not change significantly. The removal rate, which does not exceed 94% under the above conditions, can be explained by the formation of the oxyhydroxide coating of iron Fe2+ and lead Pb2+ ions. When the pH exceeds 7.2, the concentration of hydroxide ions is high enough to cause the precipitation of Pb(OH)2. Accordingly, the optimum pH is in the range of 4 to 6. Figure 12.6 shows the potential–pH diagram for lead, Pb. The degree of sorption of lead in freshwater sediments depends on the characteristics of their granulometric composition and the content of organic matter. Thus, in the absence of soluble complexing agents, the lead is almost completely sorbed and precipitated when the pH is above 6.0. In reduction conditions, lead was successfully immobilized by precipitation with sulfide minerals, complexing with insoluble organic matter. Under oxidation conditions, it is successfully removed from water by precipitation with iron oxide minerals. In the acidic environment of

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Fig. 12.6 Eh–pH (the potential–pH) diagram of the system Pb–O–H (lead– oxygen–hydrogen). ƩPb ¼ 10
10 molkg
1, 25  C, 105 Pa. (Reprinted with permission of the National Institute of Advanced Industrial Science and Technology (AIST) Japan from Naoto Takeno (2005))

humic acid, it is more efficiently absorbed than the clay particles of lead, while in a basic medium, at a pH of 6.5, humic acid can form soluble complexes with lead (Oyar et al. 2007; Zhang et al. 2010). Figure 12.7 shows the distribution of lead in a pure metal solution (a) and (b) pH [Pb] ¼ 104 mg/L at 25  C, when the ionic strength was 0.1 M.

Cadmium Cadmium is known as one of the largest pollutants of the environment. In nature, it occurs in the form of eight isotopes, and, usually, oxidation state is +2 (Table 12.4). In nature, it is located together with copper, lead, and zinc. The main causes of the presence of cadmium in the environment are copper mines, the production of nonferrous metals (processing of copper, lead, and zinc ore), iron and steel production, fossil fuel burning, cement production, and incineration of municipal waste. Phosphorus fertilizers also have a high cadmium content, and their use contributes to the increased intake of this toxic metal into the soil, from where it quickly falls into groundwater. It belongs to a group of toxic industrial pollutants. Every year about 13,000 tons of this heavy metal is produced. Provided there are no phosphates and

12

Some Effective Methods for Treatment of Wastewater from Cu Production 100

Pb2+

Pb(OH)2 (s)

80 Molar Fraction (%)

329

Pb(OH)42–

60

40

20 PbOH+ 0 0

2

4

6

8

10

12

14

pH

Fig. 12.7 Distribution of lead species present in the aqueous medium as a function of pH, a chemical equilibrium diagram by modelling the elaborated system MINEQL+: a software environment for chemical equilibrium modelling. (Reprinted with permission of the Journal of Environmental Management from Mazur et al. (2018))

sulfates that can precipitate in the aqueous environment, cadmium is always present up to pH 8 as a divalent positive ion (Valko et al. 2005). Cadmium and its compounds are poisonous even at low concentrations and accumulate in the body. Swallowing the least amount of cadmium will cause immediate poisoning and permanent damage to the liver and kidneys (Zalups and Koropatnick 2000). Cadmium is deposited in the liver, kidneys, and bones and to a lesser extent in the salivary glands, reproductive organs, and pancreas. Its increased concentration can cause lung cancer. When it enters the bloodstream, it binds to erythrocytes, albumin, and low molecular weight proteins. Chronic exposure to cadmium leads to kidney dysfunction, while high levels of exposure can lead to death (Oyar et al. 2007; Zalups and Koropatnick 2000).

Factors That Influence the Removal of Cadmium Cd2+ Ions from Wastewater Chowdhury and Yanful (2013) examined the various effects of the removal of cadmium Cd2+ ions mixed maghemite-magnetite nanoparticles and concluded that the adsorption capacity of the material increases with an increasing pH and that the maximum adsorption is reached at pH 9.3. The adsorption of cadmium is higher in alkaline environments in the pH range of 8 to 10 than in acidic environments. In highly acidic environments, there is a chance that the adsorbent will dissolve and thus reduce the number of adsorption sites. The surface of the adsorbent is highly

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Fig. 12.8 Eh–pH (the potential–pH) diagram of the system Cd–O–H (cadmium–oxygen– hydrogen). ƩCd ¼ 10
10 molkg
1, 25  C, 105 Pa. (Reprinted with permission of the National Institute of Advanced Industrial Science and Technology (AIST) Japan from Takeno (2005))

protonated in an acidic medium that is not suitable for the removal of cadmium because in such a medium, cadmium Cd2+ is the dominant ion (Singh et al. 1998). As the pH increases, the degree of surface protonation gradually decreases and achieves zero at pH 7 resulting in a gradual increase in adsorption. About pH 8–9, where cadmium Cd2+ ions and CdOH+ are present in the solution, the adsorbent surface becomes negatively charged, giving rise to an electrostatically suitable environment for greater cadmium removal Fig. 12.8, (Singh et al. 1998). Figure 12.8 shows a Pourbaix diagram for cadmium, Cd. Thermodynamic calculations have shown that different hydroxyl forms of cadmium Cd ions are represented depending on the pH of the solution. The surface of the metal oxides in the aqueous solution is usually surrounded by hydroxyl groups that can change the shape at different pH values. These groups contain double pairs of electrons together with a dissolved hydrogen atom and can represent suitable conditions for them to react both with acids and bases. Electrode oxidation of the surface of the iron oxide is dominated by adsorption or proton desorption and is caused by the dissociation (ionization) of the surface hydroxyl groups depending on the pH of the solution. Thus, the magnetite will give iron Fe2+ ions and its hydrolysis products (FeOH+, Fe(OH)2, and Fe(OH)3) depending on the pH of the solution. The

Some Effective Methods for Treatment of Wastewater from Cu Production

100

Cd(OH)–3

Cd2+

Molar Fraction (%)

80

2–

Cd(OH)4

40

Cd(OH)2

20

Cd(OH)–3 Cd(OH)2 (s)

80

Cd(OH)2 (s)

60

Cd(OH)42– 60

+

CdCl

CdOHCl Cd(OH)2

40

Cd(OH)+ 2+ 20 Cd

Cd(OH)+

0

CdCl2 CdSO4

0 0

2

4

6

8 pH

10

12

14

331

100

Molar Fraction (%)

12

0

2

4

6

8

10

12

14

pH

Fig. 12.9 Distribution of cadmium Cd solution in wastewater from India as a function of pH, a chemical equilibrium diagram by modelling the elaborated system MINEQL+: a software environment for chemical equilibrium modelling. (Reprinted with permission of the Journal of Environmental Management from Mazur et al. (2018))

precipitation of cadmium Cd2+ ions will not occur at pH values lower than 9.5 (Figs. 12.8 and 12.9), when the solution contains small concentrations of cadmium Cd (Singh et al. 1998). Figure 12.9 shows the distribution of cadmium in a pure metal solution (a) and (b) pH [Cd] ¼ 0.5 mg/L at 25  C and ionic strength calculated by MINEQL+: a software environment for chemical equilibrium modelling.

Chrome Chrome is often an industrial contaminant of water and soil (Shi et al. 2011a). In wastewater from the copper mine, there are also trivalent and hexavalent chromium anions (Rahmani et al. 2011). The main oxidation states of chromium are Cr3+ and Cr6+. More toxic and thermodynamically stable is chromium Cr6+ (Liu et al. 2010; Fu et al. 2013). Chromium Cr6+ is highly soluble in water and expresses its toxic effects on various aquatic organisms even at very low concentrations (Shi et al. 2011a; Jobby et al. 2018). The conditions of the interchange chromium Cr3+ and Cr6+ are close to the conditions governing natural waters. Chromium Cr6+ is readily reduced in the presence of Fe2+, dissolved sulfides, and certain organic compounds. Chromium is carcinogenic, mutagenic, and teratogenic. Chromium Cr6+ is 1000 times more toxic than chromium Cr3+, so chromium toxicity is primarily associated with the concentration of chromium Cr6+ in the medium, precisely because of the extreme toxicity of this chromium form. Chromium Cr6+ is more toxic because it easily passes through the cell membranes, thereby reaching the cells where it is reduced to chromium Cr3+, which then interacts with macromolecules and thus exhibits its toxic and mutagenic effect. It is certain that exposure to certain concentrations of this element presents a significant risk to both the environment and human health.

332

V. Krstić

Factors That Influence the Removal of Chromium Cr6+ Ions from Wastewater With increasing pH, the rate and efficiency of chromium (VI) removal gradually decreases, which, depending on pH, is in the form of salt H2CrO4, H2CrO4 ¯ , and CrO42 ¯ (Shi et al. 2011b; Qian et al. 2010). H2CrO4 dominates at pH NH

1.3 – 8–11 13

O S N N

SPS AA Cto, AA Cti, PG, peptide bond

– 11.6–12.6 6.0

N N N

AA AA AA

Phosphonate


C¼ONH2 >NH
C
N
H > CH H
C
N OH
P¼O
OH

Phosphodiester

>P¼O
OH

0.9–2.1 6.1–6.8 1.5

O O O

PL PL TA, LPS

Binding group Hydroxyl Carbonyl (ketone) Carboxyl Sulfhydryl (thiol) Sulfonate Thioether Amine Secondary amine Amide Imine Imidazole

Reprinted with permission of the Journal of Environmental Management from Zeraatkar et al. (2016) PS polysaccharides, UA uronic acids, SPS sulfated PS, Cto chitosan, PG peptidoglycan, AA amino acids, TA teichoic acid, PL phospholipids, LPS lipoPS

According to the metal classification presented by Pearson (1963), and then by Nieboer and Richardson (1980), the metal affinity for different ligands is shown, as shown in Fig. 12.3. Group A metals tend to form a bond with the Group I ligands (F
, O2
, OH
, H2O, CO32
, SO4
, ROSO3
, NO3
, HPO42
, PO43
, ROH, RCOO
, C¼O, ROR; where R is organic group) over oxygen. Group B metals have high affinity for Group III ligands (H
, I
, R
, CN
, CO, S2
, RS
, R2S, R3As), but also can form strong bonds with ligands from Group II (Cl
, Br
, N3
, NO2
, SO32
, NH3, N2, RNH2, R2NH, R3N, ¼N-, -CO-N-R, O2, O2
, O22
). The symbol R represents an alkyl radical, e.g., CH2
and CH3CH2
. Boundary ions can build relationships with all three groups of ligands with different affinities. Nieboer and Richardson (1980) wrote about ligands in biological systems and their affinities with respect to different classes of metal. According to HSAB (hard and soft acid and bases) principles, ions from class A such as ions of alkaline and alkaline earth metals, e.g., sodium (Na+), calcium (Ca2+), and magnesium (Mg2+), are capable of forming stable bonds with oxygen-containing ligands, for example, OH
, HPO42
, CO32
, and RCOOi¼C¼O. Ions belonging to class B (some ions of heavy metals) such as mercury (Hg2+) and (Pb2+) ions build strong links with CN-, R-S-, -SH-, RNH2
, and an imidazole group, containing atoms of nitrogen and sulfur. Most of the toxic metals (nickel Ni2+, copper Cu2+,

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Some Effective Methods for Treatment of Wastewater from Cu Production

349

zinc Zn2+, cadmium Cd2+) can build relationships with the aforementioned ligands, but also with R2NH, R3N, ¼ N-, and -CO-N- groups. Class I ions mainly comprise ionic connections, while ions of class B build covalent connections (Fig. 12.3). Differences in the efficiency of biosorption originate from the nature of the biosorbent, i.e., from the presence of various chemical components on the surface of the cells. Different polysaccharides that are components of the cell wall of mushrooms and algae play a very important role in binding ions of metal (Wang and Chen 2009). A good example is chitin, which is found in the cell mushroom wall and in the exoskeleton of the scales and contains an amine (-NH2) group that is frequent and active in terms of binding heavy metal ions (Niu and Volesky 2006), while in the cell wall of the algae, an extraordinary affinity for ions of heavy metal showed a biopolymer alginate because it contains carboxyl groups responsible for binding these ions (Davis et al. 2003). In the case of various biological materials, it has been shown that the proteins are involved in the binding of metal ions (Wang and Chen 2009). Many functional groups directly participate in the binding of ions of heavy metals, for example, the aforesaid carboxyl group or groups containing O
, N
, S
, or P
(Wang and Chen 2009). Zeraatkar et al. (2016) summarized the most important functional groups in different biological systems with pKa values and from the aspect of biosorption in Table 12.9. The pKa value is one method used to indicate the strength of an acid. pKa is the negative log of the acid dissociation constant or Ka value.

Biosorption Mechanisms Removal of metals by the biosorption process is a complex process that is often not based only on one mechanism but involves a combination of several mechanisms such as electrostatic attraction, complexing, ion exchange, formation of covalent bonds, the presence of Van der Waals attraction forces, adsorption, and microprecipitation (Abdolali et al. 2014). Which of the mechanisms will be dominant and to what extent depends on the chemical composition and structure of the biosorbent and on the chemical composition and characteristics of the solution. In order to understand the mechanism of biosorption, it is necessary not only to have information on the structure and composition of biomass but also to know the chemical composition of the solution in which the process occurs. The characterization of the material can be tested with various methods, together with the determination of chemical processes that take place on the surface of the biosorbent. Depending on the nature of the biosorbent, various techniques can be used to characterize the material. Analytical techniques such as titration or more sophisticated instrumental analyses such as Fourier transform infrared spectroscopy (FTIR), scanning electron microscopy (SEM), Raman microscopy, energy-dispersive X-ray (EDX), X-ray photoelectron spectroscopy (XPS), X-ray diffraction (XRD), and nuclear magnetic resonance (NMR) can be used (Nakbanpote et al. 2007). Each of these techniques gives a piece of information complementing the image of the heavy metal ion binding mechanism for the biosorbent. The results of some of the above

350

V. Krstić

Fig. 12.19 IR and a part of SEM (scanning electron microscopy) images of orange waste and a part of XPS (X-ray photoelectron spectroscopy) analysis. (Modified after Feng et al. (2009), Guiza (2017), and Ren et al. (2017)) Table 12.10 Range of wave number (cm
1) of Fourier transform infrared spectroscopy adsorption peaks corresponding to some functional groups

Functional group –OH; –R3C– OH –COOH; – COOM –CH3, –CH2– –HC¼O, R2C¼O –NH2, –R2NH –PO– –SO– –NO–

Range of wave number (cm
1) 3200–3600; 1000–1200 (C–O) 1670–1760(C¼O); 1000–1300(C–O); 1400–1650 2800–3000 1680–1750(C¼O) 3200–3500(–NH); 1500–1650(C–N and N–H) 1000–1400 1000–1400; 1000–1300(–SO3) 400–700

Modified after He and Chen (2014)

techniques were shown on Fourier transform infrared spectroscopy for cellulosic waste orange peel (Fig. 12.19). Characteristic spectral values of wave number (cm
1) (He and Chen 2014) of characteristic functional groups are shown in Table 12.10.

Application of Biosorption Technologies Although biosorption has been known to the scientific public for some time, its possibilities are still being explored. However, most of these tests in scientific literature were carried out exclusively at the laboratory level in batch systems or

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Some Effective Methods for Treatment of Wastewater from Cu Production

351

mini-columns. Nevertheless, several commercial products have emerged that have been designed to remove heavy metals from industrial and mine wastewater. Among them, the most famous sorbents are called BIO-FIX (biosorbent which uses biomass from a variety of sources), AMT-BIOCLAIM™ (from Advanced Mineral Technologies), and AlgaSORB™ (an algal biosorbent) (Michalak et al. 2013). The US Bureau of Mines has created a sorbent called BIO-FIX. These granules are intended for the removal of heavy metals from mining wastewaters and industrial and surface waters (Michalak et al. 2013). The biosorbent is composed of various biological materials such as Spirulina sp., Lemna sp., Sphagnum sp., yeasts, and algae, which are immobilized in porous granules. This biosorbent is efficient and useful for the treatment of wastewater containing heavy metals and is very stable and can even be regenerated after over 200 cycles (Jeffers et al. 1993). The B.V. SORBEX Inc. Company from Canada produced a biosorbent BV-SORBEXTM for the removal of heavy metals from wastewater, based on the use of various algae: Sargassum natans, Ascophyllum nodosum, Halimeda opuntia, Palmyra pamata, and Chondrus crispus. The biosorbents of this company are efficient in a wide range of pH and easy to regenerate, and the metal biosorption is also effective in the presence of calcium and magnesium (Volesky 1990; Eccles 1995). Biosorbent AlgaSORB™ based on immobilized freshwater algae Chlorella vulgaris has been developed by Bio-Recovery System Inc. (USA). This material effectively removes ions of metal from a solution of 1–100 mg/L and can withstand more than 100 regeneration cycles (Kuyucak 1990; Gupta et al. 2015a, b).

Advantages and Disadvantages of Biosorption In scientific literature, biosorption is presented as a very efficient process, which does not require high capital investments, and operating costs are not large; therefore it is interesting primarily from the economic point of view (Michalak et al. 2013; He and Chen 2014). At the same time, the ecological aspect of biosorption processes is reflected in the use of biological materials that are very often cheap and/or waste material (or by-products) from different industries. Therefore, the advantages of biosorption in relation to conventional methods are low investment, high efficiency, easy maintenance, easy regenerating biosorbents, and the possibility of separating useful components of precious metals. Biosorbents are often compared with ion-exchange resins, and the biosorption process is also referred to as the pseudoion-exchange process (Gadd 2009). However, ion-exchange resins can be designed so that they are strictly selective for the metal which needs to be isolated. The use of the polluted biomass for the purpose of eliminating pollutants also has its disadvantages, which are poor mechanical characteristics, loss of mass after regeneration, and difficult separation of small particles of used biomass from the treated effluent (Liu and Liu 2008). However, immobilization and/or granulation of biomass largely overcomes this problem. The use of natural or synthetic polymers (alginate, agar, silica gel, polysulfones, agarose gel, and polyacrylamide) has proved

352

V. Krstić

to be very effective in immobilizing various biological materials for the purpose of removing metals from water.

Biosorbents The first research related to an investigation into some microorganisms which accumulate metal ions through metabolism. This discovery initiated a large number of studies, firstly from the toxicological point of view, in order to examine the effects of pollutants on the metabolic activity of the cells and the consequences of the accumulation on the food chains (Volesky 1990; Jobby et al. 2018). However, the use of living organisms for the purpose of removing pollutants from wastewater is described in Sect. 12.8 of this paper, entitled Phytoremediation. The cheapest biomass is industrial waste or by-products of the industry, and the elimination of such waste and its reuse contribute to solving the problem of waste disposal, thus preventing contamination of ecosystems at global, regional, and local levels. Especially important are an agro-industrial waste, compared to commercial resins, the easy availability of large amounts, the fact that this waste does not have to be cultivated or produced for a special purpose, and its very low price (O’Connell et al. 2008). Although biomass is most often mentioned in the context of renewable energy sources, along with hydropotential, geothermal, and wind and solar energy, it could also be a potential source of biosorbents that could be used for wastewater treatment.

Fixed-Bed Column Adsorption The use of a fixed-bed column is another method of using biosorption. Depending on the application, the method of preparing the column with a fixed layer is individual. At the adsorption of toxic metals into the fix-bed column is influence by the nature of the adsorbent, the layer height, the pH and the flow velocity. The diversity of the manner of preparation and materials used mean that the different conditions do not promise a universal understanding of the impact of variables on the applied system. Because of this, statistical models are used to establish the relationship between the key variables and the desired results (Pedrosa Xavier et al. 2018; Kumar et al. 2016). This saves material, test time, and money. The general scheme of the fixed-bed adsorption column is presented in Fig. 12.20 (Yanyan et al. 2018). Columns are filled with adsorbents which can be regenerated and connected to the source of the retention tank through the pump. At the exit there is purified water (receiving tank). Fixed-bed adsorption columns are usually filled with different adsorption materials, and they work under different operating conditions, and the calculations for using the fixed-bed column can be made according to different models as given in Table 12.11 (Chatterjee et al. 2018). To predict the adsorption capacity, the kinetic constant, and the fixed-layer operating conditions, several models were applied (Acheampong et al. 2013; Chu

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Some Effective Methods for Treatment of Wastewater from Cu Production

353

Fig. 12.20 A general scheme of the fixed-bed adsorption column. (Reprinted with permission of the Journal of Environmental Management from Yanyan et al. (2018))

2010). One of the most commonly used models of adsorption of metals in a fixedlayer column is the bed depth service time (Izkuierdo et al. 2010). Bed depth service time is a model that can predict the relationship between metal concentration and the operating parameters of the fixed-bed system. In recent literature, several papers published on the adsorption of Cu2+ ions put forward a proposal using a fixed layer with static binary adsorption and with a dynamic binary system (Freitas et al. 2017, 2018). Combined acid wash and biosorption methods have proved in this article to be adequate for purifying Cu2+ from polluted water. Adaptive neural fuzzy inference system has been shown to effectively forecast the optimum working conditions for biosorption of toxic metals from wastewaters (Calero et al. 2018; Choińska-Pulit et al. 2018; Gao et al. 2018). Table 12.11 gives assumptions, advantages, and limitations for different fixed-bed models. Other research scientists such as Borba et al. (2008), Chen et al. (2017), Abdolali et al. (2017), Barquilha et al. (2017), Chao et al. (2014), Shah et al. (2013), Pedrosa Xavier et al. (2018), and Kapur and Mondal (2016) have investigated adsorption of Cu2+ into a fixed-bed column with different adsorbents and under different conditions.

354

V. Krstić

Table 12.11 Comparison between different models used for simulation Model Adams– Bohart

Model equations a  K K Z K C t
abD ab 0 C ¼ C o e ab 0

Thomas

C¼




1þ e

C0 K th q0 X
K th C 0 t Q

Yoon– Nelson

Co ∁ ¼ 1þerynkyn
kynt

Bed depth service time

Z min ¼ K abU oN o ln

Pore diffusion adsorption

(*)


C o C




1

Assumptions Equilibrium is not instant; adsorption of solute is directly proportional to the concentration of solute in the bulk fluid and residual adsorptive capacity of adsorbent; No axial dispersion is considered Follows Langmuir adsorption– desorption kinetics; Adsorption rate is controlled by surface reaction between adsorbate and unused capacity of adsorbent; No axial dispersion is considered Model based on theory of adsorbate probability; No axial dispersion

Model assumes that bed depth and service time of the packed bed follows a linear relationship; intraparticle diffusion and external mass transfer resistances are negligible, and adsorption takes place on the surface directly The bed is fully saturated, (i.e., all interparticle

Advantages Empirical model, hence, involves less computational time

Limitations Deviations between simulated and experimental data occur when axial dispersion is controlling phenomenon

Empirical model, hence, involves less computational time

Deviations between model predictions and experimental data occur when sorption is controlled by mass transfer rather than surface reaction

Simple form as compared to other models; involves less column parameters and computational time Empirical model, hence, involves less computational time; minimum column depth can be obtained for a particular saturation. Data obtained can be used for scale-up purpose

No detailed data concerning characters of adsorbents and adsorbate and axial dispersion Both intraparticle diffusion and external mass transfer resistances play a dominant role in adsorption process, which the model does not take into account

Realistic model involving mass balance

Requires more computational time. Liquid (continued)

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Some Effective Methods for Treatment of Wastewater from Cu Production

355

Table 12.11 (continued) Model

Model equations a

Assumptions

Advantages

Limitations

voids are filled with liquid); the velocity of the liquid phase is constant along the column

approach of liquid phase along the column. Model has strong theoretical basis

velocity does not remain uniform throughout the column generally

Reprinted with permission of the Journal of Cleaner Production from Chatterjee et al. (2018) a (symbol specifications are mentioned in nomenclature section) (*)  
     2   2 3k ∂Cp ∂C 1  i Dp ∂ C2p þ 2 ∂Cp ¼ DL ∂∂zC2
V ∂C
apf 1
ε ¼h εp ε ρs C
C p r¼ap r ∂r: ∂t ∂r ∂t ∂z ∂g 1þð1
εp Þρs ∂Cp    ∂C p where at r ¼ ap, k f C
C p r¼ap ¼ Dp εp ∂r

12.3.3 Low-Cost Adsorbents In recent years, for the removal of heavy metals, one recommendation has been to use cheap materials as the adsorbent, since in addition to the economic parameters that contribute to the preservation of the environment, these cheap materials will be given a new usage value, instead of being used for landfill (Mohan and Pittman 2007a, b; Guiza 2017; He and Chen 2014; Siti et al. 2013; Adegoke and Bello 2015). Previous applications of adsorbents for the removal of heavy metals comprise the use of cheap materials such as agricultural products and of by-products (husks, skins, sawdust, seeds, stalks, pods, scales, hair, bone), in addition to cheap materials from the industry of by-products and wastes (carbonized organic matter, coal dust, ash, red mud, slag from high furnaces, iron Fe3+, and chromium Cr3+ hydroxides), various types of soil and minerals (clay, sand, zeolites (Krstić 2020)), and bioadsorbents (chitin, chitosan, cellulose, biomass, bacteria, algae, mushrooms). Table 12.12 shows some of the low-cost adsorbents for the wastewater treatment of copper Cu2+ ions.

12.3.4 Cu and Modified Mesoporous Carbon and X-Ray Photoelectron Spectroscopy (XPS) There are various surface modification techniques, such as acid treatment, impregnation, and functional grafting of groups or molecules, which should be cited. These significantly change the properties of the adsorbent surface, such as the specific surface area, surface charge, metal dispersion, and hydrophobicity. Ren et al. (2017) supported Cu catalysts loaded on different carriers with a Cu content of 5 wt% through impregnation. In this case, the impregnation is a method

356

V. Krstić

Table 12.12 Reported data on copper Cu2+ ions removal by low-cost adsorbents Type of adsorbents Bacteria

Mushrooms

Algae

Fungi

Biosorbent Agricultural waste

Naturally occurring minerals

Industrial by-products

Low-cost Adsorbent Bacillus sp. Escherichia coli Pseudomonas putida Aspergillus niger Rhizopus oryzae Mucor rouxii Spirogyra (green algae) Ecklonia maxima (sea algae) Flammulina velutipes Auricularia polytricha Pleurotus eryngii Pleurotus ostreatus Active coal pecan shell Activated carbon Luffa Actangula carbon

qm (mg/g) or RE (%) 16.30 5.86 8.00 26.00 19.40 52.20 133.00 90.00 8.13 6.04 4.29 5.00 31.70 21.50 12.47

Shell rice Shell hazelnut Orange peel OP (orange peel) OPAA (modified orange peel) Watermelon seed hull

10.90 0.0635 86.73 44.28 289.00

Black carrot residue Peanut husk charcoal waste Zeolite, clinoptilolite

8.88 RE ¼ 60.0

Zeolite Clays HCl-treated clay

5.10 9.58 83.30

Fly ash–wollastonite Sawdust Steel-making by-product

1.18 13.80 40.00

33.90

1.64

Reference Salman et al. (2015) Bilal et al. (2013) Salman et al. (2015) Gautam et al. (2014) Bilal et al. (2013) Gupta et al. (2006) Feng and Aldrich (2004) Lia et al. (2018)

Bansode et al. (2003) Crini (2005) Siddiqui 2018 Vaghetti Pehlivan Sousa Bilal et al. 2013 Bilal et al. (2013) Feng et al. (2009)

Akaya and Güzel (2013) Güzel et al. (2008) Yargıç Babel and Kurniaw (2003) Crini (2005) Abollino et al. (2003) Vengris et al. (2001) Abdel Salam Crini (2005) Ajmal et al. (1998) Lopez et al. (2003)

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Fig. 12.21 XPS (X-ray photoelectron spectroscopy) analysis for (a) modified mesoporous silica– carbon nanocomposite and (b) catalyst of Cu. (Reprinted with permission of Chemical Engineering Journal from Ren et al. (2017))

for loading Cu on the catalyst support. They took a water solution of distilled water containing 0.1 g Cu(NO3)2 x 3H2O and impregnated with 0.5 g of MSC (modified mesoporous silica–carbon) and MC (dry mesoporous carbon). The pores were filled by Cu solution, and after drying and calcination, Cu was loaded on the catalyst support. In this way, they obtained the two catalysts, which were called Cu/MC and Cu/MSC. The pore volume and mean pore size in the modified mesoporous silica– carbon nanocomposite are significantly larger than that in mesoporous carbon. These characteristics may be beneficial to the dispersion of the loaded Cu species. The dispersion and specific surface area of copper in the Cu/MSC catalyst are higher than those of Cu/MC. Ren et al. (2017) have given a detailed explanation of the C 1s peaks of modified mesoporous silica–carbon and formation of C–C, C–O, and C¼O bonds in the free carbon matrix using XPS (X-ray photoelectron spectroscopy) analysis, probably deriving from the adsorbed carbonate species (Figueiredo et al. 1999; Liu et al. 2012; Hijikata et al. 2001). To obtain the copper–support interaction, Ren et al. (2017) used the X-ray photoelectron spectroscopy of so called Cu/MSC and Cu/MC samples, and these are shown in Fig. 12.21. The interaction between the active metal and the support is enhanced, favoring the dispersion of copper Cu2+ ions (Severino et al. 1998; Wang et al. 2002). X-ray photoelectron spectroscopy analysis is suitable as it shows the form in which copper is adsorbed in adsorbents (Huang et al. 2018; Xu et al. 2016; Deng et al. 2003).

358

12.4

V. Krstić

Different Types of Nanomaterials for Wastewater Treatment

Current advances in nanotechnology offer nanomaterials as one of the potential solutions in the removal of metals from water. Nanomaterials, made of particles having a size of 10 to 100 nanometers in diameter and having at the same time significantly improved electrical, magnetic, optical, and chemical properties, provide a much greater reaction surface, which gives a greater potential to this remediation material. In wastewater treatment, different effective environmentally friendly and cost-effective nanomaterials have been developed with the unique capacity to decontaminate industrial wastewater, surface water and groundwater, and drinking water (Gupta et al. 2015a, b). Nanomaterials for wastewater treatment can be classified into four categories: – Nano-adsorbents (activated carbon, carbon nanotubes, graphene, MgO, MnO, ZnO, TiO2, and Fe3O4). – Nanocatalysts (photocatalysts, electrocatalysts, catalysts based on fenton materials and chemical oxidants are potentially good agents). – Nanomembranes (carbon nanotubes, electrospray nanofibers, and hybrid nanomembranes). – Modern processes (alginic membrane reactors, fuel cells based on microorganisms and anaerobic processes) are used as the new technologies for wastewater treatment.

12.4.1 Nano-adsorbents In the last few years, nanoparticles have been studied as the potential adsorbents to purify wastewater. The adsorption process depends on the adsorption coefficient of pollutants, i.e., heavy metals or organic pollutants (Pešovski et al. 2018; Gupta et al. 2015a, b). Figure 12.22 shows a schematic illustration of transport, diffusion, precipitation, ion exchange, and surface adsorption of metal ions by the nanoparticles (Dubey et al. 2017). These include the metal nanoparticles, nanostructured mixed oxides, and magnetic nanoparticles. The recent development of carbon nanomaterials (CNMs) includes the carbon nanotubes, nanosheets, and nanoparticles. Different types of silicon nanomaterials are also used as the nano-adsorbents, e.g., silicon nanotubes, silicon nanoparticles, and silicon nanosheets. Factors that control the properties of nano-adsorbents are the size of nanoparticles, specific surface area, agglomeration sites, shape and fractal dimension, chemical composition, crystalline structure, and solubility.

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Fig. 12.22 A schematic illustration of the nanoparticle metal ions, by transport, diffusion, precipitation, ion exchange, and surface adsorption. (Reprinted with permission of Chemosphere from Dubey et al. (2017))

Oxide-Based Nanoparticles Oxide-based nanoparticles are inorganic nanoparticles, usually prepared with nonmetals and metals. These nanoparticles have been intensively used to remove hazardous pollutants from wastewater. These include titanium oxides, titanium oxide/dendrimer composites, zinc oxides, magnesium oxide, manganese oxides, and iron oxides. Oxide-based nanoparticles are characterized by their high specific surface area, minimal environmental impact, low solubility, and no secondary pollutants (Gupta et al. 2015a, b). Anjum et al. (2016) explain that the iron-based nanoparticles, such as ferric oxide, are a cheap material for adsorption of heavy metals. It is an environmental material and can be used directly in a contaminated environment with a lower secondary contamination. Factors that affect the adsorption of various heavy metals on the Fe2O3 nanoparticles depend on the pH value, temperature, dose of adsorbent, and incubation time. Manganese oxide nanoparticles (MnO2) show a high adsorption capacity due to the large specific surface and polymorphic structure. They are used to remove various toxic metals, especially arsenic from wastewater. Most commonly used modified MnO2 includes the nanoporous manganese oxide and hydrated manganese oxide (HMO) (Wang et al. 2011). Zaman et al. (2009) obtained the hydrated manganese oxide (HMO) by addition of MnSO4 x H2O to a NaClO

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solution. The modified HMO has a specific surface area of about 100.5 m2/g. The adsorption of various heavy metals such as lead Pb2+, cadmium Cd2+, and zinc Zn2+ ions to the HMOs (hydrated manganese oxides) usually occurs due to the effect of the internal sphere mechanism, which can be defined as the ion-exchange process (Parida et al. 1981). A high degree of efficiency in the removal of lead Pb2+ and cadmium Cd2+ from wastewater was achieved using the mesoporous ZnO nanotubes. Magnesium oxide (MgO) microspheres are the new structures that can improve the adsorption efficiency of the removal of heavy metals from wastewater. Various types of modifications of a nanostructural morphology were performed in order to increase the MgO adsorption capacity. This includes the nanotubes, nanoribbons, fractal nanostructures in the form of a fish bone, nano-wires, nano-cubes, and various three-dimensional shapes. Li et al. (2003) proved that an effective adsorption of lead Pb2+ and cadmium Cd2+ occurred on the mesoporous manganese oxide, MgO.

Carbon Nanotubes and Nano-adsorbents Based on Graphene According to Madrakian et al. (2011), the modified magnetic carbon nanomaterials can easily be removed from wastewater using a magnet. The use of multilayer carbon nanotubes eliminated lead Pb2+, manganese Mn2+, and copper Cu2+ ions. The adsorption behavior of alumina as the carbon nanomaterials coatings was studied, and it was found that the coated carbon nanomaterials showed a better ability to remove the toxic metals from wastewater compared to the uncoated carbon nanomaterials. Graphene is one of the allotropic carbon modifications, with the special characteristics to make it very favorable for wastewater treatment. Graphene oxide (GO) is a kind of carbon nanomaterial, which has a two-dimensional structure (Lingamdinne et al. 2016; Lim et al. 2018; White et al. 2018). Hummers’ method is the most commonly used method for the graphene oxide synthesis. Chowdhury and Balasubramanian (2014) have shown that, recently, graphene has attracted the attention of researchers to remove heavy metals, dyes, and organic pollutants and the adsorption of rare earth metal ions, due to its highly developed specific surface, flexibility, strength, low weight, density, and chemical stability (Table 12.13). Many researchers have used the nanomaterials based on graphene for adsorption of heavy metals from wastewater. Its bad side is a poor reduction of graphene oxide to the originally pure graphite, as it can reduce its electronic and mechanical properties. Synthesis of graphene as in the display of Lim et al. (2018) is shown in Fig. 12.23. Table 12.14 Lim et al. (2018) show the advantages and drawbacks for the removal of metal ions. Adsorption of heavy metals from wastewater with nanoadsorbents depends on several factors such as the temperature, pH value of solution, adsorbent doses, and duration of reaction. The characteristics of nano-adsorbents affect the absorption of heavy metals, and surface charge, hydrophobicity, and

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Table 12.13 Some characteristics of a single layer graphene at room temperature

Characteristic Thickness (C–C bond length) Optical transparency BET* specific surface area (theoretical) Thermal conductivity Density Carrier density Young’s modulus Fractural strength Resistivity Electron mobility

361

Value 0.142 nm 97% 2630m2 g
1 5000Wm
1 K
1 0.77 mgm
3 1012 cm
2 1100 GPa 125 GPa 10
6 Ω cm 200.000 cm2 V
1 s
1

Reprinted with permission of the Advances in Colloid and Interface Science from Chowdhury and Balasubramanian (2014) *BET, Brunauer–Emmett–Teller

addition of the new functional groups are important (Dubey et al. 2017; Ali et al. 2019a, b).

12.4.2 Nanomaterials as Catalysts and Electrocatalysts Nanocatalysts, like the semiconductors and oxide metals, are gaining increasing attention from researchers in wastewater treatment. For this purpose, various types of nanocatalysts, such as photocatalysts, electrocatalysts, and fenton-based catalysts, are used to improve the chemical oxidation of organic pollutants and antimicrobial activity (Ma et al. 2015; Yu et al. 2003). Photocatalysis is a promising technique for wastewater treatment. The nanomaterial/semiconductor requires some modification in order to reduce the energy of the narrow (band-like) gap from the UV (ultraviolet) to a visible spectrum of light. The ZnO and TiO2 nanomaterials have a narrow emission band of 3.2 eV, and their photocatalytic activities have been thoroughly studied. However, in the field of solar spectrum, both catalysts can absorb only a small fraction of the UV (ultraviolet) light, which reduces their efficiency (Anjum et al. 2017). Electrocatalysis in microbiological fuel cells is under development and is the subject of study in wastewater treatment and direct generation of electricity. In the microbiological fuel cells, the electrocatalyst has a limiting role in a fuel cell operation (Zhou et al. 2003; Chaturvedi et al. 2012). The Pt nanocatalysts, doped with the black carbon so-called XC72, are reported, showing a current density of up to 6.2 mA/cm2 in the electrocatalytic reaction of glucose oxidation, as well as a high potential for the reaction of ethanol oxidation in the fuel cells (Chen et al. 2015). The disadvantages of platinum limit its use in electrocatalysis due to its limited availability and high cost of obtaining. Also, during electrocatalysis, Pt may limit the

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Graphite

Top-down method

Bottom-up method

–Mechanical Exfoliation: Scotch tape & Atomic force microscopy tips

–Pyrolysis

–Chemical Exfoliation: Development of Hummers methods

–CVD: Low pressure CVD & Ultra-high vacuum CVD

–Chemical fabrication: Sonication process & Reduction of GO

–Plasma Synthesis: PECVD & plasma doping

–Epitaxial Growth

HO

O

HO

O O

HO

OH

O

O OH O O

OH

HO

O HO

O

HO

O

OH

Graphene

Fig. 12.23 Process flow chart of graphene synthesis. Symbols: GO, graphene oxide; CDV, chemical vapor deposition; PECVD, plasma-enhanced chemical vapor deposition. (Reprinted with permission of the Journal of Industrial and Engineering Chemistry from Lim et al. (2018))

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Table 12.14 Specific adsorption mechanisms of the graphene oxide-based nanomaterials for removal of metal ions

Graphene oxidebased nanomaterials Graphene oxide (GO)

Adsorption mechanisms included for metal ions removal Electrostatic interactions Ion exchange

Reduced graphene oxide (rGO)

Electrostatic interactions – Lewis–base–acid mechanism

Magnetic graphene oxide nanocomposites

Electrostatic interactions with graphene oxide Interactions with the particle surface

Graphene oxide materials functionalized with organic molecules

Magnetic properties of the nanoparticles Electrostatic interactions Complexation with organic molecules

Advantages Good dispersion in water Great colloidal constancy Contains rich oxygenated functional groups Restoration of sp2 domains. Better electron transport properties Bigger surface area compared to the pure GO Increased number of binding sites compared to pure GO Ease the recovery process from solutions Bigger surface area compared to pure GO Great colloidal stability Greater number of functional groups (–NH2, – OH)

Drawbacks Restricted number of sorption sites

Less oxygen-containing functional groups Weaker colloidal stability

Co-reduction of GO during the combination of the particles weakens the colloidal stability

Large variations of the stability of the loaded molecules depending on the alteration approach physically or chemically

Reprinted with permission of the Journal of Industrial and Engineering Chemistry from Lim et al. (2018)

reaction due to poisoning with the intermediate compounds (Chaturvedi et al. 2012). However, these problems can be overcome by the replacement of Pt with more affordable Pd nanoparticles. Trumić et al. (2015a, b, 2016, 2017) studied the mechanical and structural properties of the Pt–Pd system.

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12.4.3 Nanomembranes The application of nanotechnology in membrane processes is the most modern method for drinking water treatment and provides new functionalities. The main reasons for using this technology are in its advantages in terms of the treated water quality, effective disinfection, and small space for installation, making it very economical due to a simple design in relation to the other wastewater treatment techniques. Nanomaterials used for membranes also play a significant role in the chemical decomposition of isolated organic pollutants (Rajabi et al. 2013; Liu et al. 2014). The membranes made of carbon nanofibers (CNFs) showed an extremely high degree of selectivity in the filtration/removal of heavy metals. The zeolite-based materials, most commonly used in membrane making, are sodalite the MFI type and Linde type A (Liu et al. 2014). Contamination is caused by the flux reduction of filtration membranes, so due to this reason, it is necessary to clean the membranes chemically or mechanically, or in the event of a significant reduction in flux, a complete replacement of membranes is required. In this research, the main focus is on modifying the membranes by coating the hydrophilic layer of polymers (Jie et al. 2015). Research on the method of making the membrane-based carbon nanofibers, as well as composite membrane filters (CNT-MMMs), has shown the significant opportunities to achieve a high flux (Rajabi et al. 2013). Table 12.15 gives a comparison of adsorption and membrane filtration to separate zinc Zn from water. Table 12.15 Removal efficiency of zinc from water by the different functionalized CNTs (carbon nanotubes) and CNT membranes

1

CNTs and CNT membranes Functionalized MWCNTs

Removal efficiency 99%

Diameter (nm) 10–25 nm

Mode of the study Adsorption

2

Oxidized CNT sheets

58 mg/g

N/A

Adsorption

3

MWCNTs modified with chitosan Nitrogen-doped magnetic CNTs Oxidized MWCNTS

90%

60–100 nm

Adsorption

609 mg/g

N/A

Adsorption

18 mg/g

15 nm

Adsorption

34.36 mg/g

> Cd > Zn with more choice shown for copper. Pb and Cd have been remediated recently by Grenni et al. (2019b) using organic waste (dry water hyacinth and sodium alginate) after blending into 1:1 ratio, and their composite was obtained after cross-linking and then applied for removal of the two heavy metals for long durations starting from 6 h, 4 days, 11 days, and also sampling up to 126 days. Here, density functional theory adds on interesting features of summing up of dipole moment and its highly enhanced value, i.e., 63.290 D for the blend/composite developed as microspheres from the dipole moment, 6.997 D for sodium alginate, and 18.849 for water hyacinth dry matter, and consequently, such computational tools help the researchers greatly (Anglada et al. 2009; Huang et al. 2018; Malik et al. 2019). After fabricating these aforementioned blends as microspheres and their use as remediation of toxic levels of metals, people have modified and composited the easily available biomass (Dong et al. 2019). Thus, the nanofibres were made obtainable using in-situ synthesis from hydrochar got through hydrothermal carbonization of abundantly found biomass in the presence of pressurized hot water where glucose acted as its precursor in one-pot hydrothermal reaction and Ti3AlC2 derived nanosphere to remove Cd(II) and Cu(II) from wastewater. The method was also found as fluoride-free as well as scalable, and the composite was proved to be a promising candidate. Modern environmentalists are also approaching different synthetic routes to obtain microspheres of carbon obtainable from plants to improve the uptake potentiality by increasing the available surface area (Zbair et al. 2019) and hence, they have prepared carbon microspheres using walnuts shells by microwave supported pyrolysis. Thus, the product such as microspheres was depicted with Fourier transform infrared spectroscopy, Raman spectroscopy, and X-ray powder diffraction and employed for the removal of Cu(II), Cd(II), Cr(III), and Pb(II) heavy metal ions. Even more, the remediation was also authenticated by studying density functional theory and optimization of the energy of the carbon microspheres. The average diameter of the microspheres was found to be 4.55 μm using the above designated techniques like scanning electron microscopy and transmission electron microscopy. Herein, the uptake capacity was found remarkably high for Cr(III) with 792 mg/g, 638 mg/g for Pb, 574 mg/g for Cd(II), and 345 mg/g for Cu (II) at pH ¼ 5. A broad-scale study on biosorbents when applied either native or modified, for pollutant remediation including heavy metals, is being vigorously exercised for the last two-three decades. But, these materials show a noticeable drawback that they release some organic compounds soluble into the water, take longer extraction timings with metal speciation difficulty that limits the use of such biosorbents at the commercial level, with emphasis for drinking water.

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13.3.2 Adsorbents with Various Forms of Carbon Pb(II), Cd(II), and Hg(II) sequestration was procured by different forms of carbon (biochar) or processed carbon graphene oxide and carbon nanotubes (all as individual investigated adsorbent) along with density functional theory and energy optimizations for each form. In this research work, graphene oxide was found to have the strongest tendency to remove all the three heavy metals, followed by biochar 600 for Pb(II) which means biochar obtained at 600  C pyrolysis whereas biochar 300 (biochar at 300  C pyrolysis) for both Cd(II) and Hg(II) ions. Activated carbons (Deliyanni et al. 2015) and carbon nanotubes (Zhang et al. 2019a) each taken as a single adsorbent have proved comparatively low removal efficiency for the studied metals. Hence, such results may direct the upcoming researchers toward selectively designing of such adsorbents for further modifications in order to increase their removal efficiency. Some more recent work related to this class is also placed in Tables 13.3a, 13.3b, and 13.3c.

13.3.3 Nanocomposites Without Carbon Forms as Adsorbents Several variations of composites, predominantly nanocomposites, are the key attraction for scientists of the day with their new avenues of applications, covering the existing problem of dealing with the heavy metal treatment of wastewaters. A nanocomposite fabricated from nanocrystalline TiO2 as well as MoO3 at 180 C by hydrothermal method and depicted via x-ray powder diffraction and transmission electron microscopy methods. Then, the nanocomposite thus obtained was exploited to remediate Cr(VI) and two dyes where 59 mg/g was the outcome as uptake for Cr (VI) by the adsorbent (Zhao et al. 2018). Silver–yttrium oxide (Pradhan et al. 2017) ellipse like composites with nanosize was developed by solution combustion followed by their characterization through X-ray powder diffraction, atomic force microscopy, scanning electron microscopy, transmission electron microscopy, Fourier transform infrared spectroscopy, thermogravimetric analysis, and zeta potential experimentation procured for the remediation of Cr(VI) and Cu(II) from wastewater. The uptake capacity values for adsorbed Cu(II) was found  773 mg/g and for Cr(VI) was  720 mg/g as the maximized capacity for both the metals at pH ¼ 6. Some more researches’ overviews are placed in Table 3, part (b) with and without carbon forms.

13.3.4 Nanocomposites with Carbon Forms as Adsorbents A binary nanocomposite obtained via treating TiO2 with reduced graphene oxide (Vajedi and Dehghani 2019) in their different mass ratios were tried for Pb(II) and

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Cu(II) removal, out of that 1:1 ratio was observed to perform as maximized adsorption tendency for both the metals. Herein, the removal was employed using electrochemical techniques, i.e., cyclic voltammetry and square wave anodic stripping voltammetry. Garg et al. have done a genuine positioning of various nanocomposites consisting of a variety of carbon forms and their discussion toward heavy metals remediation.

13.3.5 Other Hybrids as Adsorbents Precipitation–calcination procedure was adopted to synthesize magnetic natured Fe@MgO nanocomposites (Ge et al. 2018) followed by their applications in the removal of Pb(II) and methyl orange; thus uptake capacity for Pb(II) was found to be 1476.4 mg/g, whereas for the dye, it was 6947.9 mg/g that proves a remarkable capacity. Highly efficient magnetic capturing of Pb and Cd from blood using magnetic capturing agents in nanosizes has been considered by Guo et al. (2019) in a continued adsorption mode. An improved removal efficiency for each of these heavy metals was seen by the iron nano capture candidate via interactions and along with the redistribution of heavy metals in the blood that was quantitatively studied. The findings showed that Pb was removed with 97.97% efficiency and Cd with 96.53% efficiency by involved complex from blood in 120 min via a continual multistep pathway, and thus the residual concentrations of both metals were decreased dramatically. A pristine nanocomposite comprising magnetic graphene oxide coated with SiO2 and further modified with polypyrrole-thiophene (Molaei et al. 2017), (reduced graphene oxide/SiO2@co Polypyrrole-Thiophene), a co-polymerized composite was investigated for simultaneous remediation of several heavy metals. Thereafter, the uptake capacity of the nanocomposite was reported as 201 mg/g for Cu(II) and 230 mg/g for Pb(II), and 125, 98, and 80 mg/g were found to be the order of Zn(II), Cr(III), and Cd(II), further implemented for real water also. In order to get a detailed visualization and understanding of a large class of nanocomposites (nonmagnetic and magnetic type), an erudite review is available that focuses more on the assessment of polymer-based composites/nanocomposites with the inclusion of carbon and/or metal oxides, in an estimable way. Some researchers of modern times have examined a nice variety of adsorbents consisting carbon-based, polymer-based and different metal oxides, each class individually or their mutual association providing recent micro- and nano- adsorbents like microspheres of carbon, etc. (Zhu et al. 2019b) nanohybrids, nanomembranes, nano-sized sheets, engineered nanoparticles (Garg and Kataria 2016) synthesized by a variety of chemical and physical processes. After characterizing through different techniques as mentioned above in Sect. 13.2, these adsorbents were employed for heavy metal capturing from the water and this study was reported in the literature by Garg et al. (Garg and Kataria 2016) and herein, some of the data are shown in Tables 13.3a, 13.3b, and 13.3c.

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13.3.6 Adsorbents for Radioactive Metal Removal Removal procedures for radioactive metals (Vanhoudt et al. 2018) have beautifully reviewed the literature and placed a summary focused about the remediation of radioactive metals and other isotopes with nuclear hazards such as 239Pu, 241Am, 95 Zr, 131I, 95Nb, and many more by using photosynthetically active organisms/ biosorbents and thus created an extensive database for the upcoming researchers. Vanhoudt et al. (2018) have also listed some robust microorganism/biosorbents as per their performance. Despite the fact that a remarkable research work has been exercised for vanishing of radioactive metal uranium through the numerous compounds (mono) or their composites, metal-organic framework, other organicinorganic hybrids with and without carbon forms, some important gaps are still left out. More authentic research is needed with respect to precise size, shape, and capacity measurements for different adsorbents. An exhaustive literature review comprising of almost every class of adsorbent is positioned in (Prasad and Saxena 2004) to support the future scientists, as this review covers the recent computational methods like density functional theory and molecular dynamics tools too.

13.3.7 Metal–Organic Frameworks/Covalent Organic Frameworks as Adsorbents A novel class of composites known as the metal–organic framework and covalent organic framework is being explored recently (Tables 13.3a, 13.3b, and 13.3c), and these compounds show unique applications for metallic pollutants extortion from water. Zhu et al. (2019a) have presented a very intelligent and concise summary of such nanocomposites for toxic metal sequestration in the form of a review. Metal– organic and/or covalent organic frameworks, i.e., metal–organic frameworks/covalent organic frameworks after their fabrication, are being applied for various applications by very recent research scientists (Huang et al. 2018) including heavy metal trapping too. This type of frameworks (metal-organic) based on Zr metal with magnetic nature have been implemented by (Huang et al. 2018) or for an effective elimination of dyes as well as Pb (II) with some functionalized MOFs are also in practice. Metal–organic framework with amino group modified acted for Pb(II) more effectively. Further, covalent-organic frameworks (Zhu et al. 2019b) with amide group incorporated, possessing a skeleton type design obtained through ball milling polymerization of acyl chloride and amino groups have been witnessed as promising network for Pb(II) recovery that has again proved that amide-based frameworks are good trappers for lead ions. Even more, many researchers of the present times are working on composites of several metal–organic frameworks with carbon forms (carbon nanotubes, graphene oxide, etc.) to further improve the report card of metal–organic frameworks as well as carbon-based structures.

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13.3.8 Zeolites as Adsorbents Zeolites also construct a class for good adsorption of heavy metals due to the existence of porous structure within such materials. Removal performance for toxic heavy metals (Pb, Cd, and As) and disinfection for E. coli were performed using modified zeolite by an association of wood vinegar from oak, and the results are discussed beautifully in (Li et al. 2019b) with better efficiency. Adsorption of Cd (II), Pb(II), and Zn(II) ions onto plant sand of National Aluminium Company located in Orissa, India, was also examined (Mahapatra et al. 2009). Goethite, silica, and alumina were observed as the dominating phases in the sand as authenticated by X-ray powder diffraction pattern and Fourier transform infrared spectroscopy. Adsorption process followed pseudo-second-order kinetics for each ion. The trapping of metals thus obeyed the order Zn > Cd > Pb with the loading capacity by the sand as 47.79 mg g1, 42.2 mg g-1, and 22.8 mg g1, respectively. The x-ray powder diffraction motifs with different hkl planes present in the plant sand showed quite interesting features where the adsorption has disturbed the earlier patterns of (110) plane of goethite, (011) plane in silica phase. The adsorption phenomenon has also affected the crystallinity of the silica with the transformation of alumina phase Al2O3 to σ-Al2O3 phase during Cd capturing and Zn removal was indicated by the dominance of (110) plane, that of goethite after vanishing σ-A12O3 phase too. The underlying mechanism for the uptake of the wide range of bivalent metals that are Pb, Cu, and Zn on cost-effective phosphate picked from sedimentary rock has been reported by Prasad and Saxena (2004) via batch studies and film diffusion along with external diffusion models for adsorption kinetics, with the help of which, film transfer coefficients, as well as mass transfer constants, were determined. Consequently, it was concluded that these coefficients and constants were greatly influenced by initial metal concentration. A simple heterogeneous process was followed to fabricate functionalized conjugated β-keto-enol Furan identified through 13 C nuclear magnetic resonance spectroscopy, Fourier transform infrared spectroscopy, surface area detection by Brunauer–Emmett–Teller, elemental data, thermogravimetric analysis, scanning electron microscopy, etc. The adsorbent designed was tested to remove divalent ions, namely, Zn, Cu, Cd, and Pb from T-B and G-S rivers in Morocco real waters (Radi et al. 2015), and the results were quite satisfying toward rapid removal of these ions with equilibrium time of 25 min only and with more Cd selectivity during competitive extraction of metals.

13.4

Conclusion

The present book chapter is focused on heavy metal remediation that encapsulates a concise description of the facts and findings of quite recent works in order to update the modern researchers. It also condenses the important research performed in recent decades in this field. Several modes of preparation of various classes of materials, specifically the most versatile means to remove such metals, i.e., adsorbents, with

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suitable characterizations and applications through batch or column studies are highlighted. A brief summary of the adsorbent materials as nanohybrids, micro composites, nanocomposites, modified bio sorbents and zeolites, frameworks like metalorganic, covalent-organic, etc. is also prepared herewith Table 13.3a,13.3b, and 13.3c. More importantly, the latest computational techniques being adopted to predict the experimental findings prior to procuring them in the laboratories have also been presented. Such a kind of study has become a major attraction to get updated information about the adsorbent material being applied for the removal of such pollutants. Almost each heavy metal removal including radioactive metals, too, employed by the scientists has been included herein.

13.5

Future Challenges

Though the current environmentalists’ contributions are praiseworthy in the form of efficient results for heavy metal elimination, still, there are challenges to be met out such as the need for fast techniques and more economic and eco-friendly measurements with appropriate metal selectivity. This would be made available for each human being in the society with cleaner waters. To achieve this goal, some green procedures (solvent-free reaction extrusion) for the synthesis of the pristine materials (horizon nanomaterials) with rapid removal of heavy metals in the treated water and maximized restoration are yet to be exhausted. This need is more alarming particularly in developing countries as a demanding graph for clean water is moving uphill while its availability is going downhill. Acknowledgements Authors of the chapter are grateful to Deenbandhu Chhotu Ram University of Science and Technology, Murthal, Haryana, India, for providing necessary facilities and Professor Gurmeet Singh, University of Delhi, Delhi, for every kind of support for completion of the work.

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Chapter 14

Electroanalytical Techniques for the Remediation of Heavy Metals from Wastewater Muhammad Altaf, Naila Yamin, Gulzar Muhammad, Muhammad Arshad Raza, Munazza Shahid, and Raja Shahid Ashraf

Contents 14.1 14.2 14.3

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sources of Heavy Metals and Their Detrimental Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Determination of Heavy Metals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.3.1 Atomic Absorption Spectroscopy . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.3.2 Atomic Emission Spectroscopy . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.3.3 X-Ray Fluorescence . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.3.4 Ion-Selective Electrode Method . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.3.5 Electrochemical Sensing Method . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.4 Most Commonly Found Heavy Metals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.4.1 Arsenic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.4.2 Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.4.3 Chromium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.4.4 Lead . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.4.5 Copper . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.4.6 Nickel . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.4.7 Cadmium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.4.8 Zinc . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.5 Wastewater Remediation/Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.5.1 Conventional Method for Removal of Heavy Metals . . . . . . . . . . . . . . . . . . . . . . . . . 14.5.2 Adsorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.5.3 Membrane Filtration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.5.4 Electrochemical Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.5.5 Phytoremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.6 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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M. Altaf (*) · N. Yamin · G. Muhammad (*) · M. A. Raza · R. S. Ashraf (*) Department of Chemistry, GC University Lahore, Lahore, Pakistan e-mail: [email protected]; [email protected]; [email protected] M. Shahid Department of Chemistry, University of Management and Technology, Lahore, Pakistan © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature Switzerland AG 2021 Inamuddin et al. (eds.), Water Pollution and Remediation: Heavy Metals, Environmental Chemistry for a Sustainable World 53, https://doi.org/10.1007/978-3-030-52421-0_14

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Abstract This book chapter summarizes the sources of heavy metals in water, treatment of wastewater, and diseases caused by polluted water due to heavy metals. Commonly, human interaction with heavy metal can occur through various routes, which include ingestion through food or drinks and inhalation of dust or fume. Heavy metals such as lead, mercury, copper, cadmium, zinc, nickel, chromium, and arsenic cause growth problems, nervous system damage, cancer, organ damage, and in extreme cases death. In recent years, both the type and content of heavy metals in water have gradually increased due to fast global development and human activities. In this chapter, various water purification techniques such as chemical precipitation, lime coagulation, coagulation and flocculation, flotation, ion-exchange process, adsorption, membrane filtration, electrochemical treatment, and phytoremediation have been discussed in detail. Keywords Heavy metals · Physiochemical process · Electrochemical sensing · Adsorption · Electrodialysis · Ultrafiltration · Reverse osmosis · Ion-selective electrode · Phytoremediation · Ion-exchange process

14.1

Introduction

The specific density of heavy metals ranged from 3.5 to 7.0 g/cm3, are poisonous or toxic at low concentrations, have atomic number above 20, and are adversely affecting the environment and living organisms (Gautam et al. 2014; Shawai et al. 2017). Although, no particular definition of heavy metals can be found, literature has described it as a normally happening element having a high molecular weight and density multiple times more than water. Among 69 heavy metals, 16 are synthetic. Literature revealed that heavy metals are typically present in fewer amounts in waters; however, the majority of them, for example, cadmium, lead, arsenic, mercury, nickel, selenium cobalt, zinc, and chromium, are lethal even at very low concentrations (Raouf and Raheim 2016). Presence of trace amounts of heavy metals in water is essential to normalize metabolism. However, their higher concentration may be poisonous and pose health threats. An increasing amount of heavy metals in the environment is an area of more concern since a large number of industries are releasing metal-containing effluents into water without or with insufficient treatment. It was observed that the main harmful contents of water pollution are heavy metals (both in chemically combined and elemental form). Owing to human activities, balance of heavy metals like mercury, copper, vanadium (Gautam et al. 2014), cadmium, nickel, thallium, chromium, zinc, lead, and arsenic (Smedley and Kinniburgh 2002; Katsoyiannis et al. 2004) in water has tremendously changed which is highly hazardous to the organisms and has resulted in serious environmental deterioration (Gautam et al. 2014; Shawai et al. 2017). Heavy metals aggregate in the human body at low

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Fig. 14.1 Treating wastewater by primary and secondary method. Wastewater treatment by primary treatment (remove particulate organic material) and secondary treatment (remove the biological nutrient and dissolved organic material) (QRW, 150,000 m3/d; primary sludge, 4500 m3/d; QAS, 150,000 m3/d, QAS2: 3000 m3/d). (Image is modified with permission from Ref. Karvelas et al. 2003)

concentrations, they are nonbiodegradable which can be dangerous and may cause severe illnesses, for example, nervous system damage, cancer, and other diseases, at high concentrations (Men et al. 2018). Water containing heavy metals can cause severe harm to the central nervous system, liver, lungs, kidneys, and other fundamental organs. Long-term display may bring about harm muscular, neurological and physical degenerative that copy Parkinson’s disease, muscular dystrophy, Alzheimer’s disease, and numerous sclerosis. Moreover, long-term exposure to heavy metals may cause various types of cancers and allergies (Goel 2006; Saha et al. 2017). There are numerous strategies to eliminate metals from wastewater such as expelling huge particles from the sewage by methods for sedimentation or grids (primary treatment). Primary treatment is a physical process and reduces biochemical oxygen demand and suspended solids. In the primary treatment, the material that settles is called primary sludge (Karvelas et al. 2003). Secondary treatment by oxidizing agents will further decrease the biochemical oxygen demand level in the water (Fig. 14.1). It is likewise to include both heterotrophic protozoa and microscopic organisms like bacteria. The protozoa graze the bacteria and bacteria degenerate the organic compounds and transformed into CO2 and water. Secondary treatment can be carried out with oxidation ponds or trickling filters. Tertiary wastewater treatment generally aims to expel plant supplements such as phosphorus and nitrogen (Fig. 14.1). Ion exchange, chemical precipitation, electrochemical treatment, flotation, adsorption, and phytoremediation are generally used for removing heavy metals

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from wastewater (Saha et al. 2017). This book chapter summarizes water contamination by heavy metals, their sources, and various remediation methods.

14.2

Sources of Heavy Metals and Their Detrimental Effects

Soil erosion, mining, sewage discharge, urban runoff, natural weathering of the earth’s crust, industrial effluents, metal finishing and plating, textiles dyes, printed board manufacturing, street runoffs, pesticides, and landfills are important sources of heavy metals in fresh water (Kimbrough 2009). Several industries such as surface finishing, mines, pesticide, fuel and energy production, leather, steel and iron, electroplating, photography, aerospace, and atomic energy discharge their effluents into the environment including wastewater possessing different heavy metals (Gurel 2017). Heavy metals enter into the water by industrial waste and acidic rain, from breaking down soils, and discharging wastewater into rivers, groundwater, lakes, and streams. The pollution of water by heavy metals due to quick expansion in industrialization and urbanization has brought risks which cause various impacts on natural environment such as humans, animals, soils, and plants. Literature has revealed that oxidative deterioration of biological molecules like deoxyribonucleic acid and nuclear proteins is fundamental because of their binding with heavy metals. Heavy metals aggregate in the water and represent a hazard to human well-being, and some are cancer-causing, teratogenic, and mutagenic and can cause neurological and behavioral changes. Thus, heavy metal contamination remediation deserves attention (Goel 2006).

14.3

Determination of Heavy Metals

Different techniques have been utilized for determination of heavy metals including atomic emission spectroscopy, X-ray fluorescence, inductively coupled plasma mass spectrometry, and atomic absorption spectroscopy (Al-Saydeh et al. 2017). Although, these techniques are valid in terms of accuracy, they are still not considered as alternatives as they are not economical. All of these techniques take time and need an experienced person for sample handling, apart from their complicated sample preparation. To avoid such complications, simpler and automated techniques such as ion-selective method and electrochemical sensing are preferred (Naeemullah et al. 2016).

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14.3.1 Atomic Absorption Spectroscopy In this technique, atomization of the sample takes place by using flame or graphite furnace and absorption of light by metal ions in the sample. The method takes into consideration examination of one element at any given moment, and the solid sample should be processed in aqueous solution preceding the investigation (Tang et al. 2018).

14.3.2 Atomic Emission Spectroscopy In this technique, excite the elements in a watery sample by high-temperature flame or plasma and analyze the radiated light as it falls to the ground (Tang et al. 2018).

14.3.3 X-Ray Fluorescence X-beam or ray fluorescence includes utilizing X-beams to trigger excitation of inner shell electrons to higher energy levels before decomposing and discharging energy as fluorescence, which is determined and is a characteristic of targeted heavy metal (Li et al. 2016). The previous lab instrument has exceptional sensitivity and limit of detection at lower part per billion levels. However, expensive and usually large instrumentation is needed with the expert person for the operation of testing tools. Consequently, these systems are therefore limited to laboratory testing of prepared water samples. Instead of the availability of portable field X-ray fluorescence tools with elevated detection limits, the use of X-ray fluorescence for large-scale heavy metal assessment can be time-consuming and costly (Tang et al. 2018).

14.3.4 Ion-Selective Electrode Method Ion-selective electrode is an electrochemical sensor that transforms ionic signal into electronic signal. For potentiometric detecting, ion-selective electrode ordinarily needs the utilization of a reference anode for determining the concentration of target particles electrochemically. For the synthesis of ion-selective electrodes, many inorganic, chelating, organic, composite, and intercalating substances have been considered as electroactive material. Electrochemical detection by this is preferred owing to its high sensitivity, short analytical time, low power cost, superior performance, and simple versatility for in situ and online estimation. The technology can monitor trace concentrations of Pb in drinking water. Taking into account the type of membrane material, Pb-particular electrodes can be ordered in three vital groups

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including polycrystalline, polymeric, and glass membrane ion-selective electrode (Tang et al. 2018). Ion selective strategy was fruitful in defeating the disadvantages of analytical approaches, yet it had a disadvantage of being particular.

14.3.5 Electrochemical Sensing Method Electrochemical detecting technique was utilized for estimating and checking heavy metal ion in fresh waters and for the assessment of potentiometric sensor crosssensitivity in fluid or liquid media. The issue of ion-specific detecting strategy was overwhelmed by electrochemical sensing, the nonselective methodology used by Faraday’s law to get electrical estimations for the chemical procedure. An electrochemical system’s feedback on a connected potential is estimated as an electrochemical impedance (Karkra et al. 2016). Over time, electrochemical sensing has been utilized to distinguish water quality for beverages such as tea and red wine (Gutierrez et al. 2011). Electrochemical sensing is one of the generally utilized strategies that gives the physic-synthetic data of contaminated samples (MartınezMáñez et al. 2005).

14.4

Most Commonly Found Heavy Metals

14.4.1 Arsenic Arsenic is a metalloid with an atomic number of 33, which not only occurs as pure elemental form but also in combination with metals and sulfur in over 160 minerals. Groundwater contamination by arsenic is found in numerous countries all over the world. Pollution of the water with arsenic both from anthropogenic and natural sources has happened in numerous parts of the world and seen as a worldwide issue. Primary anthropogenic sources of arsenic comprise geogenic/natural processes, metal base smelters, wood preservatives, dyes, gold mines, automobile exhaust/ industrial dust, power plants using arsenic-rich-treated wood or coals, municipal and industrial dump sites, and disposal sites for wastes from arsenic-processing plants (Katsoyiannis et al. 2004). A report found that more than 137 million individuals of 70 countries including Pakistan, India, the United States, China, and Bangladesh are most likely affected by arsenic harming of drinking water (Nriagu et al. 2007; Smedley and Kinniburgh 2002). Excess amount of arsenic present in drinking water caused human poisoning and death. Water polluted with arsenic causes respiratory, neurological, skin, and cardiovascular diseases along with bladder, lung, and kidney cancers. Arsenic assembles in keratin-rich tissues such as the skin, nails, and hair. Inorganic compounds of arsenic can cause tumor of the skin and respiratory system and neoplastic sores in different organs (Nriagu et al. 2007). Arsenic can enter the body from the GIT

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(gastrointestinal tract), skin, and respiratory system and obstruct the activities of some proteins and enzymes by developing affinity with them and disturbing the Krebs’s cycle. Literature also revealed that inorganic arsenic compounds such as As2O3 and AsH3 are more harmful than the organic ones. Investigations disclosed that 70–300 mg of arsenic trioxide is the normal deadly dosage for humans. The key impacts of arsenic in human beings are major alterations in the mucous membranes and destruction of the skin and peripheral nervous system. Arsenic related skin issues include xerosis, hyperkeratosis of palms and soles, and skin aggrevation with infection or ulceration (Hubicki and Koodynsk 2012). The World Health Organization in 1958 has declared arsenic as a poisonous element and established its upper limit to 0.01 mg/dm3. But in 1993, the maximum limit of arsenic prescribed by the World Health Organization, European Union, and US Environmental Protection Agency, was fixed to 0.01 mg/L (World Health Organization 1993). The Food and Agriculture Organization in its rules has additionally set 0.01 mg/dm3 as the most extreme limit of arsenic. Different treatment methods, for example, ion exchange (Korngold et al. 2001), chemical precipitation (Magalhães 2002), bioremediation (Katsoyiannis et al. 2004), coagulation, filtration (Wickramasinghe et al. 2004), electrocoagulation (Parga et al. 2005), reverse osmosis (Mondal et al. 2006), and adsorption are utilized to eliminate arsenic from water (Mohan and Pittman Jr. 2007). Most of these strategies have few disadvantages including sludge production, the need of costly equipment, complexity, and extra time consumption. Among the above treatment methods, coagulation and adsorption are accepted because they are low-cost methods, although coagulation requires trained operators and displays greater viability on just a few particular aluminum salts and iron. However, adsorption, because of its simplicity in design, low cost, regeneration potential, and sludge-free and easy operation, has gained fame. In addition, adsorption-based strategies are more efficient than any other method in removing arsenic to a reduced level (Kango and Kumar 2016) and display easy separation of tiny quantities of harmful components or elements from huge quantities of water. Many adsorbent types, for example, granular ferric hydroxide, ironbased sorbents, activated alumina, iron-coated sand, zero-valent iron (Hussam and Munir 2007; Su and Puls 2001, 2003), magnetic ion-exchange resins (Sinha et al. 2011; Lackovic et al. 2000), synthetic goethite (Lakshmipathiraj et al. 2006), chitosan-based adsorbents (Gupta et al. 2013; Elwakeel and Technology 2014), graphene oxide composites (Chandra et al. 2010; Kemp et al. 2013), nanoscale magnetite-coated sand, and magnetic nanochains (Das et al. 2014), are generally utilized for arsenic exclusion from water. Due to their expanded surface area and small size, the use of nanoparticles in ecological remediation processes has been recently extended (Dixit and Hering 2003; Kango and Kumar 2016). Among various magnetites (Fe3O4), iron oxide nanoparticles are potential adsorbents for arsenic removal due to their high adsorption capacity (Nriagu et al. 2007).

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14.4.2 Mercury Mercury or quicksilver is a silvery white liquid metal with an atomic number of 80, boiling point of 356.73
C, and freezing point of 38.83
C. Hg is a highly toxic metal that usually occurs in the form of mercuric sulfide. Mercury poisoning could result not only from mercury element but also from mercuric chloride or methyl mercury. Inorganic compounds of mercury can amass in the drinking water, causing health issues. Mercury in the organic form turns out to be highly risky when converted to ethyl, methyl, and butyl mercury. However, the value in industrialized territories could reach around 700 ng/dm3 (Hubicki and Koodynsk 2012). Mercury is incorporated into wastewater via different sources such as in industries like chlor-alkali plants, fluorescent lamps, household products, hospital waste (sphygmomanometers, barometers, damaged thermometers), thermal power plants, adhesives, paints, and electrical appliances. Among these, the most threatening source to fresh water and aquatic life is in chlor-alkali industry. Volcanoes are among the naturally occurring sources of mercury that produce practically 50% in the atmosphere. The other half is contributed by human activities, for example, burning (65%), generation of gold (11%) and nonferrous metal (6.8%), production of cement (6.4%), and waste disposal (3.0%) including municipal waste (Goel 2006). Mercury in the form of organic and inorganic compounds is neurotoxic and could be lethal, sometimes with severe effects when absorbed by the skin of the body. Moreover, Hg can cause chronic as well as acute poisoning. Mercury could amass in the stomach and stay nonabsorbable, leading to the development of cancercausing diseases. Long-term mercury contact may trigger severe lung irritation, nerve damage, renal failure, vomiting, skin rashes, and diarrhea (Manohar et al. 2002). The American Association of Poison Control Centers reported 3596 cases in 1997 of acute Hg poisoning. Mercury compound, like methylmercury, caused microtubule destruction, lipid peroxidation, mitochondrial damage, and aggregation of neurotoxic molecules like aspartate, serotonin, and glutamate (Jaishankar et al. 2014). Inorganic compounds of mercury caused intrinsic abnormality, unconstrained premature birth, and gastrointestinal issue, such as hematochezia and destructive esophagitis (Jaishankar et al. 2014). Organic compounds of mercury such as dimethylmercury caused erethism (Goel 2006). Ingestion of mercury salts could also result in salivation, bloody diarrhea, burning throat, kidney damage, and vomiting. The concentration of mercury vapors above 1.0 mg/m3 could cause pneumonia, mental disorders, inflammation of the gums, and damage of the lung tissues (Hubicki and Koodynsk 2012). The Clean Water Act established mercury limit at a level of 0.001 mg/L or 1.0 ppb for discharge of mercury to protect human health, fish, and wildlife (Ekinci et al. 2002). Naturally occurring level of Hg in surface water and groundwater is less than 0.5 μg/L. A small number of shallow wells and groundwaters surveyed in the Unites States were revealed to have mercury levels that surpassed the maximal impurity level of 2 ppb or 2.0 μg/L set by the US Environmental Protection Agency for drinking water (Tzanetakis et al. 2003; Okpalugo et al. 2005).

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Expelling mercury from the water can be accomplished by utilizing various techniques such as filtration, coagulation, adsorption (using granular activated carbon), biofilms, lime softening, reverse osmosis, and chemical precipitation. Coagulation/filtration is a typical treatment which uses ferric chloride or aluminum sulfate to remove mercury from water (Raouf and Raheim 2016; Hubicki and Koodynsk 2012).

14.4.3 Chromium Chromium is a steely dark gray, hard, lustrous, and among the most harmful, highly cancer-causing heavy metal. Chromium is a fundamental component needed for carbohydrate metabolism in small quantities such as glucose or starch digestion and fat and protein metabolism, yet becomes dangerous at higher concentrations. Chromium occurs from Cr(2) to +6 in different oxidation states; however, two common forms, Cr(III) and (VI), are important as far as environmental pollution is concerned (Berihun 2017). Chromium(VI) is observed to be too toxic because of its cancercausing nature. It triggers teratogenic processes and can cause malignant cancers and disturbs deoxyribonucleic acid synthesis. Discharge into the atmosphere by human activities of heavy metals like chromium, which are significant pollutants in soil and surface waters, has risen enormously owing to industrialization, which impacts the geochemical cycle (Berihun 2017). Natural sources of chromium comprise weathered rocks, volcanic emissions, and biogeochemical procedures. Chromium is a significant and widely applied element in the industry. Anthropogenic sources of chromium contamination include petroleum refining, leather tanning, electroplating industry, mining, textile industries, chrome plating, metal processing, nuclear power plants, chromium salt manufacturing such as chromate salt preparation, paints and pigments, dyeing, steel fabrication, textile manufacturing, and pulp processing units (Goel 2006). Drinking of contaminated water caused accumulation of chromium in the kidneys and liver. In the lungs, high levels of chromium are found and deposited as insoluble compounds. Binding chromium with blood components and carrying chromium through the blood primarily relies on its valency. Except the liver, Cr(III) cation dominates in many tissues. Hexavalent chromium quickly crosses red blood cell layers and reduced to the trivalent state. The interaction with DNA initiates cancercausing chromium. Clinical indications of serious harm by chromium compounds are described by vomiting, abdominal pain, kidney damage, gastrointestinal ulceration, and bloody diarrhea (Hubicki and Koodynsk 2012). Hexavalent chromium is water soluble which makes more it hazardous, carcinogenic, and mutagenic. Antagonistic impacts of the hexavalent form on the skin may comprise ulcerations; dermatitis; gastric damage; kidney, liver, and lung cancers; and hypersensitive skin reactions. Most industries use chromium compounds in an endeavor to enhance human expectations for everyday comforts, but releasing these chemicals into the environment without

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adequate treatment turns around the expectations for everyday comforts that are planned. The most significant contributors of chromium pollution are leather tanning industries (Berihun 2017). Permissible level of chromium in groundwater is 0.02 mg/L. The upper limit of chromium in drinking water is 0.05 mg/L (Goel 2006). Adsorption strategy is well-known encouraging method and is regarded to be highly strong financially and operationally extremely powerful to expel heavy metals from wastewater, particularly cheap and high productivity adsorbents. Numerous adsorbents such as wheat grain, rice hull, peel of banana, wheat (Triticum aestivum), activated carbon, henna leaves, and so forth have been affirmed powerful for chromium expulsion by formation of chelates (Mitra et al. 2017; Song et al. 2016). Other less economical methods are coagulation, solvent extraction, filtration, precipitation, reverse osmosis, biosorption, and ion exchange (Raouf and Raheim 2016; Goel 2006).

14.4.4 Lead The atomic number of lead is 82 and naturally present in the earth’s crust, and its dangerous concentration in industrial and natural wastewater is currently a worldwide problem. Pb is available in various minerals, the important ore being Galena (PbS), and goes into the body via water, food, and air and is accumulated in essential organs of animals and human. Pb is utilized to produce pigments, fuels, leaded glass, matches, photographic materials, explosives, and batteries (Goel 2006). Various industrial wastewater effluents, like those from ceramics, lead–acid batteries, mining, leaded gasoline, electroplating, mobile batteries, bangle industry, lead smelting, petrol-based materials, electronic waste, brass corrosion, paints, metal finishing, pesticides, and coal-based thermal power plants, release critical amounts of lead ion in water bodies (Kimbrough 2009). Lead is a lethal and cancer-causing metal in nature. Inorganic compounds of lead affect the central nervous system, gastrovascular cavity, kidneys, reproductive system, and gastrointestinal tract. A remarkably serious impact of lead poisoning is its teratogenic impact. Lead poisoning inhibits the ability to produce hemoglobin and triggers hematological damage, anemia, kidney dysfunction, and damage to the central nervous, cardiovascular, and reproductive systems (Goel 2006). Chronic contact to lead causes extreme sores in the liver, lungs, kidney, and spleen. Pb-based inorganic compounds seldom cause severe harm to live organisms. The intense type of lead harming is spasticity of inner organs, and the neurological harm in peripheral organs is known as lead colic. However, in case of an occurrence of intense harming in human, the symptoms are vomiting, oliguria, diarrhea, proteinuria, high blood pressure, abdominal cramps, central nervous system damage, and hematuria. Tetraethyl lead are alkyl organic Pb compounds which are more poisonous than inorganic compounds. In youngsters and adults, the poisonous impact of Pb is observed in the brain and its poisoning damage in the peripheral nervous system. Characteristic signs comprise blue–black border on

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the gums and pale gray skin color (Hubicki and Koodynsk 2012). As per the World Health Organization standards, its extreme level in water is 0.01 mg/L, and the most significant release limit is 0.5 mg/L (Raouf and Raheim 2016; Hubicki and Koodynsk 2012). Different procedures for removal of lead are greatly used like ion exchange, adsorption, reverse osmosis, precipitation, coagulation, membrane separation, and electrodialysis; however, adsorption is the most versatile and broadly utilized technique because it has many advantages over other methods including recovery of metal value, sludge-free operation, selectivity, and cost viability. Adsorption is an effective method which can remove lead from wastewater up to 95% by using appropriate adsorbents. Different commercial adsorbents and biomaterials show great adsorption efficiency for removal of Pb(II) ion from olive cake, sea nodule, bone powder, tea waste, polymerized banana stem, horticultural waste (palm and coconut shell carbon), phosphogypsum, activated carbon, bentonite, and carbon aerogel (Mondal 2008). Adsorption of Pb(II) ions on sulfur-treated activated carbon is higher due to the greater affinity of lead with sulfur (Tripathi and Ranjan 2015; Raouf and Raheim 2016). Moreover, weakly basic anion exchangers can be utilized for particular expulsion of lead chloride from the solutions of pH ranging from 4 to 6 (Mondal 2008).

14.4.5 Copper Copper (having an atomic number of 29) is usually found in the earth’s crusts, as sulfides, for example, chalcopyrite, bornite, covellite, and chalcocite. Even at low concentrations, copper is a harmful transition metal, and copper-polluted water must be processed before releasing into the atmosphere. Copper is released in water from various sources such as mining, metal refining, plastic and electroplating industry, smelting operations and plating baths, paints and pigments, petroleum refining, brass corrosion, fertilizers, printed circuit board production, and wood pulp (Al-Saydeh et al. 2017). Copper is also a vital life element, being an essential part of enzymes and human blood. The assessed adult dietary intake of copper is somewhere in the range of 2–4 mg/day. The requirement of Cu is expanded in kids, pregnant ladies, and old peoples. Cu is involved in various procedures such as oxidation–reduction in the body, the activity of hemoglobin acting as stimulant, hair keratinization, melanin synthesis, solidifying of collagen, and lipid metabolism. Cu is principally gathered in the mitochondria, ribonucleic acid, deoxyribonucleic acid, and nucleus of animal cells. Copper promptly forms an association with different proteins, particularly those of sulfur-containing amino acids. Although copper is a fundamental metal, still it can lead to poisonous impact along with gastrointestinal tract disturbances and liver damages, for example, Wilson’s disease and cirrhosis which are characterized by a gathering of granules of copper inside the liver. Moreover, absorption of copper causes blockage of the nasal mucosa, diarrhea, sore throat, gastritis, inflammation,

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corneal opacity, lung damage, conjunctivitis, ulceration, itching, nasal congestion, and nasal septum. The copper content in river and saline waters ranges from 0.9 to 20 and 0.02 to 0.3 μg/dm3, respectively (Hubicki and Koodynsk 2012). According to the World Health Organization, the maximum amount of copper in industrial effluents should not exceed to 1.3 mg/L. Similarly, the World Health Organization expressed that copper ion content in drinking water should not surpass to 2.0 mg/L (Al-Saydeh et al. 2017). A large number of water treatment strategies have been utilized for the expulsion of copper ions including adsorption, ion exchange, chemical precipitation, coagulation–flocculation, flotation, electrodialysis, photocatalysis, membrane filtration, and electrochemical technologies (Al-Saydeh et al. 2017).

14.4.6 Nickel Nickel is comparatively a less toxic metal than other heavy metals. It is commonly found in the earth’s crust, water, soil, food, and air. Nickel has a wide range of applications in stainless steel, and electroplating represents 11% of the total nickel processing (Djaenudin et al. 2017). The wastewater produced from nickel processing still contains a significant amount of nickel that should be expelled before the final release to the environment (Djaenudin et al. 2017). Nickel is present in crude oil (50–350 mg/kg), seams of coal (4–60 mg/kg), and creases of coal (4–60 mg/kg) (Hubicki and Koodynsk 2012). Nickel sources are; industrial methods like plastics manufacturing, connector, pigments, galvanization, nickel cadmium batteries, lead frame, tableware, metal finishing, electroplating, super phosphate fertilizers, paints, thermal power plants, smelting operations, burning of coal, diesel fuels and mining and metallurgical operations. Moreover, brass is also a major source of nickel in wastewater (Kimbrough 2009). Acidic rain expands versatility of nickel in dirt or soil and is the major cause of increased nickel concentrations in groundwater. Major diseases caused by nickel are, kidney and lung cancers, vomiting, nausea, gastrointestinal tract disorders, diarrhea, asthma, pulmonary fibrosis, renal edema, conjunctivitis, and skin irritation, aggregated for the most part in bones, the heart, and different organs, are caused by ingestion of nickel in higher concentration through the water. Given the fact that nickel is a poisonous element, it was suggested that it is essential for animals and plants and for human nourishment. Nickel ingested with water and food is inadequately assimilated and quickly discharged from the body. Nickel inhalation by air is to a great extent aggregated in the lungs. Lethal or intense harming of Ni isn’t seen. Tetracarbonylnickel, which occurs mostly in nickel refineries, is the most hazardous. Day by day, human nickel absorption ranges from 0.3 to 0.5 mg. In human beings, nickel intake from the gastrointestinal tract is less than 10% (Hubicki and Koodynsk 2012). In municipal and industrial wastewater, the content of the metal varies from 20 to 3924 mg/kg (Hubicki and Koodynsk 2012). It is expected that the nickel concentration in river water must be around 1 μg/

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dm3. A lot of nickel are transferred from municipal wastewater to surface water where the concentration surpasses 3000 ppm (Hubicki and Koodynsk 2012). The acceptable level should be 20 μg/dm3 (Hubicki and Koodynsk 2012). As far as possible, the limit suggested by the World Health Organization and Food and Agriculture Organization is 0.02 mg/dm3 (Hounslow 2018). Nickel is removed from wastewaters using several techniques such as adsorption, reverse osmosis, solvent extraction, ion exchange, coagulation, flocculation, biosorption, electrocoagulation and electrodialysis (Djaenudin et al. 2017), floatation, and membrane separation. The decision of a specific strategy for the removal of the metal relies on concentration in wastewater, cost, nature of effluent, percentage removal required, and accessibility of the equipment (Djaenudin et al. 2017).

14.4.7 Cadmium Cadmium is a metal of great toxicological study due to its nonbiodegradable and bioaccumulation nature even at very small concentrations. Whenever ingested, cadmium causes chronic or acute dangerous impacts. Cadmium is commonly utilized in the manufacture of phosphate fertilizers, pigments, nickel–cadmium batteries, amalgams, or alloys and metal plating (Kumar et al. 2015). Water is polluted by cadmium via e-waste, zinc smelting, paint pigments, battery waste, sludge, fuel combustion, refined petroleum products, galvanized pipes, corrosion of galvanized pipes, pesticides, copper refineries, and plastics. Cadmium might be brought into water typically by volcanic ejection. Water and food are an important cause of human interaction with cadmium, particularly for the resident population of the industrial region from which cadmium is released. Cadmium is mainly absorbed through inhalation, and the gastrointestinal tract adsorbs less than 10% of cadmium (Hubicki and Koodynsk 2012). Cadmium is a popular nephrotoxic agent (Wu et al. 2008). In humans, cadmium has a lengthy half-time of 10–30 years (Swaddiwudhipong et al. 2012) and aggregates essentially in renal cells with advancement of chronic disease. Long-term presentation of Cd in humans results in lung disease, renal dysfunction, and tubular proteinuria. It’s portrayed by cough with bloody and foamy sputum, chest pain, and pulmonary tissue death due to watery fluid aggregation. Cadmium is also linked to bone imperfections (osteoporosis, osteomalacia, and unconstrained breaks), expanded blood pressure, and myocardial dysfunctions. Cd amasses in kidneys, pancreas, digestion tracts, and organs that alter the digestion of vital components in the body, for example, iron, copper, zinc, calcium, magnesium, and selenium (Raouf and Raheim 2016). Kidney and respiratory destructions are the basic impacts of cadmium compounds on human beings. Cd likewise causes impedance of smell, pain in the spine, changes in the skeletal framework, trouble in walking, and the development of hypochromic anemia. Long-time exposure to cadmium and extreme intake of cadmium cause genuine sickness, for example, itai-itai illness in individuals. Furthermore, cadmium

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has been recognized as a human cancer-causing agent (Mahalik et al. 1995). Cadmium occasionally causes diabetes, liver disease, hypertension, nervous system damage, and urinary stones (Fowler 2009). According to the Environmental Protection Agency and World Health Organization, the average concentration of Cd in water should be below 0.001 and 0.003 mg/ dm3, respectively (Gray 2008; WHO 1993). Several strategies are accessible for the removal of cadmium ions from wastewater. Some important strategies incorporate solvent extraction, ion exchange, precipitation, adsorption, and reverse osmosis. Activated carbon adsorption and carbon-based nanosorbent adsorption are well-known methods for the removal of cadmium (Fowler 2009).

14.4.8 Zinc Zinc is a shiny blue–white metal, with an atomic number of 30. Zn has reasonable reactivity that will allow it to bind with oxygen and dilute acids and other nonmetals. Zinc occurs in nature in the form of willemite, sphalerite, smithsonite, franklinite, zincite, and so forth. Zinc metal is used in galvanization of steel and some alloys, in construction of the cathodic plate in electrical batteries, and as a pigment in scanner, plastics, cosmetics, inks for printing backdrop, and so on, while in the rubber industry, it acts as a catalyst during synthesis and as a heat disperser. Zinc metal is incorporated into most single tablet and is recognized as having antioxidant characteristics that protect the muscles of the body and skin from aging (Hubicki and Koodynsk 2012). Zinc is generally utilized in industries, for example, paint, galvanization, batteries, pigment, smelting, electroplating, polymer, detergents, rubber, paints, dyes, ointments, fertilizers, and pesticides industries. Zinc is generally considered as nontoxic, particularly whenever taken orally. Nonetheless, higher amount can cause dysfunctions of the reproductive system and stop its growth. Higher concentration of zinc metal can cause harm to numerous biochemical processes and can cause disorders in the kidneys, liver, and gonads. Moreover, high level of zinc cause indistinguishable indications of illness caused by Pb and can mistakenly analyzed as Pb poisoning. Kidneys play a significant role in keeping up zinc homeostasis in the body. Zinc toxicosis clinical signs were reported as diarrhea, lethargy, liver failure, vomiting, depression, bloody urine, icterus, anemia, kidney dysfunction, neurological signs, and increased thirst. Zinc is a fundamental metal and is the major component of human diet as its day-by-day prerequisite ranges from 10 to 20 mg and necessary for carbohydrate and protein metabolism; however, high amount of zinc dosage causes various diseases. Laborers who are exposed to zinc fumes experience health issues like high fever, shuddering, and dryness of mouth. Zn compounds are destructive to the skin, eyes, and mucous membranes, resulting in a type of dermatitis known as “zinc pox” (Neetesh and Lal 2018).

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There are a wide range of strategies for removal of zinc such as adsorption, reverse osmosis, electrodialysis, ion exchange, electrochemical treatment, membrane separation, ultra-nanofiltration, chemical precipitation, solvent extraction, and biological methods. The biological approach is among these, the best for zinc removal because the method has decreased necessity for chemicals, low operating costs, and high efficiency and is eco-friendly and non-generative of toxic sludge. These techniques likewise offer conceivable outcomes for metal recovery and biosorbent recovery afterward (Arif Tasleem Jan 1 2015). Table 14.1 summarizes sources of heavy metals and methods of treatment.

14.5

Wastewater Remediation/Treatment

Purification of wastewater containing heavy metals needs exceptional consideration to secure water, air, and soil quality and the health of animals and human beings all over the world (Kobielska et al. 2018).

14.5.1 Conventional Method for Removal of Heavy Metals The conventional methods for removal of heavy metals from water involve reverse osmosis, chemical precipitation, and lime coagulation. The details of water purification techniques are given below. Physicochemical Processes

Chemical Precipitation Chemical precipitation by coagulation and flocculation and sedimentation has commonly been utilized for a long time to treat wastewater. Precipitation is used mainly to convert dissolved metals into particulates that can be isolated by coagulation and sedimentation or filtration. The method is widely utilized for removing heavy metals from wastewater by converting dissolved materials into solid particles. The method is used for chemical precipitation of sulfates, sulfide, carbonates, lime, and hydroxides. It is also utilized to expel ionic components from water by the inclusion of counterions to decrease solubility metal ions. The strategy is generally used for the removal of cations of heavy metal as well as anions, for example, cyanide, fluoride, and phosphate (Wang et al. 2008). By insertion of coagulants, for example, ferrous sulfate, lime, ferric chloride, potash alum, polyaluminum chloride, iron salts, and other organic polymers, precipitated metals are separated by filtration (Shawai et al. 2017). However, the process depends on the kind of metal and reagent used along with the concentration of the metal. General scheme for heavy metals is given below (Eq. 14.1) (Wang et al. 2008).

Sources Metal smelters, geogenic/natural processes, wood preservatives, dyes, gold mines, automobile exhaust/industrial dust, power plants, industrial and municipal dump sites

Industries; volcanoes; fluorescent lamps; household products; thermal power plants; hospital waste; damaged thermometers, barometers, and sphygmomanometers; and electrical appliances

Volcanic emissions, weathered rocks, industries, mining, nuclear power plants

Metal Arsenic

Mercury

Chromium Ingestion, inhalation, and absorption through the skin

Absorption through the skin, ingestion, inhalation

–

Required for carbohydrate, certain protein, and fat metabolism

Entry in the body Ingestion, inhalation

Importance in the body –

Neurotoxin, cancerous diseases, nervous system damage, lung irritation, kidney disease, skin rashes, eye irritation, rheumatoid arthritis, diarrhea, vomiting, loss of hearing and muscle coordination Headache, carcinogenic, abdominal pain, gastrointestinal tract ulceration, kidney failure, dermatitis, ulcerations,

Toxicity effect Skin manifestations; human poisoning; respiratory, cardiovascular, and neurological diseases; bladder, lung, renal, and visceral cancers; vomiting; anorexia; and nausea Coagulation, filtration, adsorption, chemical precipitation, reverse osmosis

Adsorption method, coagulation, precipitation, biosorption, filtration, ion exchange, reverse osmosis, and solvent extraction

0.05 mg/ L

Removal methods Chemical precipitation, ion exchange, bioremediation, coagulation, filtration, electrocoagulation, reverse osmosis, and adsorption

0.001 mg/ L

Amount permitted in water 0.01 mg/ L

Table 14.1 Summary of the sources, entry in the body, and toxicity effect of heavy metals and methods of treatment

Arif Tasleem Jan (2015) and Raouf and Raheim (2016)

References Raouf and Raheim (2016), Katsoyiannis et al. (2004), Magalhães (2002), Wickramasinghe et al. (2004), Parga et al. (2005), and Mondal et al. (2006) Raouf and Raheim (2016), Manohar et al. (2002), and Hubicki and Koodynsk (2012)

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Industries, ceramics, electroplating, lead– acid batteries, mining, leaded gasoline, mobile batteries, petrol-based materials, electronic waste, brass corrosion, paints, pesticides, coal-based thermal power plants

Mining, industries, paints and pigments, brass corrosion, wood pulp, fertilizers, paperboard, and printed circuit board production

Industrial processes, superphosphate fertilizers, paints, thermal power plants, combustion of coal and liquid fuels primarily by diesel engines, and mining and metallurgical

Lead

Copper

Nickel

Part of several enzymes and components of human blood. Stimulant on the amount and activity of hemoglobin, hair keratinization, hardening of collagen, melanin synthesis –

–

Inhalation and ingestion

Inhalation, absorption

Inhalation, ingestion, absorption through skin

Damage to lungs and kidneys, gastrointestinal tract disorders, pulmonary fibrosis, asthma, conjunctivitis, renal edema, skin irritation

allergic skin reactions Carcinogenic; central and peripheral nervous system impairment; renal, gastrovascular, gastrointestinal, and reproductive system disorders; hematological damage; anemia; and renal failures Insomnia, Wilson’s disease, congestion of the nasal mucosa, gastritis, diarrhea and chronic lung damage, conjunctivitis, ulceration, corneal opacity

0.02 mg/ dm3

2.0 mg/L

0.01 mg/ L

Djaenudin et al. (2017), Kimbrough (2009) and Raouf and Raheim (2016)

Adsorption, ion exchange, coagulation and flocculation, solvent extraction, chemical reduction, reverse osmosis, biosorption, electrocoagulation

Electroanalytical Techniques for the Remediation of Heavy Metals from Wastewater (continued)

Al-Saydeh et al. (2017) and Raouf and Raheim (2016)

Kimbrough (2009) and Raouf and Raheim (2016)

Chemical precipitation, membrane filtration, ion exchange, adsorption, flocculation, coagulation–flotation, and electrochemical technologies

Ion exchange, reverse osmosis, precipitation, coagulation, adsorption, membrane separation and electrodialysis, adsorption

14 487

Electronic waste, zinc smelting, waste, Cd– Ni batteries, paint, sludge, incinerations, volcanic eruption, refined petroleum products, paint pigments, pesticides, plastics

Smelting, batteries, wood preservatives, electroplating, detergents, phosphate fertilizers, dyes, paints, and ointments

Zinc

operations, brass, acid rain

Sources

Cadmium

Metal

Table 14.1 (continued)

Metabolism of proteins and carbohydrates. Antioxidant properties, delayed premature aging of the skin and muscles

–

Importance in the body

Ingestion and inhalation

Entry in the body

Nephrotoxic, renal dysfunction, obstructive lung disease, cadmium pneumonitis, bone defects, increased blood pressure and myocardial dysfunction, irritation of respiratory tract, hypochromic anemia Vomiting, diarrhea, lethargy, impairment of reproduction and growth, bloody urine, depression, icterus, liver failure, kidney failure, anemia

Toxicity effect

0.003 mg/ L

Amount permitted in water

Adsorption, ion exchange, reverse osmosis, electrodialysis, electrochemical treatment, membrane separation, ultrananofiltration, chemical precipitation, solvent extraction, and biological methods

and electrodialysis, floatation, membrane separation Solvent extraction, precipitation, ion exchange, reverse osmosis, and adsorption

Removal methods

Hubicki and Koodynsk (2012) and Neetesh and Lal (2018)

Wu et al. (2008), Gray (2008), WHO (1993), Fowler (2009), Hubicki and Koodynsk (2012) and Mahalik et al. (1995)

References

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M2 þ 2OH ! MðOHÞ2 #

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ð14:1Þ

Here, (M2+ ¼ heavy metals and OH counterions), and M(OH)2 represents insoluble metal hydroxide. A new kind of coagulant, polyelectrolyte amphoteric in nature, was synthesized by grafting sodium xanthogenate to polyethylenimine (Wang et al. 2005) that efficiently removed both dissolvable heavy metals and insoluble substances. Though with the above-mentioned coagulant, negatively charged colloidal substances can also be coagulated even at lower pH; however, the cationic nickel ion can’t be removed efficiently. Generally, at higher pH, the turbidity of solution decreases, and nickel removal increases. Precipitation is frequently used for heavy metal removal from wastewater. Silver, calcium, and barium ions are precipitated by chloride, oxalate, and sulfate ions, respectively.

Lime Coagulation Lime precipitation is a commonly used method in most of the countries worldwide because it is an economical, easily available, convenient, and safe operational method, and it has a simple process. Lime precipitation can be used adequately to treat inorganic effluents such as coil coating, metal finishing, ore mining and dressing, battery manufacturing, porcelain enameling, inorganic chemical manufacturing, organic and inorganic wastes, non-ferrous metal manufacturing, copper coating, paint and ink formulation, aluminum forming, textiles, steam electric power plants, auto, and laundries, where the concentration of the metal is above 1000 mg/dm3 at basic pH (9–11) (Wang et al. 2008). The methodology is efficiently used for removal of iron (Fe+2) and chromium (Cr+3) from wastewater (Wang et al. 2008; Wang et al. 2005). However, the process is ineffectual when metal ion concentration is low. A large amount of sludge production is another drawback of the technique.

Coagulation and Flocculation Flocculation is the process in which small suspended particles bind with each other to make big flocs. When suspended ions are flocculated into bigger particles, they could be isolated by floatation and filtration. Nowadays, numerous types of flocculants are commercially available for wastewater treatment, for example, polyacrylamide, polyferric sulfate, and polyaluminum chloride. Mercapto-acetyl chitosan macromolecule produced by reaction of mercapto-acetic acid with chitosan is very efficient for removal of turbidity and heavy metals from water (Fenglian and Wang 2011). Generally, flocculation and coagulation could not wholly remove heavy metals from water, hence must be accompanied by other treatment techniques. From industrial wastewater, tungsten metal removal takes place by using ferric

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chloride in acidic conditions more efficiently (Fenglian and Wang 2011). The process is not very efficient due to chemical consumption, excessive sludge production, slow metal precipitation, aggregation of metal precipitates, unsettling, and the long-term impact of sludge on the environment (Wang et al. 2008; Wang et al. 2005).

Flotation Currently, various types of flotation processes like dissolved air flotation and precipitation flotation are commonly used to remove metals from water (Rubio et al. 2002). In the process of flotation, chemical coagulant such as ferric salts, aluminum, or polymer coagulant is often used to increase the flocculent structure of floated particles to entrap the air bubbles easily. Dissolved air flotation (DAF) has been broadly utilized to remove metal ions (Rubio and Tessele 1997). In this flotation process small air bubbles to bind with suspended ions in water, creating clump having a density lower than water, which results in the flocs to float on the surface of the water and removed as sludge (Figs. 14.2 and 14.3) (Kyzas and Matis 2019). Ion flotation is another type of flotation process in which metal ions are removed via air bubbles and is based on attachment of hydrophobic-charged metal species in wastewater on the gas bubbles by using the surfactants (Hoseinian et al. 2015). The appropriate amount of gas bubbles in the flotation process is achieved by adding frothier such as polypropylene glycol, methyl isobutyl carbonyl, methyl

Fig. 14.2 Dissolved air flotation (DAF) unit at laboratory scale. (Image is modified with permission from Ref. Rubio and Tessele 1997)

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Fig. 14.3 Heavy metal aggregates between the Cu++–chabazite and precipitates of Fe(OH)3. The removal of heavy metal ions almost >98% by precipitates of Fe(OH)3. (Image is modified with permission from Ref. Rubio and Tessele 1997)

ethers, and ethanol. The species which are skimmed in flotation process are the metal particles. Removal of metal particles correlates with a vast metal particle/water proportion in the froth stage. It pursues through the selectivity of particles and evaluated on the basis of overall recuperations of both the metal particles and water from the flotation cell. The ion flotation process has been applied to remove lead (89.95%), cadmium (71.17%), and copper (81.13%) successfully from aqueous solution of plant-inferred biosurfactant tea saponin when the ratio of collector to metal was 3:1. Similarly, ion flotation process was used to expel Cr3+, Ag+, Cu2+, and Zn2+ from water up to 74% and 90% in acidic and alkaline medium, respectively. Generally, for the process, hexadecyltrimethylammonium bromide, sodium dodecyl sulfate as amassers, methyl isobutyl carbonyl, and C2H5OH as frothier are used. The method is efficiently used for removal of Cr+3 from wastewater up to 96.2% around pH 8.0 (Fenglian and Wang 2011; Kyzas and Matis 2019).

Ion-Exchange Process Ion-exchange technique is utilized for the removal of heavy metals from wastewater by the interchange of metal ions among substrate and encompassing medium. In the procedure, metal particles from dilute solutions are interchanged with particles held by electrostatic force on the substrate. A particle exchanger is an efficient technique for exchanging anions and cations from the surrounding materials (Acheampong et al. 2010). Generally, synthetic ion-exchange resins are used for this purpose. The most commonly used resins to remove metals from aqueous solutions are natural zeolites, synthetic resins, and silicate mineral. The resins are interconnected polymer

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framework that contains various functional groups incorporated by covalent bonding (Acheampong et al. 2010). These polymers are insoluble in water and mostly organic solvents. These materials are stable and easy to work over a broad spectrum of pH at elevated temperatures. Polystyrene cross-linked with 3–8% divinylbenzene is commonly used for water purification. Highly cross-linked polymer matrix is specially made for industrial applications, as the type of polymers is very expensive but easy to work. Typical ion exchangers are manufactured with pore size range of 20–50 nanometer. Regeneration of resins is easy due to small affinity of the trapped metal ions with resin. The ion exchangers with 50 and 100 meshes are used on commercial scale and analytical purposes, respectively. The type of ion exchangers is categorized as acidic and basic based on various functional groups. Acidic exchangers are further categorized as weakly acidic and strongly acidic having – COOH and –SO3H functional groups, respectively. The hydrogen ions of carboxylic or sulfonic acid groups are replaceable with heavy metal ions. As the solution containing heavy metal goes through the cations’ bed of column, metal ions are exchanged with hydrogen on the resin with the following process. nR‐SO3 H þ Mnþ ! ðR‐SO3  Þn Mnþ þ nHþ nR‐COOH þ Mnþ ! ðR‐COO Þn Mnþ þ nHþ Many types of ion exchangers possessing different functional groups like imino diacetate, phosphonic, phenol, and phosphine are used for water purification. Many scientists and researchers have reported that under variable experimental conditions, zeolites show high cation-exchange capacities for heavy metal ions. Clinoptilolite is considered a typical zeolite for water purification due to its high selectivity for metal ions (Saleh 2017). Similarly, basic cation exchangers, for example, weakly basic (secondary and tertiary amine) and strong (quaternary ammonium) exchangers, are effectively employed to remove heavy metal ions. Similarly, amphoteric exchangers have been investigated for water purification under different pH conditions (Saleh 2017). Recently, these exchangers are also called zwitterionic ion exchangers or bipolar electrolyte exchange resins (Jaishankar et al. 2014). Heavy metal uptake by exchangers is moderately influenced by specific factors, for example, temperature, pH, contact time, and initial metal concentration (Raouf and Raheim 2016). Ion-exchange techniques are broadly used for metal removal from water because of their numerous points of interest, for example, greater removal efficacy and kinetics and regeneration capability. The strategy is highly expensive and cannot be used at large scale for the removal of specific ions from wastewater. The other drawbacks of the technique are regeneration capability, secondary product production, nonselectivity, and high sensitivity to pH of the solution (Barakat 2011).

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14.5.2 Adsorption Adsorption is an economical and powerful technique for the expulsion of metals from water. The technique has gained attention due to its recovery capability of adsorbents and design and operational ease for the treatment of high-caliber pollutants in water (Chen et al. 2019).

Activated Carbon Adsorbents Activated carbon adsorbents are generally used for the removal of metal contaminants. Its convenient use comes from its large mesopore and micropore volumes and high surface area zone. To make the process cost-effective, certain additives like tannic acid (Üçer et al. 2006), alginate (Park et al. 2007a), surfactants (Ahn et al. 2009), and magnesium (Yanagisawa et al. 2010) are used, or activated carbon composite could be an alternative (Dias et al. 2007). Kongsuwan et al. (2009) investigated the utilization of activated carbon from bark of eucalyptus in the double sorption of Pb2+ and Cu2+ with high sorption of 0.53 and 0.45 mmol/g, respectively. A noteworthy scheme for the take-up of heavy metals was exhibited by adsorption (Vilela et al. 2016). Guo et al. (2010) investigated poultry waste to fabricate activated carbon for treating heavy metal (Guo et al. 2010). It was revealed that poultry waste activated high carbon adsorption capacity for heavy metals than commercial activated carbons derived from coconut shell and bituminous coal (Guo et al. 2010). The process is advantageous due to its easy availability and inexpensive sources of activated carbon (Peng et al. 2017).

Carbon Nanotube Adsorbents Iijima in 1991 discovered carbon nanotubes. These are good adsorbents for removal of heavy metals, for example, chromium (Pillay et al. 2009), copper (Liu et al. 2013), cadmium (Kuo and Lin 2009), lead (Kabbashi et al. 2009), and nickel (Kandah and Meunier 2007) from wastewater. Carbon nanotubes are multiwalled and singlewalled (Ihsanullah et al. 2016; Okpalugo et al. 2005; Rao et al. 2007). The metal particles are adsorbed onto carbon nanotubes due to sorption, electrostatic affinity, chemical interaction, and surface functional groups of carbon nanotubes (Rao et al. 2007). The adsorption limit of metal ions by crude carbon nanotubes can be remarkably increased by NaClO4, KMnO4, and HNO3 as oxidizing agents revealed adsorption of Pb2+ (75.3%) by using acidified multiwalled carbon nanotubes that contain oxy functional groups (Wang et al. 2007). Non-functionalized multiwalled carbon nanotubes have shown remarkable adsorption capacity (98%) of chromium (VI) from wastewater (Pillay et al. 2009; Desai et al. 2016). Both functionalized and non-functionalized multiwalled carbon nanotubes demonstrated better adsorption ability than that of AC (Xu et al. 2018). To avoid health hazards of carbon nanotubes

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Fig. 14.4 Removal of heavy metal ions (M2+) by carbon nanotubes. (Image is modified with permission from Ref. Rao et al. (2007)

to humans, an ecological well-disposed adsorbent, in which carbon nanotubes are clamped with calcium alginate, was synthesized, which can remove 67.9 mg/g of Cu by adsorption (Fig. 14.4) (Mubarak et al. 2014; Gupta et al. 2015; Vilardi et al. 2018).

Biosorption Biosorption is a relatively favorable and promising heavy metal exclusion method from water. Biosorption processes are especially useful when metal concentration in the water is low. Major sources of biosorbents are algal biomass, nonliving biomass (crab shell, bark, shrimp, lignin, squid, krill) (Apiratikul and Pavasant 2008a, b), and microbial biomass (fungi, bacteria, yeast). Algae, bacteria, yeasts, and fungi are also good candidates as metal biosorbents. Different types of reasonable nonliving plant materials, for example, potato peels (Aman et al. 2008), sawdust (Kaczala et al. 2009), eggshell (Park et al. 2007b), seed shells (Amuda et al. 2009), coffee husks (Oliveira et al. 2008), citrus peels (Schiewer and Patil 2008), and sugar beet gelatin gels (Mata et al. 2009) have been widely explored as alternative sources of biosorbents for heavy metal removal. Algal biosorbents are frequently used due to their economic and easy availability and high metal adsorption capacity (Apiratikul and Pavasant 2008a, b). Generally, different types of biosorbents are used for removal of different heavy metals like Zn and Cu by Chaetomorpha linum (Ajjabi and Chouba 2009; Ajjabi and Chouba 2009) (a green macro alga); Cr by Ulva lactuca (El-Sikaily et al. 2007) (green alga); Pb, Cd, Cu, and Zn by Caulerpa lentillifera (Apiratikul and Pavasant 2008a, b; Pavasant et al. 2006) (marine green macro alga); and Pb by Cladophora fascicularis (Deng et al. 2007) (green algae). Microbial removal of heavy metal ions has been found viable. Biosorption of metal

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Fig. 14.5 The biological system for the treatment of wastewater. (DP dephosphatation, DN I dinitrification 1, DN II dinitrification 2, N nitrification, DA deaeration). (Image is modified with permission from Ref. Chipasa 2003)

particles from water by bacteria includes B. cereus (Pan et al. 2007) and Pseudomonas aeruginosa (Joo et al. 2010). Yeasts and fungi are easy to grow and produce significant amount of biomass. Fungi biosorbents include A. niger (Tsekova et al. 2010) and Lentinus edodes (Bayramoglu and Arıca 2008). Heavy metals like Cu ion can be removed efficiently by caustic soda-treated fungus R. oryzae in a batch reactor. However, regeneration of biosorbents would be difficult (Bhainsa and D’souza 2008). Unexpectedly, sorption utilizing biochars and biosorbents is noteworthy when recovery of heavy metals and biosorbents is demanding. In that regard, biowaste (maize cob, wheat bran) is a best source for adsorption of heavy metals from wastewaters (Thaçi and Gashi 2018). Peels of vegetables and fruits and industrial and agricultural biowaste are cheap biomaterials for treatment of wastewater (Raouf and Raheim 2016). For the removal of lead ions from wastewater, wheat bran is being used. For adsorption of Ni, Zn, and Mn from aqueous solutions, maize cob, olive cake, and corncob were used. Biosorption is beneficial compared to other techniques due to low cost, high efficiency, and simplicity of the process. Efficacy of biosorption depends on various types of adsorbents, functional groups, and concentrations of heavy metals in wastewater (Fig. 14.5) (Chojnacka 2010; Chipasa 2003; Wang et al. 2019; Yuna 2016).

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14.5.3 Membrane Filtration Membrane filtration with various types of membranes displayed great ability for heavy metal uptake due to smooth operation, high efficacy, and space-sparing characteristics. However, process complexity, high cost, membrane fouling, and low penetration have restricted their use (Fenglian and Wang 2011). Membrane filtration processes (ultrafiltration, electrodialysis, and nanofiltration) are designed on the basis of reverse osmosis mechanism.

Ultrafiltration Ultrafiltration is similar to reverse osmosis water purification process but with membranes containing larger number of pores. Ultrafiltration membranes are pressure-driven which work efficiently at low transmembrane pressure for the removal of soluble particles from wastewater. Ultrafiltration layers of pore size are bigger than dissolved metal ions, small molecular weight compounds, or hydrated ions (Gupta et al. 2012). These particles are removed effectively by ultrafiltration membranes. Micellar-enhanced ultrafiltration and polymer-enhanced ultrafiltration were suggested for maximum removal of wastewater metal ions (El Zeftawy and Mulligan 2011). In the presence of surfactants, micellar-enhanced ultrafiltration has been declared a successful parting strategy to remove metals from water ultrafiltration membrane with pore sizes more diminutive than micelle which can remove metal ions containing micelles. However, the un-trapped particles rapidly go through the ultrafiltration layer (Fig. 14.6) (Huang et al. 2017). Generally, an anionic surfactant,

Fig. 14.6 Flowchart showing micellar-enhanced ultrafiltration at laboratory scale. (Image is modified with permission from Ref. Huang et al. 2017)

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sodium dodecyl sulfonate, is chosen for successful elimination of metal particles in micellar-enhanced ultrafiltration. Heavy metal ion removal proficiency by micellarenhanced ultrafiltration highly depends on the concentrations and characteristics of metals, ionic strength, solution pH, surfactants, and various variables associated with membrane process. Zinc was successfully removed from wastewater by micellarenhanced ultrafiltration in the presence of sodium dodecyl sulfate with rejection coefficient up to 99% (Landaburu-Aguirre et al. 2009). Similarly, micellar-enhanced ultrafiltration effectively removed Pb2+, Cu2+, Zn2+, Cd2+, and Ni2+ from wastewater with two anionic surfactants (sodium dodecyl sulfate and straight alkyl benzene sulfonate) (Shen and Hankins; Landaburu-Aguirre et al. 2010). It is vital to regenerate and reuse the surfactant to make it a cost-effective and eco-friendly process (Samper et al. 2009). Li et al. (2009) tried chelation-pursued ultrafiltration for the separation of Zn2+ or Cd2+ from sodium dodecyl micelles. In the process, ethylenediaminetetraacetic acid was used as chelating agent at pH 4.4. It was considered as perfect combination for heavy metal removal (Zn2+, Cd2+) and retake of sodium dodecyl sulfate (68.5% for Zn2+ and 65.5% for Cd2+). The heavy metal removal efficiency of recovered sodium dodecyl sulfate in micellar-enhanced ultrafiltration was 89.6 and 90.3% for Zn2+ and Cd2+, respectively. In the technique, sulfuric acid at pH 1 was used most effectively for isolating of metal particles (96.1% for Zn2+ and 98.0% for Cd2+) and recovering of sodium dodecyl sulfate (54.3% for Zn2+ and 58.1% for Cd2+). The efficiency of recovered sodium dodecyl sulfate was 87.8 and 88.1% for Zn2+ and Cd2+, respectively (Li et al. 2009). Polymer-enhanced ultrafiltration has been projected as a feasible strategy to isolate several metal ions from wastewater (Huang et al. 2016). Polymer-enhanced ultrafiltration uses water-soluble polymers to form high molecular weight compounds of metal ion, having molecular mass higher than membrane. Macromolecules are pumped through ultrafiltration membrane, and retentate is used to isolate metal ion particle (Trivunac and Stevanovic 2006a, b). Different complexing agents like polyethylenimine (Shen and Hankins 2017), diethylaminoethyl cellulose (Trivunac and Stevanovic 2006a, b), polyacrylic acid (Zhu et al. 2016), and humic acid have been tested to accomplish removal and isolation process at low cost. The fundamental factors influencing polymer-enhanced ultrafiltration are polymers and metal types, the proportion of polymer to metal, pH, and presence of other metal particles in the solution (Llanos et al. 2008). The ultrafiltration water purifier works without electricity with high selectivity. Moreover, the process is economical and produces no pollution due to un-involvement of chemicals (Petrinic et al. 2015). However, ultrafiltration cannot remove the dissolved salts from the wastewater, thus poisonous species even herbicides will pass through the membrane challenging the water quality. The process also produces a greater amount of sludge and foul odor of the water like other techniques (Petrinic et al. 2015; Llanos et al. 2008).

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Reverse Osmosis Reverse osmosis involves partially permeable membranes like thin-film composite and cellulose acetate membrane to remove heavy metal ions, large particles, and unwanted molecules from wastewater at 150–600 pounds per square inches. Reverse osmosis is a well-known technique for the removal of metal ions from water by applying the abovementioned pressure from an area of high solute concentration through a membrane to a region of low solute concentration. In the process, heavy metals are removed by a semi-porous membrane, enabling the wastewater to go through it while discharging the contaminants at high pressure than osmotic pressure (Tripathi and Ranjan 2015). Reverse osmosis membranes have shown many advantages by decreasing the production of organic pollutants, total dissolved solids, viruses, heavy metals, and microscopic organisms. Reverse osmosis is an everincreasing common wastewater treatment technique in chemical and ecological engineering covering over 20% of the world’s desalination. It was explored that Ni2+ and Cu2+ ions were effectively removed by reverse osmosis with dismissal efficacy up to 99.5% by using disodium ethylenediaminetetraacetic acid (Thaçi and Gashi 2018). Reverse osmosis is examined that enhancing feed flow rate and working pressure improves efficiency by decreasing the concentration of ions. The process is preferred due to reduced discharge and purchases and the conservation of water resources. However, the process is expensive with high power consumption for pumping pressure and regeneration of the membranes.

Electrodialysis Electrodialysis emerged as an efficient technique for the uptake of heavy metals from water. Electrodialysis is an alternative membrane procedure to remove charged particles (cation-exchange and anion-exchange) from solution by utilizing the electric potential as driving force. They allowed to pass the wastewater from the cell compartments and observed that the ions were attracted according to their affinity toward electrodes (Li et al. 2008). For electrodialysis of Co(II) and Ni(II), cation-exchange membranes, like sulfonated polyvinyldifluoride membrane and perfluorosulfonic nafion, were used (Tzanetakis et al. 2003). For removal of Ni(II) and Co(II) (69 and 90%), perfluorosulfonic nafion membrane was used. However, its working efficacy depends on temperature, flow rate, and voltage at different concentrations (Tzanetakis et al. 2003). High voltage and temperature enhanced electrodialysis cell execution, though, with rising flow rate, the separation rate is reduced. The use of membranes with greater exchange of particles led to improved cell efficiency (Mohammadi et al. 2004). Electrodialytic removal of Cd(II) from water was also performed using electric field. It was examined that execution of an ion species does not depend on electrodialysis cell species but relies on the working conditions and structure of electrodialysis cells (Jakobsen et al. 2004). Similarly, expulsion of Cr

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(+VI) using an assembled electrodialysis production plant containing an IE membrane was performed with affordable results and highest pollutant level of 0.1 mg/ dm3 for Cr (Nataraj et al. 2007). Moreover, electrodialysis process was successfully used to get rid of Cu, Cr, Pb, and Fe metals from wastewater (Abou-Shady et al. 2012; Cifuentes et al. 2009). Since electrodialysis is a membrane process, it requires careful operation, clean feed, and periodical perpetuation to avert any stack damages. The procedure has been extensively used for the removal of Cu and Cr from industrial effluents and salt production (Sadrzadeh and Mohammadi 2009). The major drawback of electrodialysis process is the formation of hydroxides of metals which obstruct the membrane and corrosion of the electrodes (Cifuentes et al. 2009).

14.5.4 Electrochemical Treatment Electrochemical treatment strategy is based on metal ion deposition on a cathode surface. These metal particles can be easily collected in elemental state using simple mechanical process. The electrochemical processing requires high power consumption of electricity, which makes it a useless technique for large-scale wastewater treatment. However, strict environmental rules for industries for wastewater release have made the technique significant over the past two centuries (Wang et al. 2007). Electrochemical treatments include different methods such as electrocoagulation, electrodeposition, and electroflotation. Electrocoagulation includes the production of coagulants in situ by dissolving electrically either iron or aluminum ions with respect to electrodes. Hydrogen gas is discharged from the cathode and the metal ion generation happens at the anode. The H2 gas can skim the flocculated ions out of the water (Tchamango et al. 2010). Electrocoagulation has been effectively used to remove As+5, Ni2+, and Mn2+ (Hasieb 2012). Aluminum electrodes are used for removing heavy metal ions like Zn, Cu, Ag, and chromate ions (Heidmann and Calmano 2008) by hydrolyzing and co-precipitating as hydroxides. Electrocoagulation is an equally powerful strategy for removing suspended solids, tannins, and dyes (Wang et al. 2007). Electroflotation use a basic procedure that hovers contaminants to the water surface by small air bubbles of H2 and O2 gases produced from water electrolysis. It means the electrochemical reactions at the anode and cathode are O2 and H2 evolution. The effectiveness of the removal of contaminants relies on the size of air bubble pockets. Successful electroflotation needs the production of uniform and small air bubbles. The uniform and small size air bubbles produced electrically give much preferred execution. The development of steady, active, dynamic, and inexpensive new materials for oxygen evolution will slowly replace the traditional flotation techniques (Chen 2004). Major advantages of the wastewater treatment are small amount of chemical consumption and less sludge production. The main disadvantage of the process is costly electric supply and arrangement of high-cost instruments (Wang et al. 2007).

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14.5.5 Phytoremediation Phytoremediation is also known as botano-, vegetative, or green remediation. In the process, specific plants are used to clean up soil and water polluted by metals. Phytoremediation involves the use of specific plants to rectify the selected substances in polluted sludge, surface water, sediment, and wastewater groundwater and soil (Moosavi and Seghatoleslami 2013). Depending on the type of pollutant and basic procedure used, phytoremediation is generally classified as phytoaccumulation/phytoextraction, phytodegradation/phytotransformation, phytovolatilization, rhizo-filtration/phytostimulation/rhizostimulation, and phytostabilization. Phytoremediation technique generally uses three types of procedure, rhizo-filtration, phyto-stabilization, and phytoextraction to remove heavy metals from contaminated water (Moosavi and Seghatoleslami 2013).

Phytoextraction In phytoextraction process, plant roots are involved to translocate and store contaminants in different parts of plant. The strategy is useful for the purification of waste containing metals (Moosavi and Seghatoleslami 2013). All plants have variable metal uptake ability depending on the type and concentration of impurities. Compounds of metals have been effectively phytoextracted, and lead, zinc, copper, chromium, and nickel compounds have been effectively extracted by using the technique. The process proceeds naturally using hyperaccumulators or can be initiated artificially by using chelates to enhance bioavailability. Sunflower and mustard plants have been successfully used for phytoextraction of heavy metals (Raskin et al. 1997).

Rhizo-filtration Rhizo-filtration is like phytofiltration process, but plants used for remediation are grown in greenhouse nurseries with roots in water. Rhizo-filtration process uses suitable plants with fast root development system for removal of heavy metals from wastewater. The process is based on breakdown of natural contaminants in water through microbial movement in the rhizosphere. The process is suitable for remediation of polluted groundwater not for polluted soils. In some cases, some metals were removed by binding to roots without the addition of chelating agents, for example, ethylenediaminetetraacetic acid derivation. Alginate–disodium ethylenediaminetetraacetic acid is prepared as adsorbent with high tendency for heavy metal particles, for example, Pb2+, Cr3+, Cu2+, Co2+, and Cd2+. Alginate– disodium ethylenediaminetetraacetic acid can adsorb more than 85% of metal ions from contaminated water. Adsorption increases with increase in pH (Wang et al. 2019). In the process, metal ions are absorbed on the roots of plants. A wide variety

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of plant species are known as a best source for removal of dangerous metals, for example, Pb2+, Cd2+, Cu2+, Cr(VI), Zn2+, and Ni2+, from aqueous solutions (Varkey et al. 2012).

Phyto-stabilization Phyto-stabilization is introduced for the remediation of sludge and sediments. During the process, contaminants are consumed and adsorbed in plant roots. The process of phyto-stabilization depends on the resistance capacity of a plant to certain impurity. Phyto-stabilization is a useful process for removing with variable efficacy heavy metals, as plants are cultivated and developed under the contaminated conditions. The significant advantages of phytoremediation over ordinary treatment strategies are minimal effort, high productivity, less problematic to the environment, and no need for new plants. There is only a minimum production of chemical and biological sludge. However, removal of metals depends on the development of plant growth, and recovery of the plant for further phytoremediation requires a long time. Phyto-stabilization also depends on the conditions like atmospheric climate, temperature, topography, altitude, resistance of the plant to the contamination, and probability of natural harm due to leaching of dissolvable contaminants (Tangahu et al. 2011). Table 14.2 describes various water treatment techniques with their drawbacks and benefits.

14.6

Conclusions

Generally, industrial effluents and domestic wastewater contain heavy metals as significant pollutants. The study produced significant information about heavy metal pollution in wastewater, their sources, and remediation techniques. Human being and nature are two primary sources of heavy metal pollution in wastewater. Natural factors include weathering of rocks, soil erosion, volcano eruptions, aerosol particulate, and urban runoffs, whereas human sources include mining, industrialization (textile, metal finishing, and electroplating), and coal and nuclear power production. Water bodies receiving untreated water effluents from industries and domestic resources cause contamination and pose severe threats to human being, plants, and animals. A particular focus is given to physicochemical removal processes such as chemical precipitation, adsorption, ion exchange, membrane filtration, reverse osmosis, electrodialysis, adsorption, and phytoremediation. Their advantages and disadvantages have been evaluated in details. It is evident from our current studies that new adsorbents and membrane filtration techniques are widely applied for the treatment of wastewater. General criteria for the selection of suitable technique for treatment wastewater effluents are based on cost, applicability, and plant simplicity. Reverse osmosis is the most common process used for removal of metal ions with some

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Table 14.2 Summary of water purification techniques with their advantages and disadvantages Technique Chemical precipitation

Advantage Ease of operation, inexpensive capital investment Adapted to treat heavy metal ion concentration

Coagulation and flocculation

Sludge settling capacity and dewatering characteristics are very good

Ion exchange

Widely applied to expel heavy metals from wastewater and selective for metal ions High removal efficiency High treatment capacity Ion-exchange resin can be regenerated Removal efficiency is very high High metal selectivity Low detention period Low-cost adsorbent

Flotation

Adsorption

Simplicity High efficiency

Electrochemical methods

Membrane filtration

Less production of sludge Rapid techniques and need few chemicals Low pressure Low space requirement High separation efficiency to heavy metals or selectivity

Disadvantage Not economical

Ineffective with low metal ion concentration Produce large amount of sludge, sludge disposal problems Increased sludge volume generation High consumption of chemicals Costly and complex, available for less number of metal ions

References Lawrence K. Wang (2005) and Wang et al. (2008)

Wang et al. (2008) and Fenglian and (2011) Barakat (2011)

Nonselective and sensitive to the pH of the solution High maintenance and operation cost High initial investment cost

Fenglian and Wang (2011)

Capacity of adsorption relies on the kinds of adsorbent Performance depends upon adsorbent Removal of heavy metals from low wastewater concentrations Electric supply is expensive

Farooq et al. (2010), Gurel (2017), and Wang et al. (2019)

Very costly techniques to set up the instrument High operation cost and complex Membrane fouling has limited heavy metal removal

Wang et al. (2007)

Fenglian and Wang (2011)

(continued)

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Table 14.2 (continued) Technique Membrane process and ultrafiltration

Advantage Less chemical consumption Less solid waste produced High efficiency

Electrodialysis

Reverse osmosis

Phytoremediation

High selectivity

Broadly utilized for the treatment of industrial effluents Reduced discharge Conservation of the water resources Low cost, high efficiency Minimization of chemical and biological sludge No additional nutrient requirement No need for disposal sites, hence reducing risk of spread of contaminants

Disadvantage Removal (%) decreases with the presence of other metals Odor of the water still remains the same High initial and running cost Can’t expel the dissolved salts from the water Low flow rates High operation cost due to membrane fouling and energy consumption Formation of metal hydroxide, which clogs the membrane Expensive High power consumption due to pumping pressure For further phytoremediation, regenerate the plant and it requires long time to remove metals

References Llanos et al. (2008)

Cifuentes et al. (2009) and Sadrzadeh and Mohammadi (2009) Thaçi and Gashi (2018) Moosavi and Seghatoleslami (2013) and Raskin et al. (1997)

limitations such as high cost of materials and limited pH range. Our studies suggest that there is always a growing need of new materials and alternative methods for the removal of heavy metals.

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Chapter 15

Mechanisms and Approaches for the Removal of Heavy Metals from Acid Mine Drainage and Other Industrial Effluents Vhahangwele Masindi

, Muhammad S. Osman, and Memory Tekere

Contents 15.1 15.2

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Various Mechanisms Used for the Removal of Heavy Metals . . . . . . . . . . . . . . . . . . . . . . . . . 15.2.1 Precipitation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.2.2 Adsorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.3 Filtration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.3.1 Microfiltration, Ultrafiltration, and Nanofiltration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.3.2 Reverse Osmosis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.3.3 Electrodialysis and Electrodialysis Reversal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.3.4 Ion Exchange . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.3.5 Phytoremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.4 Valorization Options . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.5 Challenges and Failures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.6 Future Perspectives . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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V. Masindi (*) Magalies Water, Scientific Services, Research & Development Division, Brits, South Africa Department of Environmental Sciences, School of Agriculture and Environmental Sciences (CAES), University of South Africa (UNISA), Florida, South Africa Council of Scientific and Industrial Research (CSIR), Built Environment (BE), Hydraulic Infrastructure Engineering (HIE) Group, Pretoria, South Africa e-mail: [email protected] M. S. Osman Council of Scientific and Industrial Research (CSIR), Built Environment (BE), Hydraulic Infrastructure Engineering (HIE) Group, Pretoria, South Africa e-mail: [email protected] M. Tekere Department of Environmental Sciences, School of Agriculture and Environmental Sciences (CAES), University of South Africa (UNISA), Florida, South Africa e-mail: [email protected] © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature Switzerland AG 2021 Inamuddin et al. (eds.), Water Pollution and Remediation: Heavy Metals, Environmental Chemistry for a Sustainable World 53, https://doi.org/10.1007/978-3-030-52421-0_15

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Abstract Contamination of surface and underground water resources with heavy metals has been an issue of paramount concern in recent decades. This is attributed to the magnitude of impacts that heavy metals pose to terrestrial and aquatic living organisms on exposure. Furthermore, mining activities and metal processing interventions generate huge volumes of heavy metal-laden streams to different receiving environments. Similarly, the nature of chemical species present in the water bodies depends on the nature and type weathered and host minerals. On that regard, scientists and engineers have developed a number of prudent and pragmatic technologies for the removal of these pollutants from water bodies. Different techniques that rely on different mechanisms and approaches are often employed for the removal of heavy metals from wastewater, and they include neutralization, adsorption, filtration, and phytoremediation, among others. These technologies have been reported to present different advantages and disadvantages on application. This chapter will give an in-depth discuss on different mechanisms and approaches employed for the removal of pollutants from wastewater. Furthermore, sustainable approaches for valorization and beneficiation of waste effluent streams will also be discussed in this chapter, including preliminary challenges of beneficiation. This will impart value to wastewater via the conversion of waste into a resource, thus enabling them to perceive these wastewater streams as a resource not waste. Recovery of heavy metals from wastewater can aid in offsetting the running cost of the treatment process, hence making it self-sustainable and eco-friendly. This will also foster the concept of circular economy and sustainable development. This chapter will also highlight the advances made in terms of acid mine drainage management and heavy metals treatment. In particular, the advancements, failures, challenges, and future research avenues will also be unpacked in this chapter. This will be used as an avenue to guide future research and to identify potential opportunities for future research in wastewater treatment. Keywords Heavy metals · Acid mine drainage · Industrial effluents · Removal mechanisms · Adsorption · Precipitation · Phytoremediation · Crystallization · Filtration · Ion exchange

15.1

Introduction

Different metallurgical and mining processes generate wastewater streams, which are rich in heavy metals. The discharge of this effluent to different receiving environments may pose serious environmental problems if not properly managed (Amos et al. 2015; Baker and Banfield 2003; Dold 2017). In response to that, stringent environmental regulations have been developed, and they require wastewater to be properly managed prior contamination of surface and groundwater resources. According to different environmental regulations, their prescribed limits

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range from
0.05 to 0.5 depending on their ecotoxicological impacts on exposure (Sheoran and Sheoran 2006; Sheoran et al. 2011a; Sheoran et al. 2011b; Simate and Ndlovu 2014; Skousen et al. 2017). Acid mine drainage is predominated by Fe, Al, Mn, and SO4 among traces of other heavy metals such as Cu, Zn, As, Cr, Ni, etc. (Masindi 2016; Masindi et al. 2015, 2018a; Sahoo et al. 2013; Sheoran and Sheoran 2006; Sheoran et al. 2011a, b; Simate and Ndlovu 2014). The composition of acid mine drainage depends on the chemical properties of the host rock which is been weathered including the leachable fractions in the mother rock. Acid mine drainage mainly forms from the oxidation of sulfide bearing minerals such as pyrite, arsenopyrite, and marcasite, among others. As an example, the following equations can be used to denote the oxidation of pyrite (Masindi et al. 2016, 2018b, c, 2019): 2 þ 2FeS2ðsÞ þ O2ðgÞ þ 2H2 Ol ! 2Fe2þ ðaqÞ þ 4SO4ðaqÞ þ HðaqÞ

ð15:1Þ

Essentially, the oxidation of sulfide to sulfate promote the oxidation of (Fe(II)) to (Fe(III)) (Eq. 15.2): þ 3þ 4Fe2þ ðaqÞ þ O2ðgÞ þ 4HðaqÞ ! 4FeðaqÞ þ 2H2 OðlÞ

ð15:2Þ

These reactions are also mediated by a number of microorganisms, which have the potential to oxidize both sulfur and iron. Moreover, the produced ferric ions can further oxidize pyrite into ferrous ions (Eq. 15.3): 2þ þ FeSðsÞ þ 14Fe3þ ðaqÞ þ 8H2 OðlÞ ! 15FeðaqÞ þ 2SO4ðaqÞ þ 16HðaqÞ 2

ð15:3Þ

The aforementioned reactions will produce H+ which will then be essential in maintaining the solubility of Fe(III). Furthermore, acidic pH will create conditions, which are suitable for the leaching of metals from the surrounding materials, hence leading to high levels of dissolved ions. On the other hand, metallurgical processes lead to the generation of effluent stream that is rich in heavy metal residues from the separation processes. This also enriches the product effluent stream with high levels of dissolved ions. In most instances, these streams are rich in high levels of undesired heavy metals, contaminants, and residues, hence leading to high load of dissolved components (Amos et al. 2015; Baker and Banfield 2003; Bwapwa et al. 2017; Dold 2017; Fernando et al. 2018; Gazea et al. 1996; Johnson and Hallberg 2005). This induces a notable challenge to the receiving environment if poorly managed (Masindi et al. 2019). These effluent streams have been documented to contain some toxicological impacts to living organisms on exposure. According to epidemiological and toxicological studies, the presence of heavy metals in water can lead to numerous cancers, affect human productivity, reduce biodiversity, reduce plant productivity, and affect ion exchange between plants and impair metabolic activities in living organisms. This need to be managed prior the degradation of the environment and its resources.

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As such, scientists developed numerous technologies to remove these pollutants from water bodies before it could be released into the receiving environment (Anawar 2013; Candeias et al. 2018; Jain et al. 2016; Masindi et al. 2018a; Nordstrom et al. 2015; Parbhakar-Fox and Lottermoser 2015; Smirnova et al. 2011; Truter et al. 2014; Zumarán Farfán et al. 2003). These techniques rely on different mechanisms and approaches, and they include precipitation of metals at different pH ranges, adsorption of pollutants from aqueous medium using an adsorbent, filtration using different membrane technologies, ion-exchange process using different resins, electrodialysis, and phytoremediation (Gazea et al. 1996; Kalin et al. 2006; Kefeni et al. 2017; Lewis 2010; Liu et al. 2017; Masindi et al. 2017b, 2018a, b, c; Mayes et al. 2009; Nleya et al. 2016; SánchezAndrea et al. 2014; Sheoran and Sheoran 2006; Sheoran et al. 2010; Skousen et al. 2017; Valenzuela et al. 2009). These techniques and their mechanisms will be discussed in detail. Their success, limitations, and failures will also be highlighted including future research avenues.

15.2

Various Mechanisms Used for the Removal of Heavy Metals

Different mechanisms are employed to attenuate, remove, recover, and valorize heavy metals from aqueous environments as depicted below. The mechanisms and approaches employed for heavy metals attenuation are shown in Fig. 15.1.

15.2.1 Precipitation Chemical precipitation has been a technique, which is prevalently employed for the removal of heavy metals from the aqua-sphere. This process is proceeded by binary mechanisms, which include co-precipitation and adsorption (Benatti et al. 2009;

Fig. 15.1 Schematic presentation of the mechanisms used for the removal of heavy metals from aqueous solution

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Blais et al. 2008). Furthermore, these processes are dependent on the solubility of the metal and adsorption sites for binding using weak van der Waal forces. Often, heavy metals precipitate to beyond their solubility and subsequently get proceeded by co-precipitation and adsorption concomitantly (Ahmad and Idris 2013; Zhu et al. 2007). Precipitation is dependent on pH, temperature, solid to liquid ratios, concentration, affinity/ionic strength, and coexisting ions. Nucleation followed by particle growth is the process that governs the precipitation process. Specifically, nucleation is comprised of molecules between 10 and 100 and particle growth which is influenced by the addition of more molecules which lead to an increased particle size due to electrostatic forces. This process is also driven by the temperature, reaction kinetics, and chemical characteristics of the source water bodies (Barakat 2011; Blais et al. 2008; Fu and Wang 2011; Westholm et al. 2014). Weathering processes and anthropogenic activities have gradually and enormously deteriorated the quality of water. This is attributed to their ability to load high levels of heavy metals into the aqua-sphere. As highlighted initially, the type and nature of these metals are dependent on the host rock and the upstream activities. As an initial step toward the management of heavy metals, sulfides, hydroxides, oxides, and carbonates are added to promote the formation of insoluble species (Bwapwa et al. 2017; Gazea et al. 1996; Kefeni et al. 2017; Lewis 2010; Mayes et al. 2009). However, pH is the driving factor for this process. The desludging process or filtration is the techniques used to remove the precipitate out of the reactor. Sequence required to remove metals as precipitates includes neutralization by an alkaline agent, precipitation after nucleation, coagulation and flocculation process, and desludging (Fig. 15.1). Other chemical components in aqueous medium affect the precipitation process. Chemical precipitation using alkalis to raise the pH of heavy metal-rich water, hence, promotes the formation of hydroxides, sulfides, and carbonates depending on the feed. Commonly used alkalis are rich in Mg, Ca, and Na. On that regard, dissolved metal will precipitate as the pH gradually increase, hence exceeding their solubility limit (Table 15.1). Furthermore, metal hydroxide is susceptible to redissolution, but this is also subjected to the pH. Low and high pH values mainly favor redissolution. Therefore, the optimum pH for precipitation can inversely affect other precipitated metals since one metal may cause another metal to solubilize, hence getting reintroduced into the aqua-sphere. To be particular, wastewater solution/effluent mainly contain multi-metals composition hence making it unsuitable

Table 15.1 Typical chemical reagents used in the precipitation of metals from the aqueous solution (Masindi et al. 2017a; Johnson and Hallberg 2005; Naidu et al. 2019; Nleya et al. 2016; PozoAntonio et al. 2014; Sahoo et al. 2013; Sheoran and Sheoran 2006; Simate and Ndlovu 2014) Magnesium Magnesite Brucite Periclase Magnesium bicarbonate

Calcium Limestone Hydrated lime Lime Dolomite

Sodium Soda ash Caustic soda Sodium bicarbonate Sodium sulfide

Others Mg-rich waste materials Ca-rich waste materials Na-rich waste materials Ca, mg, and Na waste

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Fig. 15.2 Schematic presentation of the process used for the precipitation of heavy metals using alkaline agents

for selective and stepwise recovery (Fig. 15.2). In addition, the generated sludge will be complex, heterogeneous, and difficult to handle. Metal oxides/hydroxides, sulfides, carbonates, and phosphates are forms in which metals can be removed from aqueous medium (Masindi 2016; Masindi et al. 2017a, b, c). In a response toward the circular economy phenomena, recovery of heavy metals via selective precipitation has been explored extensively (Blais et al. 2007, 2008). Redox reactions are the key factors that govern the removal of metals. Noteworthy, cost factor also play a key role in the selection of the alkaline agent to be used for the removal of contaminants from wastewater, as such, biological option has received massive attention but the challenge is the maintenance of biota and their sensitive nature that discourages end users. To further save the costs and acquire a synergy, hybrid systems that combine two or three mechanisms or materials are employed. The schematic presentation of the precipitation of heavy metals using an alkaline agent is shown in Fig. 15.2. However, due to numerous factors such as quick time, effectiveness, and simplicity, precipitation using chemical agents gained attention in the removal of heavy metals.

Reagents Used for Metals Precipitation For valorization purposes, researchers and engineers rely on sequential and fractional precipitation of metals from aqueous environment (Masindi et al. 2018c). This is the best approach to eliminate the generation of waste that triggers additional environmental challenges. To attain that, different alkaline agents are used to raise the pH of the water bodies on a stepwise fashion, thus systematically attenuating heavy metals from the aqua-sphere. Specifically, for acid mine drainage treatment,

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chemical treatment has been a prime mechanism for heavy metal attenuation from acid mine drainage and other industrial effluents. Reagents of Mg, Ca, and Na nature are used to systematically remove heavy metals (Zvimba et al. 2012; Simate and Ndlovu 2014; Sheoran and Sheoran 2006; Kefeni et al. 2017) (Table 15.1). The ultimate goal is to raise the pH of the water until metals selectively precipitates. The dosage and temperature are the driving factors for this process since enough alkalinity is needed to raise the pH of the water, hence ensuring the formation of insoluble hydroxide. The typical alkaline materials used in precipitation processes are depicted in Table 15.1. Thermo-mechano-activation of magnesite can produce brucite and periclase; similarly, limestone can be calcined to produce hydrated lime and lime (Magagane et al. 2019). Residues from the extraction of these minerals are also rich in these mineral; thence, they can partially perform the functions of these minerals when exposed to acidic water.

Heavy Metals Precipitation as a Function of pH Variation in pH push the metals to an unsoluble state, hence leading to metals precipitating as hydroxides, carbonates, and sulfides depending on the seeding material. The removal of co-precipitive metals during precipitation of the soluble metals is aided by the presence of hydroxide, which acts as an adsorbent during the precipitation reaction. For example, hydroxide precipitation of metals is charged, and they can be used for co-precipitation and adsorption reactions for metals which are positively charged. The following reactions, i.e., precipitation, co-precipitation, and adsorption, generate particles and solids that require filtration for them to be removed. The precipitation process is ideal when metal recovery needs to be pursued. pH is a valuable variable which must be optimized for optimal metal removal and recovery. The precipitation of metals varies from different pH gradients (pH  3 to 12) (Table 15.2). Precipitation of metal hydroxides varies depending on pH and feed seeding material (Table 15.2) (Yang et al. 2006; Skousen et al. 2006; Demchak et al. 2001; Johnson and Hallberg 2005). The physicochemical properties are principal components that influence the selection of alkaline agents for the neutralization process. Other factors include the receiving aquatic ecosystem, availability of electrical power, proximity of the feedstock, cost of the material, treatment time, and quantity of resultant solids. The Table 15.2 The pH levels at which metals are precipitated from the aqueous media (Masindi et al. 2018c)

Metal ion Al3+ Fe3+ Mn2+ Cr3+ Cd2+ Fe2+

pH 4.5 3.5 8.5 5.5 7.0 5.5

Metal ion Hg2+ Na+ Pb2+ Zn2+ Cu2+ Mg2+

pH 7.5 7.0 6.0 7.0 5.5 10.0

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employment of chemicals for the removal of metals via neutralization of acidity has been proven effective at industrial scale (Maree et al. 1997, 2013; Maree and Du Plessis 1993; Gitari et al. 2006).

15.2.2 Adsorption In colloidal chemistry, components of different charges attract each other using physisorption or chemisorption. This mechanism has been widely employed for the removal of pollutants from different media. It is a surface and charge-driven phenomenon whereby a compound that is negatively charged will attract positively charged surface (Langmuir 1997). This process relies on adsorbent and adsorbate (concentration). Moreover, the process also depends on the charge balance of the solution (zeta potential or the point of zero charge). A number of charged materials such as clay mineral, biochar, activated char, and metals (Al, Fe, Mn, Mg, Ca, etc.); modified compounds; and their composites are widely employed for the removal of contaminants from aqueous media. Valence of the adsorbent is the main component that determines the capacity of the adsorbent to scavenge pollutants from the aquasphere (Sparks 1995; Sparks and Sparks 2003). Biochar has been extensively explored for the removal of contaminants from the aqua-sphere. Their net negative charge on the surface enables them to scavenge cationic pollutants from feedwater (Anawar et al. 2015). As an example, the stepwise process for the synthesis of biochar from organic materials is shown in Fig. 15.3. Combusting organic matter in the absence of oxygen (anaerobic conditions) leads to the formation of a highly reactive biochar (Fig. 15.3). This material has a very

Fig. 15.3 A schematic presentation of the process used for the synthesis of biochar from organic materials

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Fig. 15.4 Different mechanisms that governs the removal of pollutants from aqueous solution using biochar

high surface area, which is also proportional to its high adsorption capacity. Due to that virtue, biochar has the ability to remove organic and inorganic contaminants from the aqua-sphere by employing differing mechanisms (Fig. 15.4). It can also remove odor (Geosmin and 2MIB) causing compounds and organic contaminants. Mechanisms of pollutants removal from aqueous media using biochar are shown in Fig. 15.4. Ion exchange via isomorphous substitution and valence difference can be the mechanism that governs the ability of the adsorbent to adsorbed pollutants onto its surface due to charge difference, functionalization, and affinity. Complexation can also play a role in the removal of contaminants from aqueous solution. If the material is exchanging base metals, the precipitation of metals can also take place. Scientists have developed a number of models to point out modes that governs the removal of pollutants from aqueous solution (Godlewska et al. 2017; Oliveira et al. 2017; Sophia and Lima 2018). Adsorption kinetics, isotherms, and thermodynamics are models used to predict the mechanisms governing the removal of heavy metals from wastewater (Tran et al. 2017).

Percentage Removal and Adsorption Capacity The amount of contaminants adsorbed by the adsorbent can be determined using the following Eqs. (15.4 and 15.5):

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 Removal percentage ð%Þ ¼

Ci  Ce Ci

Adsorption capacity ðqe Þ ¼

  100

ðC i  Ce ÞV m

ð15:4Þ ð15:5Þ

These equations determine the percentage removal of the contaminants and adsorption capacity of the adsorbent for adsorbate, where: Ci ¼ Initial concentration Ce ¼ Residual concentration V ¼ Solution volume m ¼ Adsorbent mass

Adsorption Kinetics To determine mechanisms governing the removal of contaminants from aqueous solution, adsorption kinetics are used. Mainly, the pseudo-first- and second-order equations are commonly employed (Tran et al. 2017; Ho et al. 2005; Ho and McKay 1999; Ho et al. 2000). Pseudo-First-Order Kinetic The rate at which contaminants are removed from the aqua-sphere, pseudo-firstorder kinetic model is usually employed. The following equation is mainly employed to determine the rate governing the reaction (Lagergren 1898): log ðqe  qt Þ ¼ 

k1 t þ log qe 2:303

ð15:6Þ

where: k1 refers to the rate coefficient (min1). qe refers to the amount of heavy metals adsorbed at equilibrium. qt refers to the amount of the heavy metals/contaminants adsorbed at time t. Pseudo-Second-Order Kinetic Pseudo-second-order kinetic is also used to point out mechanisms that govern the adsorption of contaminants from aqueous media by an adsorbent over time. The linear form of pseudo-second-order model is expressed as (Ho and McKay 1999): t ¼ ðk 2 qe Þqt þ k 2 q2e qt where: k2 refers to the adsorption rate coefficient [g (mg min1)].

ð15:7Þ

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qe refers to the amount of heavy metals adsorbed at equilibrium. qt refers to the amount of heavy metals adsorbed at time t. Intra-Particle Diffusion Model Adsorption can follow multistage diffusional processes, where the rate of contaminant removal is determined by using a square root of time as shown below: qt ¼ kid t 1=2 þ C i

ð15:8Þ

where: kid refers to the intra-particle diffusion coefficient (mgg1 min1/2). Ci refers to the rate constant of the intra-particle diffusion. Those parameters are used to determine the rate-limiting step for intra-particle diffusion. A plot will be used to determine the rate governing the removal of contaminants from the aqua-sphere. If a plot passes through the zero point and has one stage, one mechanism is the rate-limiting step, and when multistage is seen, multistage adsorption is the rate-limiting step (Tran et al. 2017).

Adsorption Isotherms Adsorption of contaminants from aqueous solution onto an adsorbent can be governed by a mono-site or multi-site adsorption. The removal of contaminants from aqueous solution can be determined using the adsorption isotherms. Mainly, Langmuir and Freundlich adsorption isotherms are prevalently used (Tran et al. 2017; Lima et al. 2015). Langmuir Adsorption Isotherm Langmuir developed a model that predicts the monolayer adsorption of contaminants from aqueous solution. The Langmuir adsorption isotherm can be expressed by the following linear equation: Ce 1 C ¼ þ e Qe Qm b Qm

ð15:9Þ

where: Ce ¼ Final concentration (mg L1) Qe ¼ adsorption capacity at equilibrium (mg g1) Qm ¼ Langmuir constants denoting adsorption capacity (mg g1) A plot of Ce vs. Ce/Qe can be used to determine the constant. The straight-line equation will be used to determine the Qm (slope) and C, the intercept.

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Freundlich Adsorption Isotherm Heterogeneous adsorption of contaminants from aqueous medium can be modelled using the Freundlich adsorption isotherm. The Freundlich adsorption isotherm can be expressed by the following linear equation (Freundlich, 1906): log qe ¼ n log Ce þ log k f

ð15:10Þ

where: Ce ¼ Adsorbate at equilibrium (mg L1) qe ¼ Contaminants adsorbed by an adsorbent (mg g1) K ¼ The partition coefficient (mg g1) n ¼ Adsorption degree A plot of log Ce vs. log qe can be used to determine the fitness of the adsorption mechanism. The K value denotes that the energy required for the adsorption process is independent of surface coverage and n is employed to highlight the rate of adsorption. Thermodynamics To determine the nature of adsorption, thermodynamic model was employed. The free energy change (ΔG) needs to be evaluated. The Gibbs free energy change for the adsorption of contaminants can be expressed by the following equation: ΔG ¼ RT ln K c

ð15:11Þ

where: R refers to the gas constant (8.314 J mg1 K1). T refers to the temperature in Kelvin (K). Kc refers to the equilibrium constant (Kc ¼ qe/ce). The positive ΔG value indicates a spontaneous and feasible reaction whereas the negative value indicates a non-spontaneous and unfeasible.

Desorption Study After adsorption, a desorption process can be adopted to remove contaminants from the adsorbent. As such, the amount of contaminants desorbed from the matrices of the adsorbent is calculated using the following equation (Vu et al. 2018): %Desorption ¼

qe qr   100 qe qd

ð15:12Þ

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Table 15.3 Comparison of the adsorption capacity of different adsorbent used for pollutants removal Adsorbent Biochar Magnetite Zeolite Magnetic Fe3O4– chitosan@bentonite (Fe3O4CS@BT) composites Granular activated carbon (GAC) Chitosan–lysozyme biocomposite (CLC) Kaolinite nanotubes Vermiculite

Metals Pb2+, Cu2+, Ni2+, and Cd2+ Fe2+, Pb2+, Zn2+, Ni2+, Cd2+, and Cu2+ Pb2+, Cu2+, and Ni2+

Adsorption capacity 200 mmol/ kg 101 mg/g 299 mg/g

Cd2+, Cr3+, Cu2+, Fe2+ Zn2+, Ni2+, and Pb2+.

62 mg/g

Cd2+, Cu2+, and Zn2+

7.36 mg/g

Cr3+

216 mg/g

Zn2+, Cd2+, Pb2+, and Cr6+ Mn2+, Ni2+, Zn2+, Cd2+, and Cu2+

100 mg/g 33 mg/g

Reference Inyang et al. (2012) Karami (2013) Hong et al. (2019) Feng et al. (2019) Eeshwarasinghe et al. (2019) Rathinam et al. (2018) Abukhadra et al. (2019) Abollino et al. (2003)

where: qr refers to the mass of the unremoved contaminants after desorption the desorption process (mg/g). qd refers to the mass of contaminants desorbed from the adsorbent (mg/g).

Comparative Study Different adsorbents have different adsorption capacity for different contaminants. However, this is also dependent on the surface area and available adsorption sites. The affinity of the material toward the adsorbate also plays a critical role. As shown in Table 15.3, different materials have different adsorption capacity. This is also linked to the surface area of the material and their ability to scavenge contaminants.

15.3

Filtration

The removal of heavy metals from acid mine drainage via filtration is primarily achieved with membrane filtration. Membrane filtration utilizes the combination of a driving force and a membrane to achieve separation of dissolved ions from water (Flynn 2009). In wastewater treatment, the most common and well-known

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Table 15.4 Typical operating parameters and conditions for membrane technologies Membrane technology Microfiltration Ultrafiltration Nanofiltration Reverse osmosis Electrodialysis

Operating species removal range

Operating Pressure (kPa)

Flux range (L/m2d)

Energy consumption (kW/m3)

0.1 μm – 10 μm 10 nm – 100 nm 1 nm – 100 nm calcareous mudstone which is attributed to the different geometries of hydrated copper. Similar work was reported by (Kennedy et al. 2010) for the removal of copper and zinc from raw and acid activated clay. The adsorption effeciency of zinc was studied by untreated Algerian bentonite clay (Mohellebi and Lakel 2016), although the maximum adsorption capacity was achieved at 50%

Compositions Istanbul (Turkey) Tainan-Taiwan Guandong-China Shenzhou-China Egypt China Gafsa-Tunisia Turkey Al-Azraq-Jordan Jordan Algeria Morocco Jordan Mexico Caldin est Turkey China

SiO2 92.8 92 90.1 89.6 83.6 82.95 76.8 76.5 75.57 72.5 72.1 72 72 70.38 69.7 65

Al2O3 4.2 3.3 – 2.5 4.24 5.75 3.66 7.25 9.79 11.24 5.3 7.3 11.42 13.52 11.5 17.5

Fe2O3 1.5 1.3 0.3 1.8 1.07 1.14 1.49 3.85 5.08 5.81 3.3.8 4.3 5.81 3.37 0.65 4.8

TiO2 – – 0.4 – – 0.69 – 0.5 – – 0.37 – – – 0.65 –

Table 16.2 Composition of diatomite from different sources Na2O – – – 1.5 – 0.06 1.45 0.45 6.15 7.21 0.65 1.8 7.21 0.17 0.8 0.5

K2O – – – – – 0.06 0.26 0.85 0.62 0.69 0.54 1.2 – 0.3 1 –

CaO 0.6 – 0.5 – 6.17 0.24 6.61 – 1.39 1.48 7.2 10 1.48 0.66 – 1.1

MgO 0.3 – 0.2 1 – 0.12 1.42 – 0.24 0.25 2.6 1 – 0.42 – –

P2O – – – – – – – – – – – – – – – –

LOI 0.5 – 8.5 4.5 4.86 7.93 – 4.86 – 0.64 7.44 2.4 – 11.18 15.3 11.1

References Sarı et al. (2010) Tsai et al. (2006) Peng Yuan et al. (2010) Sheng et al. (2011) Ibrahim and Selim (2012) Yuan et al. (2004) Saidi et al. (2012) Koyuncu (2012) Al-Ghouti et al. (2009) Al-Degs (2000) Safa et al. (2012) Liva et al. (2007) Al-Ghouti et al. (2003) Miretzky et al. (2011) Al-Ghouti et al. (2003) Al-Ghouti et al. (2003)

16 Removal of Dyes and Heavy Metals with Clays and Diatomite 545

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after 5 min at room temperature and pH 7.42. This endothermic removal of Zn2+ was fitted by a monolayer and homogeneous surface according to results of the Langmuir isotherm (R2 ¼ 0.9980). Dankova Z et al. described the removal of Cu2+ by using natural and modified siderite and kaolin. The value of the adsorption capacity of kaolin modified by MnO2 was achieved 39.5 mg/g which is more important than siderite (10.8 mg/g) (Danková et al. 2015). In another study, the authors reported a comparison of the adsorption of copper, cobalt, and cadmium on natural and pillared clay. Pillared clay is more concentrated and richer on cerium than natural clay which present the better adsorbent for removing these heavy metals (Mnasri-ghnimi and Frini-srasra 2019). In fact, to improve the removal of pollutants from water, the surface of clay is crucial to facilitate its combination with other adsorbents. The following subsections describe and discuss some examples of grafting groups introduced on clay during the recent years for dye and heavy metal removal. Many studies have been reported on using a biopolymer for modified clay such as cellulose (Delhom et al. 2010; Kamel 2018). Adsorption of Cr (II) has been reported by using a cellulose/montmorillonite (Kumar et al. 2012) and showed a maximum adsorption capacity of 22.2 mg/g. Recently the complex cellulose/clay was used for cadmium and lead removal (Kim and Tripathi 2019). The maximum of the adsorption capacity reached 389.78 and 115.96 mg/g for Pb (II) and Cr (II), respectively. Langmuir and pseudo-second-order kinetic model were used to describe the adsorption model. Chitosan is a derivative of chitin and present the second natural polysaccharide in earth (Putro et al. 2017). The abundance, the biocompatibility, the low cost, and the high reactivity of this renewable polymer present the most characteristics to be an important adsorbent combined with clay and used for heavy metals removal (Labidi et al. 2019). The adsorption of copper and chromium were reported recently (Kameda et al. 2019) by using chitosan modified clay. The adsorption of Cu (II) and Ni (II) increased after addition of montmorillonite at high temperature and at efficient time. The maximum adsorption capacity was between 0.18 and 0.33 mmol/g for Cu (II) and 0.23–030 mmol/g for Ni (II) and followed Langmuir adsorption isotherm. Recently nanotube of kaolinite synthesized by ultrasonic scrolling was used for removal of zinc, cadmium, lead, and chromium. The maximum adsorption capacity achieved was 103 mg/g, 116 mg/g, 89 mg/g, and 91 mg/g for Zn (II), Cd (II), Pb (II), and Cr (II), respectively (Abukhadra et al. 2019a). Alginate present one of the most biopolymer used for modified clay by several studies (Ely et al. 2011). Lead and cadmium were removed by using a new formulation of hydrogel-based Poly (methyl methacrylate) grafted alginate onto bentonite. The result showed the best correlation with a Langmuir model with a maximum monolayer adsorption was 60.24 mg/g and 48.78 mg/g for Pb (II) and Cd (II) respectively (Hasn, Imran, 2018). Table. 16.3 summarize the maximum adsorption capacity of the most recent research for removal of heavy metal ion from natural and modified clay. In another work, (Kong et al. 2019) proposed a lignin xanthate resin-bentonite clay composite as efficient adsorbent of Hg (II) in aqueous solutions. Kong’s group used this clay derivative for Hg (II) removal, the authors showed that the adsorption capacity of Hg (II) on lignin xanthate resin-bentonite clay composite (36.45 mg/g) was much higher than that on bentonite (21.29 mg/g) due to lignin

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Table 16.3 Recent studies for heavy metal removal onto raw and modified clay Heavy metal adsorbate Adsorbent Unmodified clay Cd (II) Bentonite Ni (II) Cu (II) Pb (II) Ni (II) Kaolinite clay Mn (II) Pb (II) Raw bentonite Zn (II) Raw bentonite Cu (II) Raw montmorillonite Cu (II) Bentonite clay

Conditions

Qmax (mg/g)

References

pH 8 pH 6

Anna et al. (2015)

pH 8 pH 8 pH 3–9 –

31.25 26.32 32.26 85.47 166.67 111.11 20.68 1.75 18.69 11.89

Modified clay Cu (II) Sodium alginate/hectorite clay Pb (II) Xanthan gum/montmorillonite

pH 55 pH 4

160.28 187.08

Hg (II) Pb (II)

Nanocrystalline cellulose/ bentonite

pH 2 pH 3

Cu (II)

Dactyloctenium aegyptium biomass/montmorillonite Lignin/montmorillonite

pH 6

0.23 0.22 (mmol/g) 18.51

Chitosan/silver nanoparticles/ montmorillonite Lignocellulose/ montmorillonite Chitosan/bentonite Starch/rectorite clay Alginate/bentonite Alkaline/montmorillonite

pH 7

1.0803 mmol/ g 181.50

pH 6.8

94.86

pH 2 pH 5 pH 4–5 pH 6

89.13 180.8 94.04 34.58 32.35 68. 66 68.39 62.15 77.80 49.08 38.8 52.9 55.5 119

Pb (II) Cu (II) Ni (II) Cr (II) Pb (II) Cu (II) Mn (II) Ni (II) Cu (II) Zn (II) Cd (II) Pb (II) Hg (II) Cr (II) Hg (II) Pb (II) Cu (II)

pH 6

pH 10

Lipopeptides produced from solid-state fermentation/NAmontmorillonite

pH 5

4-Aminoantipyrine/bentonite

pH 4

2-Oxyhydrazino-N(2-methylen-ylhydroxyphenyl)pyridinium/ Na- montmorillonite

pH 6

Dawodu and Akpomie (2014) Guerra et al. (2013) Taylor et al. (2013) Taylor et al. (2012) Bertagnolli et al. (2011) Tong et al. (2019) Mirza and Ahmad (2018) Putro et al. (2017)

Umran Khan et al. (2017) Yanli Ma et al. (2017) Azzam et al. (2016) Zhang and Wang (2015) Liu et al. (2015) Wang et al. (2015) Tan et al. (2014) Akpomie and Dawodu (2018) Zhu et al. (2013)

Qihui Wang et al. (2011) Abou-el-sherbini and Hassanien (2010) (continued)

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Table 16.3 (continued) Heavy metal adsorbate Hg (II)

Adsorbent 2-(3-(2-Aminoethylthio) propylthio)ethanamine/ hectorite 2-(3-(2-Aminoethylthio) propylthio)ethanamine/ montmorillonite

Conditions pH 4

Qmax (mg/g) 54.7

References

46.1

Phothitontimongkol et al. (2009)

xanthate resin intercalated bentonite clay composite. This new inorganic adsorbent exhibited a high removal degree (over 99%) using mercury (II) solution of 10 mg/L. They confirmed that the adsorption mechanisms of Hg (II) was mainly due to the complexation of Hg (II) with the functional groups in the lignin xanthate resin– bentonite clay composite. The authors concluded that lignin xanthate resin–bentonite clay composite used for mercury (II) removal was environmentally friendly and appropriate material for water purification. Regarding natural and synthetic dye removal by clay derivatives, many works have been reported in the literature. A cationic dye such as basic yellow 2 was reported to be removed by montmorillonite clay. Langmuir model was reported and showed a good correlation with this study and present a maximum monolayer adsorption capacity in the order of 434.196 mg/ g. The results showed also a endothermic characteristic and fitted by the pseudosecond-order models (Sözüdo et al. 2015). The adsorption of Rhodamine 6 G was evaluated by using three different forms of clay such as montmorillonite, kaolinite, and polygoskite (Li et al. 2018). The removal of Rhodamine 6 G was physically endothermic and present the highest adsorption value in case of using montmorillonite clay form. To evaluate the removal of anionic and cationic dyes such as acid brown 75 and basic yellow 28, a Tunisian researcher’s group’s used a natural smectite clay (Chaari et al. 2019). The results fitted with Langmuir model and present the highest adsorption capacity for basic yellow 28 (76.92 mg/g) in the basic pH rang compared with acid brown 75 (8.33 mg/g) in acid conditions. The adsorption of methylene blue was reported recently (Allam et al. 2018) by using raw and activated clay. The results fitted with a Langmuir model with increasing of adsorption capacity of raw clay from 30 mg/g to 50.22 mg/g for activated clay. Several studies used surfactant modified clay for dye removal (Anirudhan and Ramachandran 2015; Gamoudi and Srasra 2019; Xiang et al. 2019). Reactive red 120 was removed by an electrochemical conductivity technique using two surfactant-modified montmorillonite (Mahmoodi et al. 2019). Natural clay can be modified using ceramic ultrafiltration membranes made (Ouaddari et al. 2019). Purified clay was used for the removal of direct red 80 dye. A new methodology was developed recently by using graphene oxide to modified clay by three different methods (Gogoi et al. 2019). This combination of graphene and clay present an effective adsorbent for methylene blue removal compared with raw clay and by only graphene oxide. This adsorbent present 85%, 75%, and 75% of adsorbent yields for

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three cycles, respectively. Methylene blue was removed by using cellulose modified clay (Hu et al. 2019; Liu et al. 2018). The result of using carboxymethyl/cellulose/ carrageenan/activated montmorillonite presents 92% yield of methylene blue adsorption at pH between 6 and 10 and at 45  C. This study was fitted by Langmuir model and followed the pseudo-second-order model of adsorption. Table. 16.4 summarizes the maximum adsorption capacity of the most recent studies of dyes removal by natural and modified clay. In agreement with the results reported by (Gamoudi and Srasra 2019), a modified clay obtained from Na+-exchanged clay can be used as a potential adsorbent for methyl orange, indigo carmine, and phenol red removal. The authors have used the X-ray diffraction analysis, Fourier transform infrared spectroscopy, and surface charge density δH to confirm the successful modification of clay. They noted that the modification of the Na +
exchanged clay by cationic surfactants hexadecylpyridinium (HDPy+) by integration into the interlayer space can be monitored by following the clay expansion. The X-ray diffraction analysis analyses show the d001 basal spacing of modified clay which has moved toward higher value after the addition of the surfactant that demonstrated the effective intercalation of surfactant in the clay layers. The Na +
exchanged clay showed peaks at 1.23 nm. Two peaks were perceived for hexadecylpyridinium–clay corresponding to the d002 and d001 reflection: the first peak (d002) is showed at 2.09 nm and the second peak (d001) at 4.39 nm. The Fourier transform infrared spectrophotometry spectra of dye-loaded hexadecylpyridinium–clay disclosed the existence of a novel bands which are more intense and the organoclay structure. The novel peaks and the change of peaks intensities confirm the adsorption of methyl orange, indigo carmine, and phenol red dyes. This adsorbent was characterized also by using the surface charge density δH; the authors suggested that the rise in PZC after treatment can be explicated by the significant positive charge of hexadecylpyridinium–clay with respect to clay. Adsorption experiments showed that hexadecylpyridinium–clay unveiled remarkable adsorption capacities. In 30 and 60 min, 0.1 g of material adsorbed 227.27, 326.40, and 344.82 mg/g for methyl orange, indigo carmine, and phenol red, respectively, with a dye concentration of 200 mg/L. The author noted that the adsorption of methyl orange dye on hexadecylpyridinium–clay was pH-independent. They indicate that the indigo carmine and phenol red adsorption capacity decreased slightly with the increasing of pH solution. Based on the proposed mechanism, dye adsorption in this study is governed by physical and electrostatic forces which is related to the dye structure, the molecular size, and the functional groups. In other works (Xiang et al. 2019), the removal of bromophenol blue using organoclays modified with imidazolium-based gemini surfactant (C14-4-C14im-Vt, C14-4-C14im-SiNSs and C14-4-C14im-Mt) was studied. The adsorption capacities were 400.19, 230.77, and 220.27 mg/g, respectively. This clearly demonstrated the important role of imidazolium surfactants to enhance adsorption capacity of organoclays. Moreover, C14–4-C14imSiNSs possesses present a higher adsorption efficiency (R ¼ 99.53% at pH ¼ 2, C0 ¼ 200 mg/L). All these findings were important for the forthcoming advance of clay-based material for heavy metals and dyes adsorption from wastewater.

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Table 16.4 Recent studies for dyes removal onto raw and modified clay Dyes adsorbate Adsorbent Unmodified clay Methylene Raw clay blue Methylene Natural clay blue Direct Raw clay orange 34 Methylene Raw clay blue Methylene Natural clay blue Malachite green Methyl orange Cationic Natural untreated clay basic yellow Basic red 2 Raw clay Basic red 18 Methyl violet Modified clay Methyl orange

Acid orange Methylene blue Reactive violet Auramine O Amido black (10B) Direct rose FRN Toluidine blue Crystal violet

Conditions

Qmax (mg/g)

pH 12

30

pH 10

322.58

pH 2

2.68

pH 9

56.85

pH 7.5 pH 6.5

95.50 109.82 88.70

pH 5.2

833.33

–

200

Natural montmorillonite

pH 4

530.645

Ghassoul

pH 10

625

Dioctadecyl tetrahydroxyethyl dibromopropane diammonium and octadecylmethyldihydroxyethyl ammonium bromide surfactant modified montmorillonite Cetyltrimethylammonium bromide / Montmorillonite Cellulose/ Montmorillonite

pH 5

250.63 91.11

Peng et al. (2019)

pH 2

98.59

pH 7

227

Bentonite silylated aminopropyltrimethoxy silane Cellulose/montmorillonite

pH 2

107.4

pH 7 pH 3 55  C

1338.2 232

Shen et al. (2019) Wang et al. (2019b) Queiroga et al. (2019) Pan et al. (2019)

Chitosan/clay

pH 10

17.18

Na-montmorillonite

pH 11

5.80 5.40 (mmol/g)

References Allam et al. (2018) Bentahar et al. (2018) Islem et al. (2016) Bendaho and Driss (2015) Elmoubarki et al. (2015)

Öztürk and Malkoc (2014) Mohammed (2013) Fil et al. (2013) Elass et al. (2011)

Kausar et al. (2019) El Haouti et al. (2019)

(continued)

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Table 16.4 (continued) Dyes adsorbate Methylene blue Methylene blue Basic blue 41 Congo red Methyl orange Acid blue 113 Acid red Crystal violet Telon dyes (orange– red–blue) Methylene blue Congo red Acid orange 7 Acid red 17 Reactive dye Crystal violet Methylene blue Crystal violet Safranin T Acid scarlet G Methyl orange Alizarin red Basic yellow 28

Adsorbent Chitin/clay

Conditions –

Qmax (mg/g) 156.7

Lignin/montmorillonite (hydrogel)

pH 7

9646.92

Acid activated clay

–

73

Tragacanth gum/poly (methyl methacrylate)/bentonite

pH 8.5

900 750 8.5

Hexadecyltrimethylammonium bromide/bentonite Kappa-carrageenan (poly vinyl alcohol)/Na-montmorillonite Imidazolium/bentonite

pH 3 –

140.84 (μmol/g) 151

pH 7.3

88–108

pH 12 30  C pH 10

303

Plasma surface modified bentonite TiO2/acid activated kaolinite

12.36

Dimethyl dialkyl amine/ montmorillonite

–

89.31% 82.13%

Starch–montmorillonite/polyaniline

pH 2

91.74

Alginate acid activated bentonite

pH 8–10

582.4

Sulfuric acid activated clay

pH 12 30  C

223.19

Kappa–carrageenin/Na-alginate/Namontmorillonite N-Vinyl/2-pyrrolidone itaconic acid/ organoclay Cellulose acetate/organo montmorillonite Glycol/bis-N-cetylnicotinate dibromide/bentonite Iron oxide activated clay

pH 6.85

88.8

pH 6

550

pH 1

85.7

–

99.02%

–

32.7

pH 6

514

Granular inorganic organo pillared clays

References Xu et al. (2018) Zhao et al. (2017) Kooli et al. (2015) Sadeghi et al. (2015)

Mullassery et al. (2014) Hosseinzadeh et al. (2015) Makhoukhi et al. (2015) Şahin et al. (2015) Hai et al. (2015) Aydin Hassani et al. (2016) Olad and Azhar (2014) Oladipo and Gazi (2014) Auta and Hameed (2013) Reza et al. (2013) Gök and Güçlü (2013) Zhou et al. (2012) Kan et al. (2011) Fu et al. (2011) Cheknane et al. (2010)

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Diatomite Based-Material

Due to its specific properties, diatomaceous earth has been extensively applied in adsorption process to remove pollutants from water. It has been shown that diatomite contains different impurities, to make it more inert, diatomaceous earth was purified in HCl (Goren et al. 2002) and calcined (Khraisheh et al. 2005) for further use. Recently, the surface modification of diatomite by several methods has been investigated. Diatomaceous earth is a potential decontaminating agent for heavy metals. FloresCano et al. used natural diatomite from a mine in Jalisco (Mexico) and investigated its adsorption capacity toward heavy metals (Cd (II), Zn (II), Pb (II), and Cr (III)) (Flores-Cano et al. 2013). Before used, this diatomite was washed for several times and dried in an electric oven at 100  C. The characterization of their diatomite was carried out with different techniques and present the pore diameter of diatomite was 0.33 mm. The chemical composition by X-ray diffraction technique showed that this Mexican diatomite contains 95.91% of SiO2 (quartz and cristobalite). The maximum uptake of heavy metals onto the purified diatomite was 0.734 mg/g for Cd (II) at pH 7, 0.232 mg/g for Zn (II) at pH 6, 0.169 mg/g for Pb (II) at pH 4 and 0.162 mg/g for Cr (III) at pH 4. This adsorption process was endothermic because the adsorption capacity of diatomite increases by increasing the temperature. The increase of pH from 4 to 7 shows that the maximum capacity increased 6.3 times. In recent works modified diatomite was modified with chitosan gel to obtain a novel adsorbent for the removal of Hg (II) (Caner et al. 2015). This natural diatomite was obtained from the Beg-Tug Industrial Minerals Mines Company in Turkey; it contained essentially of 92.8% SiO2, 4.2% Al2O3, 1.5% Fe2O3, and other metal oxides. The BET surface area of the natural diatomite was 0.56 m2/g, while it was measured as 0.96 m2/g for the sorbent. The monolayer sorption capacities of diatomite before and after modification were 68.1 and 116.2 mg/g, at pH 5, respectively. The mean adsorption energy was calculated from the Dubinin–Radushkevich (D–R) model which was founded 8.2 kJ/ mol indicating a chemisorption process. The single-point sorption total pore volume of diatomite/chitosan was 0.0088 cm3/g as it was measured as 0.011 cm3/g for the prepared diatomite/chitosan biocomposite sorbent. The surface of diatomite modified with chitosan consists of micro and meso particles. During the sorption process, these holes were occupied by Hg (II). Another modification with chitosan was carried out by Salih et al. for the removal of Zn (II) ion (Salih and Ghosh 2018). The chitosancoated commercial diatomaceous earth (beads) was synthesized by a dropwise method. Their diatomaceous earth powder contained 92% SiO2, 4% Al3O5 and 2% Fe2O3. The BET surface area of diatomite/chitosan beads was 6.4 m2/g, which is more important than the Turkish diatomite modified by chitosan gel by Caner et al. The pore size was 3.23 10–6 mm and the pore volume was 7.4 10–9 m3/g indicating a high porous structure. The maximum uptake adsorption of Zn (II) was found at pH 6 (127.4 mg/g). The most used and successful modification was that of metal oxide. Metal oxide was introduced onto the matrix of diatomaceous earth due to their adsorption

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properties. It represents the common and effective method because it showed an important effect of the adsorption of heavy metals. Al–Degs tested the modification of natural Jordanian diatomite by oxide manganese for the adsorption of Pb2+, Cu2+, and Cd2+ (Al-Degs et al. 2001). In other works, H. Ma et al. have used natural and modified diatomite for the removal of Cr (VI). They confirmed that the Cr(VI) adsorption effect on acid modified diatomite was better than on natural diatomite (Ma et al. 2019). The maximum adsorption capacity obtained from Langmuir isotherm model was 14.4 mg/g at 298 K. Furthermore, the values of thermodynamic parameters indicate an endothermic and non-spontaneous adsorption process. Recently, Nefzi. H and co-workers reported the preparation of palygorskite clay from Tunisian diatomite for the adsorption of Ni2+ and Cu2+. The effects of contact time, initial concentration, temperature, and pH on the adsorption process were investigated. The results showed that the (%) of maximum adsorption capacity of diatomite was 78.44% for Cu2+ at pH 4 and 77.3% for Ni2+ at pH 7, while the (%) of the maximum adsorption on palygorskite reached 91% for Cu2+ and 87.05% for Ni2+, under the same condition (Nefzi et al. 2018). The results of the various works confirmed that diatomite has a good and important potential to be used as an efficient adsorbent to remove heavy metals from effluents. Due to its characteristics, cellulose –diatomite beads were used in various works to remove dyes using adsorption process. The modification of diatomite by maleic anhydride was used to enhance the adsorption capacity of dye (Li et al. 2014). The new material had a strong adsorption capacity toward basic dyes. The maximum adsorption capacity increased from 51.6 to 116.6 mg/g for methylene blue, depending on the initial concentration of the dye. A similar trend was also found for methylene violet adsorption increased from 30.5 to 61.1 mg/g. Hongjie Dai and co-workers confirmed that the use of cellulose increased the capacity adsorption of diatomite (Dai et al. 2019). In their study, they reported a green method to elaborate hydrogels with pineapple peel cellulose/magnetic diatomite. The combination of the magnetic iron oxide with diatomite was performed to make an hydrogels more efficient for the adsorption process. The adsorbents based on cellulose obtained from agricultural residues has been used for the adsorption of methylene blue. The results showed that the prepared hydrogels showed an excellent stability and reusability for adsorption. Compared with the first adsorption, the recovery efficiency of the hydrogels surpassed 90% after four cycles of adsorption/desorption. Therefore, these hydrogels can be employed as adsorbents for efficient removal of dyes. Chitosan was used in different researchers; Peng Zhao and co-workers prepared a new chitosan/diatomite composite obtained by simple mixing in the mass ratio for the removal of methyl orange from aqueous media (Zhao et al. 2017). These results revealed that the chitosan/diatomite might be considered as an excellent adsorbent. Thus, the removal rate reached 88.37% and the desorption efficiency reached 98.1%. Most importantly, the data obtained in this study showed that there is no significant decrease after six times, particularly in the sorption capacity of the chitosan/

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diatomite composite, with satisfactory adsorption performance. Compared to these results, other work confirmed the efficient of a novel diatomite/chitosan-Fe(III) composite which was synthesized successfully to be used as an adsorbent for the removal of anionic azo dye. The new material was proven to be a promising adsorbent for the removal of anionic azo dyes from wastewater (Zheng et al. 2014). The modification of diatomite with metal oxide represents the most used and successfully one. It is the common and effective method due to the important adsorption of pollutants. Ningning Liu, for example, used magnesium oxide to modify diatomite waste as an adsorbent for organic dye removal (Liu et al. 2019). The diatomite was dispersed in MgCl2 solutions in a 100 mL beaker performing ultrasonic oscillation for 30 min. Then, NaOH solutions (2.25 mol/L) were added into the mixture. Afterward, the product was stored at room temperature, followed by being filtered and washed with distilled water. This prepared adsorbent was used for the removal of cationic and anionic dyes. The results of this work showed that diatomite modified by metal oxide was an effective adsorbent for dyes, which is predominantly attributable to its high surface area and active sites. Thus, the maximum adsorption capacity of modified diatomite of methylene blue reached 25.02 mg/g and 35.13 mg/g for the removal of acid orange. Numerous works have been carried out to enhance the adsorption capabilities of natural diatomite by integrating it in composites with various hydroxides nanostructures. Previous study confirmed that the incorporation of nickel in the porous materials produced a hybrid materials which enhances adsorption capacity. They used Ni/NiO nanoparticles as hybrid material to increase the adsorption of malachite green and also hexavalent chromium (Abukhadra et al. 2019b). The results showed that adsorption equilibrium was reached after 660 min for both compounds. The adsorption equilibrium curves of dye by Ni/diatomite were categorized as L-type isotherm which reflected the high affinity between the ions and the adsorbents. Moreover, the equilibrium data fitted principally with the Langmuir model suggesting a monolayer adsorption phenomenon. The estimated parameters of Dubinin–Radushkevich gave adsorption energies values of 5.64 KJ/mol, corresponding to physisorption mechanism. The values of free energy and enthalpy determined by the thermodynamic studies showed that the adsorption of both ions are spontaneous endothermic reactions of physical nature. Recently, Du et al. (2018) performed the experiment of the adsorption and photoreduction of Cr (VI) using diatomite modified by Nb2O5 nanorods. In this study, the authors investigated serval characterization techniques of the synthesized material and others for the visualization and detection of the phenomenon of Cr (VI) adsorption (Du et al. 2018). Scanning electron microscopy and transmission electron microscopy analyses indicate that Nb2O5 first heterogeneously nucleates on the surface of diatomite, before it propagates with a trend for oriented crystallization. The morphology of Nb2O5 changes from flower-like to rod-like. Nb2O5 nanorods deposited on the edge of diatomite. The adsorbents resulting was amorphous, after the hydrothermal treatment diffraction peaks of crystalline Nb2O5 were observed, signifying that pure Nb2O5 crystals occur on the diatomite. The Nb2O5 crystalmodified diatomite is a performant adsorbent for rapid removal of Cr (VI) from

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555

wastewater at ambient temperature. Cr (VI) adsorption capacity was 115 mg/g and can be improved with ultraviolet irradiation, due to the synergistic effect of surface adsorption and Nb2O5 photoreduction. Du et al. studied the adsorption of Cr (VI) using Fourier transform infrared spectrophotometry and XPS of the adsorbents before and after Cr (VI) photoreduction to improve the capacity of the synthesized material toward Cr (VI) adsorption and photoreduction. After the adsorption, a new peak at 813 cm
1appears due to the stretching vibration of Cr -O bonds. The XPS indorses that Cr (VI) is transformed to Cr (III) by Nb2O5/diatomite photoreduction. For the degradation of Cu2+ in aqueous solutions, (Sun et al. 2015) reported the synthesis of nano-TiO2/diatomite composite materials by the hydrolysis deposition of titanium tetrachloride. The composite analysis confirms that the TiO2 was immobilized on diatomite to obtain the TiO2- diatomite for Cu2+ photoreduction. The removal efficiency of 10 mg/L Cu2+ solution by the composite reaches 96.63% after irradiation for 3 h under UV light.

16.5

Adsorption Kinetics, Equilibrium, and Thermodynamic

16.5.1 Kinetics Studies To study dye and heavy metal removal, adsorption kinetics is one of the important parameters to describe the adsorption process. Adsorption of dyes and heavy metals on clay and diatomite derivatives was studied by means of the pseudo-first-order and the pseudo-second-order and the intra-particle diffusion models. The pseudo-first-model was related to the kinetics of one-site adsorption governed by the rate of the surface reaction and is expressed as follows (Eq.16.1): dq ¼ K 1 ð qe
q t Þ dt

ð16:1Þ

The linear form of the equation (1.1) is given as follow (Eq. 16.2): log ðqe
qt Þ ¼ log qe


K1 t 2, 203

ð16:2Þ

qe and qt are the adsorption capacity of each pollutant (mg/g) at equilibrium and time (t), respectively; K1 (1/min) is the adsorption constant. The plot of log (qe
qt) against time (t) was used to determine the constant (K1) values for each pollutant removal. The pseudo-first-model was generalized to two-site-occupancy adsorption to form a pseudo-second-order and can be expressed by the next equation (Eq. 16.3):

556

N. Tahari et al.

dq ¼ K2 ðqe
qt Þ2 dt

ð16:3Þ

The linear form of equation (1.3) is given by the following formulation (Eq. 16.4) and at the boundary conditions, q ¼ 0 at t ¼ 0 and q ¼ qt at t ¼ t: t 1 1 ¼ þ t qt k 2 qe 2 qe

ð16:4Þ

where k2 is the pseudo-second-order rate constant (g/mmol min). The plot of the variation of (qt ) vs. (t) permits to determine the constant K2. t Weber and Morris (1963) suggested that a plot of the adsorption capacity. The square root of contact time would give a straight line if the pore diffusion was the rate limiting step of the adsorption process. It can be expressed in the following form (Eq. 16.5): qt ¼ kp t =2 þ C 1

ð16:5Þ

kp (mg/g min1/2) is the constant of intra-particle diffusion model and C (mg/g) is a constant, which is related to the thickness of the boundary layer. The variation of the curve of qt ¼ f (t) at various concentrations of each pollutant allows the determination of the controlled adsorption process parameters.

16.5.2 Adsorption Isotherms The adsorption equilibrium of each dyes and heavy metals is evaluated by means of isotherm curves. The representation gives the relationship between the mass of dyes and heavy metal adsorbed on the clay and diatomite-based material (qe, mg/g) and the residual quantity of dye and metal in the liquid phase (Ce, mg/L), under equilibrium conditions at a set temperature. Curve shapes can provide indication regarding the adsorption mechanism. A large number of models of adsorption isotherms have been developed to describe the adsorption equilibrium results. Different parameters of the adsorption were calculated from the mathematic models developed by Langmuir, Freundlich, Dubinin–Redushkevich (D–R), Temkin, Redlich–Peterson (R–P), and others using the linear and nonlinear equations (Table. 16.5). Langmuir isotherm model (Langmuir 1918) is the simplest theoretical model for monolayer adsorption onto a surface with finite number of identical sites. The adsorbent is homogeneous, and the interaction forces among adsorbed compounds are insignificant.

qe ¼

Redlich–Peterson (R-P)

K RP C e 1þARP C e α

2

qe ¼ qm e
βε

Temkin

Dubinin–Radushkevich (D-R)

1 n

qe ¼ K F C e qe ¼ RT bT ln ðKT C e Þ

qe ¼

K l qm C e 1þK l C e

Nonlinear form

Freundlich

Isotherm Langmuir

Table 16.5 List of adsorption isotherm models

m

m

¼ qce þ kl1q

qe ¼

ln ðKT Þ þ

RT bT ln ðce Þ 2

e

Ln qe ¼ Ln qm
βε   ln K RP Cq e
1 ¼ αlnðC e Þ þ ln ðARP Þ

RT bT

log qe ¼ 1n log ce þ log kF

ce qe

Linear form e

e

Ln qe vsε2   ln K RP Cq e
1 vs ln ðC e Þ

qevsln(Ce)

logqevsCe

Plot of linear form ce q vsC e

KR, AR, α

qm, β, ε

b T, K T

kF, n

Parameters Kl, qm

16 Removal of Dyes and Heavy Metals with Clays and Diatomite 557

558

N. Tahari et al.

Langmuir model is defined by (Eq. 16.6) qe ¼

K l qm C e 1 þ K l Ce

ð16:6Þ

where qm is the maximum adsorption capacity (mg/g), Kl (L/mg) is the Langmuir coefficient, and Ce is the equilibrium of each pollutant concentration in solution (mg/L). Freundlich isotherm (Freundlich 1906) is an empirical equation appropriate to nonideal adsorption on heterogeneous surfaces (Eq. 16.7): 1

qe ¼ K F C e n

ð16:7Þ

where qe is the equilibrium dye concentration on adsorbent (mg/g), Ce is the equilibrium concentration of each pollutant in solution (mg/L), KF (mg/g) (L/g) and 1/n is the Freundlich constant related to sorption capacity, and n is the heterogeneity factor. Temkin isotherm (Temkin 1940) is based on the assumption that the heat of adsorption decreases linearly with the increase of coverage of adsorbent due to adsorbate/adsorbate interactions. Temkin’s equation can be expressed as follows (Eq. 16.8): qe ¼

RT ln ðK T Ce Þ bT

ð16:8Þ

where qe is the equilibrium of each pollutant concentration on adsorbent (mg/g), Ce is the equilibrium concentration of each pollutant in solution (mg/L), and KT (L/g) and bT are the Temkin constants. Dubinin–Radushkevich (R–D) isotherm model (Dubinin 1955) does not, unlike the Langmuir isotherm, consider a homogeneous surface or constant adsorption potential and is defined by the next equation (Eq. 16.9): qe ¼ qm e
βε

2

ð16:9Þ

where qe is the amount of each pollutant adsorbed per unit weight of adsorbent (mg/g), qm is the maximum adsorption capacity, β is the activity coefficient useful in obtaining the mean sorption energy E (Kj mol
1), and ε is the Polanyi potential. The Redlich–Peterson isotherm (Redlich 1958) is a three-parameter isotherm which includes the features of Freundlich and Langmuir models (homogeneous or heterogeneous adsorbent surface). The Redlich–Peterson isotherm has the following formulation (Eq. 16.10):

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Removal of Dyes and Heavy Metals with Clays and Diatomite

qe ¼

K RP C e 1 þ ARP Ce α

559

ð16:10Þ

KRP (l/g) and ARP (mg/L)-g are the Redlich–Peterson constants, α is dimensionless exponent whit value between 0 and 1, and Ce is the equilibrium concentration of each pollutant in solution (mg/L). In addition to the nonlinear form, the different parameters in the five isotherm models were assessed by means of the linear method. In each case, corresponding correlation coefficients (R2) were determined and compared to identify the best isotherm form for each pollutant adsorption (Nebaghe et al. 2016). For the Langmuir, Freundlich, Dubinin–Redushkevich (D–R), Temkin and Redlich–Peterson (R–P) models, the different linear plottings for each model are given in Table 16.5. In utmost adsorption studies, the linear approach has been employed more than the nonlinear method in assessing the quality of fitting of an isotherm model to experimental data principally owing to its easiness and effectiveness (Safwat and Matta 2018). On the other hand, the study of dye and heavy metal removal can be studied in non-binary system (only one dye or heavy metal in solution) and also in binary system (more than one adsorbate in solution). It is essential to explore the aptitude of clay and diatomite derivatives uptake in binary system with the purpose to study the selectivity of each adsorbents toward the different pollutant removals in the solution. In this case, the used isotherm in the previous section will be transformed into other forms to study the adsorption of each pollutant in solution which contains more one adsorbate. The equation of Langmuir isotherm results from the hypothesis that two adsorbates compete for the same adsorption sites. The extended isotherm from the Langmuir model is given as follows (Eqs. 16.11 and 16.12): qe1 ¼

K l1,1 qm1 C e1 1 þ K l1,1 C e1 þ K l1,2 C e2

ð16:11Þ

qe2 ¼

K l1,2 qm2 C e1 1 þ K l1,1 C e1 þ K l1,2 C e2

ð16:12Þ

where qe1 and qe2 are amount of components 1 and 2 adsorbed at equilibrium; qm1 and qm2 their maximum adsorption capacities (mg/g); Kl1, 1 and Kl1, 2 (l/mg) affinity constants of the adsorbent for each adsorbate (1 and 2), respectively; and Ce1 and Ce2 the concentrations of each adsorbate at equilibrium (mg/L). The extended Redlich–Peterson isotherm has been also used for multi-component systems. The isotherm has the following equation (Eqs. 16.13 and 16.14): qe1 ¼

K RP1 C e 1 þ ARP1 Ce1 α1 þ ARP2 C e2 α2

ð16:13Þ

qe2 ¼

K RP2 C e 1 þ ARP1 Ce1 α1 þ ARP2 C e2 α2

ð16:14Þ

where α reflects the heterogeneity of the adsorbent surface.

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Table 16.5 is a compilation of some of the highest reported adsorption capacities (qm is expressed in mg of pollutant per g of adsorbent).

16.5.3 Thermodynamic Studies Thermodynamic parameters including the Gibbs energy (ΔG0), enthalpy (ΔH0), and entropy (ΔS0) are the indicators to predict adsorption mechanism (i.e., physical or chemical). The physical adsorption results from weak interactions (i.e., van der Waals force); chemical adsorption implicates higher interactions (i.e., coordination) with consequent transfer of electrons between the adsorbent and adsorbate. The  thermodynamic parameters (ΔG0), (ΔH ), and (ΔS0) were determined using the following equations (Eqs. 16.15, 16.16, 16.17, and 16.18): Kc ¼

Ca Ce

ð16:15Þ

ΔG0 ¼
RTLnK c

ð16:16Þ

ΔS0 ΔH 0
R RT

ð16:17Þ

Ln K c

ΔG0 ¼ ΔH 0
TΔS0

ð16:18Þ

Kc is the equilibrium constant, Ca is the adsorbent phase concentration at equilibrium (mg L
1), and Ce is the equilibrium concentration of each pollutant in solution  (mg/L). (ΔG0), (ΔH ), and (ΔS0) were from the linear variation of LnKc versus T1. The standard values of enthalpy change (ΔH0) are used to confirm the adsorption nature (exothermic or endothermic) and to give indication about the interactions between clay and diatomite-based material-dyes/heavy metals. The standard entropy change (ΔS0) is associated with the randomness that happens in the adsorbent surface during the adsorption. The standard values of Gibbs free energy change (ΔG0) are used to substantiate whether the adsorption is spontaneous or favorable.

16.6

Conclusion

In summary, a wide variety of clays and diatomite derivatives have been successfully applied for the removal of dyes and heavy metals from wastewater. Their surface modification by adding different ligands was studied by serval analytic techniques, offering generally good removal percentages of organic and inorganic pollutants. Clay and diatomite are used in decontamination processes due to their great advantages, i.e., reduced cost, biodegradability, and eco-friendly character. Clay and diatomite comprise a diversity of functional groups that might be modified via

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chemical and physical methods. These modifications enhanced the adsorption capacity of clay and diatomite derivatives due to the augmentation of binding sites and inclusion of new moieties on the surface of the adsorbent.

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Index

A Abdel Salam, 356 Abderrabba, M., 540 Abdolali, A., 127, 353 Abdullah, N., 295 Abollino, O., 356, 525 Abukhadra, M., 525 Acid mine drainage, 189, 515, 518, 519, 525, 527, 528, 530, 531 Adsorbents, 21, 39, 44, 51, 58, 60, 61, 64, 125–138, 189, 190, 196–201, 222, 224, 226, 247, 250–252, 303, 317, 329, 330, 342, 343, 345–347, 352–360, 370, 412, 414, 416, 444, 445, 448, 454–456, 458–462, 477, 480, 481, 493–495, 500–502, 516, 519–525, 530, 541–544, 546, 548, 549, 551–556, 558–561 Adsorption, 11, 21, 39, 41, 50, 54–62, 92–94, 106, 124, 125, 128, 130–138, 157, 167, 179, 184, 187, 189–190, 193–201, 213, 215, 218, 222, 224–228, 233–235, 249, 251, 252, 282, 286, 295, 299, 302, 316, 320, 329, 330, 332, 336–338, 340, 342–347, 349, 350, 352–354, 358–360, 363, 364, 367, 368, 374, 375, 377–381, 383, 399–401, 407, 414, 443–457, 459, 461, 473, 477, 479–488, 493–495, 500–502, 516, 517, 519–525, 531, 541, 543, 544, 546, 548, 549, 552–561 Adsorptionent, 541 Adsorption on new adsorbents, 317 Aguiar, I., 305

Ahmad, R., 449 Ahmady-Asbchin, S., 132 Ahmaruzzaman, M., 127 Ahsan, M.A., 446 Ajmal, M., 356, 444 Akaya, G., 356 Akinwekomi, V., 530 Akter, M., 124–138 Albaaj, F., 18 Aldrich, C., 356 Al-Ghamdi, S.G., 165 Alhaj, M., 165 Ali Khan, M.Z., 260–274 Ali, A., 127, 133, 167, 181 Ali, H., 344, 405, 406 Ali, S., 364 Allègre, C.J., 10 Al Omar, M.K., 130 Al-Othman, A., 167 Al-Sa’ed, R., 165 Al-Saydeh, S.A., 487 Altun, T., 134 Aman, T., 236 Amiri, M.J., 54 Ammari, F., 396, 398, 399 An, Y.J., 181 Anantha Singh, T.S., 86–113 Anastopoulos, I., 127, 131 Ancsombe, N., 8 Anjum, M., 359 Anthropogenic sources, 4, 9, 18, 43, 81, 180, 181, 265, 476, 479

© The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature Switzerland AG 2021 Inamuddin et al. (eds.), Water Pollution and Remediation: Heavy Metals, Environmental Chemistry for a Sustainable World 53, https://doi.org/10.1007/978-3-030-52421-0

571

572 Antimicrobial, 4, 44, 158, 167, 283, 304–308, 361 Aparicio, J., 165 Aquino, J.M., 97 Arias-Andres, M., 344 Arief, V.O., 456 Aris, A.Z., 127 Arsenic, 359, 370, 375, 378, 379, 381, 389, 408, 411, 414 Asgher, M., 137 Ashraf, R.S., 408, 472 Aubert, D., 10 Awual, M.R., 135 Ayadi, S., 540 Ayangbenro, A., 181 Azarova, Y.A., 127 Azzam, A.M., 54

B Babalola, O., 181 Babel, S., 249, 356 Badot, P.M., 127 Bagbi, Y., 56 Baiju, A., 105 Balasubramanian, R., 361 Baltrenaite, E., 135 Bansode, P.R., 356 Barakat, M.A., 343, 444, 502 Barbosa Neto, A.G., 344 Barbusinski, K., 127 Barquilha, C.E.R., 353 Basu, A., 13 Beer, H.B., 391, 392 Bello, A.O., 447 Bellona, C., 444 Benaissa, H., 132 Benguella, B., 132 Bertagnolli, C., 457 Bhatia, M., 213 Bhatnagar, A., 127 Bhaumik, M., 134 Bilal, M., 356 Bilton, A.M., 166 Biochar and adsorption, 21, 39, 92, 124, 157, 179, 213, 249, 282, 316, 443, 473, 516, 541 Biofiltration systems, 237, 243–255 Blais, J.F., 181 Borah, R., 131 Borba, C.E., 103, 353 Bostani, A., 56 Boye, B., 102

Index Braun, J.J., 10 Braun, M., 12 Buddhi, D., 127 Bünzli, J.C.G., 8 Burda, K., 14 Burkes, S., 13 Buzzi, D.C., 528 Byrne, R.H., 10

C Cadmium (Cd), 39–44, 51, 52, 57–58, 130, 131, 133, 138, 179, 181, 184, 185, 188–190, 197–200, 222, 244, 245, 253, 260, 261, 263–265, 269–271, 282, 284, 288, 294, 319, 322–324, 328–331, 349, 360, 370, 375, 381, 389, 408–411, 414, 443, 456, 472, 482–484, 488, 491, 493, 546 Calderon, J., 268 Campbell, M., 8 Cañizares, P., 111 Carbon footprint, 162, 171 Cardona, O.D., 19 Carpenter, D., 13 Castro, L., 446 Chandio, T.A., 181 Chaney, R.L., 405 Chang, J.S., 295 Chang, Q., 444 Chao, H.-P., 353 Charalampides, G., 8 Chatterjee, S., 355 Chaudhary, V.K., 260–274 Chauhan, D., 260–274 Chen, G., 353 Chen, L., 268 Chen, P., 396, 399, 400 Chen, S.S., 343, 350 Chen, Z., 17 Chi, T., 131 Chibuike, G.U., 181 Chmielewská, E., 531 Chowdhury, S., 361 Chowdhury, S.R., 329 Chromium (Cr), 39–44, 51, 52, 56, 59–61, 78, 133–134, 138, 179, 181, 182, 184–186, 188, 192, 197, 199, 222, 244, 245, 253, 260, 261, 265, 272, 274, 282–284, 294, 296, 299, 322–324, 331–333, 355, 370, 371, 375, 377, 378, 385, 389, 408–412, 414, 443, 472, 479–480, 486, 489, 493, 500, 546

Index Chuah, T.G., 127 Cifuentes, G., 503 Cifuentes, L., 444 Ciocarlan, R.-G., 343 Cirelli, A.F., 127 Clays, 190, 194, 197–198, 226–228, 248, 249, 251, 321, 324, 328, 334, 355, 356, 382–384, 520, 541–544, 546–551, 553, 555, 556, 559–561 Comninellis, C., 88 Composites, 21, 110, 130, 131, 133, 134, 167, 168, 197, 198, 200, 222, 286, 291, 293, 299, 300, 303, 359, 364, 365, 370, 395, 445, 447, 450, 451, 455, 457–460, 475, 477, 493, 498, 520, 525, 546, 548, 553–555 Crane, R.A., 384 Crini, G., 125, 127, 136, 137, 356 Cristaldi, A., 344, 408 Crystallization, 379, 554 Cui, J.A., 13 Culhane, F., 344 Cutting, S., 344

D Daghrir, R., 99, 100 Dalai, T.K., 10 Danila, V., 54 D’Aquino, L., 13 Das, P., 178–201 Das, S., 181 Dasa, D.P., 371 Davis, A.P., 248 De Almeida Silva, R., 531 Debnath, S., 343 Degradation, 21, 44, 50, 51, 75, 86, 94–99, 102–113, 193, 194, 196, 201, 218–220, 227–231, 235–237, 272, 282, 286, 288, 289, 292–294, 299, 341, 366, 368, 370, 372, 373, 392, 401, 407, 415, 515, 541, 555 Dehghani, H., 450 Dehghani, M.H., 134 Dehghanian, C., 400 Demirbas, A., 127 Deng, F., 291, 292 Deng, J., 135 Deng, L., 292 De Quadros Melo, D., 127 Desalination, 149–171, 192, 199, 365, 498, 528

573 Detection, 24, 80, 135, 461, 475, 554 Devi, P., 127 Dhir, B., 127 Dialynas, E., 100, 444 Diamadopoulos, E., 129, 444 Diatomaceous earth, 543, 544, 552 Diebold, U., 391, 404 Díez-Pascual, A.M., 308 Díez-Vicente, A.L., 308 Dimensionally stable anodes (DSA), 105, 112, 319, 389–405, 414, 415 Diseases, 17, 38, 75, 78–79, 81, 124, 126, 154, 169, 184, 185, 214, 264, 266–270, 274, 283, 303, 322, 413, 416, 443, 473, 476, 478, 481, 483, 484, 486–488 Disinfection, 76, 79, 80, 151, 154, 218, 340, 341, 364, 461 Divyapriya, G., 86–113 Djaenudin, W., 487 Domínguez, J.R., 109 Dong, X., 451, 453 Douglas, M., 10 Du, Y., 554 Dubey, S., 359 Dubinin, M.M., 346, 347 Dyes, 42, 43, 50, 52, 64, 96, 124, 126, 127, 137, 138, 212–215, 218, 220, 222, 225, 228–238, 272, 282–284, 286, 287, 289, 291, 292, 294, 300, 304, 306, 360, 403, 442, 445, 456, 458, 474, 476, 484, 486, 488, 499, 540–544, 546, 548–551, 553–556, 558–560

E Ecological effects, 12, 13, 15, 16, 167 Ecotoxicology, 4, 22, 23, 25, 408 Eeshwarasinghe, D., 525 Electrochemical advanced oxidation processes, 87, 89, 100, 113 Electrochemical sensing, 474, 476 Electrodialysis, 150, 153, 155, 157, 162, 163, 179, 187, 192, 317, 444, 481, 483, 485, 487, 488, 496, 498–499, 501, 503 Eliseeva, S.V., 8 Elmolla, E.S., 133 Elrouby, M., 452 El Samrani, A.G., 444 El-Sayed, H.E.M., 127 Energy security, 160 Englisch, M., 399

574 Environmental reservoirs, 5 Environments, 4, 8, 9, 11, 12, 19, 20, 23, 24, 39, 41, 43, 45, 46, 50, 51, 60, 64, 75–78, 82, 83, 96, 98, 124, 126–128, 134, 135, 137, 138, 149, 152, 154, 156, 162, 169, 178–180, 182, 184, 186, 187, 190, 193–195, 200, 201, 212–214, 221, 224, 244, 245, 247, 260, 262, 263, 274, 282, 283, 295, 304, 317–322, 327–331, 333, 335–340, 343, 347, 355, 359, 364, 366, 379–383, 386, 387, 389, 413–415, 472, 474, 479, 482, 490, 501, 514–516, 518, 530, 531, 540 Eslami, A., 97 Esposito, E., 405

F Fakhri, A., 308 Farooq, U., 502 Farrell, J., 379 Fashola, M., 181 Feng, G., 525 Feng, N., 350, 356 Fenglian, F., 502 Fenton processes, 86, 93, 94, 98, 110–112, 228, 237 Filter media, 245–255 Filtration, 21, 24, 39, 59, 81, 124, 136, 151, 153, 154, 179, 188, 189, 191–192, 200, 201, 216, 218, 222, 251, 252, 283, 295, 317, 336, 342, 364, 366, 409, 414, 444, 477, 479, 480, 482, 485–487, 489, 496, 501, 502, 516, 517, 519, 525–531, 541 Flores-Cano, J.V., 552 Forward osmosis, 153, 155–157, 168, 171 Fowler, B.A., 488 Franca, A.S., 127 Freire-Gormaly, M., 166 Freundlich, H., 556–558 Fu, F., 444, 456 Fu, L., 343 Fu, R., 55 Fudala-Ksiazek, S., 104 Fukumura, T., 308 Fukushima, T., 269

G Gagol, M., 229 Gałuszka, A., 17 Gamoudi, S., 549 Gandhimathi, R., 97

Index García-Rodríguez, L., 161 Garcia-Segura, S., 99, 100 Garg, V.K., 130, 459 Gashi, S., 503 Gautam, R.K., 356 Ge, L., 452 Germ, M., 181 Ghanbari, F., 98 Ghasemzadeh, P., 56 Ghazouani, M., 99, 100 Ghosh (Nath), S., 336 Gillham, R.W., 50 Gogate, P.R., 229 Gonçalves, M.R., 112 Goyal, P., 137 Grace, M.A., 250 Gray, N.F., 488 Grenni, P., 447, 457 Grishkewich, N., 127 Gueneau, M.J.P., 404 Guiza, S., 350 Guo, X., 343, 459 Gupta, S.K., 134 Gupta, V., 250 Gupta, V.K., 356 Gurel, L., 502 Güzel, F., 356

H Hadiani, M.R., 448 Hakke, V.S., 212–238 Han, Z., 449 Hanafiah, M.A.K.M., 127 Hankins, N.P., 168 Haque, M., 129 Haque, N., 8 Hasan, I., 449 Hatt, B.E., 253 Hawthorne, J., 13 Hayati, B., 450 Haynes, R., 250, 251 He, F., 381 He, J., 343, 350 He, X., 453 Heavy metal pollution, 415 Heavy metal removal, 39, 52, 462, 497, 502, 526, 529, 541, 547 Heavy metals, 38–45, 50–54, 56–61, 64, 65, 75, 77, 82, 105, 178–201, 212, 213, 215, 219, 244–255, 260, 261, 263, 269–274, 282, 283, 294, 295, 302, 316–324, 328, 340, 343, 344, 347–349, 351, 355,

Index 358–360, 364–366, 370, 372, 381, 385, 405–407, 409, 410, 412, 415, 442–445, 448, 455–462, 472–476, 479, 480, 482, 485, 486, 489, 491–503, 514–519, 521–523, 525, 527–531, 540–544, 546, 552, 553, 555, 556, 559, 560 Helmers, E., 10 Hems, R., 100 Henry, J.R., 273 Henryk, K., 253 Hermawan, A.A., 250 Herrmann, J.-M., 366 Heterostructures, 286, 288, 290–292, 294–296, 298–304, 306–308 Hiu Liu, 300 Hoag, G.E., 389 Hong, M., 525 Horzum, N., 379 Hossain, K.F.B., 134 Hosseini, M., 308 Hu, X., 288, 289 Huang, C.H., 13 Huang, L., 453 Hubicki, Z., 486, 488 Human health, 4, 16–20, 24, 25, 39–43, 64, 76, 78, 82, 124, 126, 179, 183, 186, 187, 200, 212, 214, 215, 260–262, 264, 274, 282, 283, 294, 316, 319, 324, 331, 335, 338, 343, 388, 389, 398, 414, 478 Human health effects, 18, 23–25, 41, 42, 134, 322 Human intake pathways, 4, 24 Humphries, M., 8 Hurwitz, G., 99, 101 Hutchison, A., 18 Hybrid treatment technologies, 211–238 Hydrodynamic cavitation, 212, 213, 221, 224–226, 228–231, 236, 237 Hydroxyl radicals, 86–89, 91–95, 97, 99, 103–110, 113, 220, 224, 235, 284, 305, 368, 369, 378

I Iervolino, G., 295 Industrial effluents, 86, 124, 212, 216, 444, 474, 482, 499, 501, 503, 519, 531 Industrial wastewater, 86, 87, 109, 111, 112, 130, 154, 155, 216, 218, 219, 224, 227, 229, 262, 333, 358, 416, 445, 480, 482, 489 Inyang, M., 525 Ion exchange, 11, 21, 23, 39, 59, 137, 152, 153, 187, 191–193, 197, 200, 246, 282, 295, 316, 321, 340, 342, 347, 349, 363, 383,

575 444, 473, 477, 480–488, 502, 515, 521, 528, 529, 541 Ion exchange process, 137, 191, 351, 360, 491, 492, 516 Ion-selective electrode, 475, 476 Isarain-Chavez, E., 102 Itoh, H., 368, 370

J Jamil, S., 168 Jan, A., 486 Jiang, D., 383 Jiang, J., 449 Jiang, X., 451 Jin, W., 15 Jingjing, C., 318–320 Johannesson, K.H., 10 Johnson, R.L., 385 Joseph, L., 322, 324, 343, 344, 409, 412 Josih, S.M., 229 Jun, Y.S., 166 Jung, I.L., 444

K Kaczala, F., 444 Kajitvichyanukula, P., 371 Kan, C.-C., 134 Kanel, S.R., 379 Kapur, M., 353 Karami, H., 525 Karunanayake, A.G., 133 Kataria, N., 130 Katsoyiannis, I.A., 486 Kaur, R., 530 Kavamura, V.N., 405 Kavitha, E., 447 Keane, C., 8 Kennedy, J.U., 544 Khadim, H.J., 448 Khan, S.A., 308 Kharat, D.S., 127 Khimaniv, A.J., 308 Khosa, M.A., 129 Khoufi, S., 112 Kim, K.-H., 10 Kim, M., 181 Kim, S., 168 Kimbrough, D.E., 487 Knappe, A., 10 Kobielska, P.A., 134, 295 Kola, A.K., 212–238 Komkiene, J., 135 Kong, Y., 546

576 Kononova, O.N., 444 Koodynsk, D., 486, 488 Krstić, V., 316, 318, 323, 343, 396, 400, 416 Ku, Y., 444 Kumar, J., 127 Kumar, M.S., 229 Kumar, S., 308 Kumar Nair, K., 308 Kümmerer K., 10 Kuppusamy, S., 133 Kurniaw, T.A., 356 Kurniawan, T.A., 249, 343 Kurt, U., 102

L Labhane, P.K., 219 Labidi, A., 540 Labidi, J., 540–561 Laine, J., 252 Langmuir, I, 556–558 Lanthanides, 3, 7, 12, 178 Larous, S., 127 Lawrence, M.G., 21 Leads (Pb), 3, 39–44, 51, 52, 56, 57, 61–62, 78, 82, 89, 103, 104, 106–109, 112, 134–135, 138, 166, 179–186, 188–190, 196–200, 212, 215, 220, 224, 226, 244, 245, 250, 252, 253, 261–268, 270–272, 274, 282, 284, 294, 317–319, 322–329, 338, 345, 347, 360, 369, 370, 374–377, 380, 386, 389, 390, 406, 408–411, 414, 415, 443, 456, 472, 480–482, 487, 491, 493, 500, 515, 517, 542, 543, 546 Lee, J., 447 Leong, Y.K., 295 Lewis, A.S., 54 Li, B., 396, 400 Li, G., 453 Li, H., 131, 135 Li, L., 55 Li, S., 166 Li, T., 379 Li, X., 343, 497 Li, X.Q., 377 Li, Y.H., 360 Li, Z., 317, 451 Lia, X., 356 Lim, A.P., 127 Lim, H.S., 248, 253 Lim, J.Y., 343, 360, 362, 363 Limonti, C., 56 Lin, G., 339

Index Lin, H., 447 Ling, R., 343 Liu, D., 16 Liu, H., 343 Liu, J., 301 Liu, T., 54 Liu, X., 129 Liu, Y., 132 Llanos, J., 503 Lodhi, R.S., 183 Lofrano, G., 135 Long, J., 447 Lopez, A., 356 Lorenzen, L., 129 Lu, V.M., 8 Luo, J., 290, 292 Luo, M., 446

M Ma, H., 343 Ma, J., 451 Machacek, E., 8 Macpherson, J.V.J.L., 404 Madrakian, T., 360 Mahalik, D., 488 Mahalik, M.P., 488 Mahmoud, M.E., 451 Mali, J.M., 287, 289 Malik, D.S., 133 Malik, R., 446 Malkoc, E., 444 Manasiet, S., 448 Manenti, D.R., 98 Manju, G., 129 Manjula, N., 308 Manohar, D., 486 Maranon, E., 444 Marselli, B., 88 Martínez-Huitle, C.A., 97 Martino, C., 13 Marzec-Wróblewska, U., 17 Mazloom, J., 308 Mazur, L.P., 326, 327, 329, 331, 333, 336, 340, 342, 343, 381 McCarthy, D., 245 McCleskey, C.S., 13 McKay, G., 131 McLennan, S.M., 10 Melignani, E., 412 Membrane fouling, 21, 155, 157, 159, 166, 171, 342, 496, 502, 503, 527 Meniai, A.H., 127

Index Meschke, K., 527 Metal concentrations, 132, 191, 194, 274, 342, 353, 461, 492 Metalloids, 52, 152, 178–181, 183–188, 191, 193–197, 199–201, 261, 372, 476 Metals, 3, 6, 14, 15, 24, 39–41, 44, 46–48, 51–54, 57, 64, 77, 81, 88, 108, 109, 124–128, 130–135, 137, 138, 153, 168, 178, 179, 181, 182, 184–190, 192–197, 199–201, 244, 245, 247, 249, 251, 254, 260–263, 265, 269–274, 283–286, 294, 296, 299, 300, 304, 306, 316, 320–344, 347–349, 351–353, 355, 357–361, 363, 370, 372–377, 379–383, 387–399, 401, 402, 404–406, 409, 412, 414–416, 443–445, 455–462, 472–476, 478–486, 488–494, 496–503, 515–521, 525, 526, 528–530, 540, 541, 552, 554, 556 Michalak, I., 137 Migaszewski, Z.M., 17 Miretzky, P., 127 Mishra, U., 260–274 Miu, A.C., 270 Mohammadi, T., 444, 503 Mohan, D., 135 Mohan, K.G.V., 134 Mohapatra, P., 371 Mohsen-Nia, M., 444, 447 Molaei, K., 452 Molinari, R., 444 Mondal, M.X., 353 Mondal, P., 295, 486 Monteagudo, J.M., 444 Moosavi, S.G., 503 Moradi, F., 400 Moradi, M., 98, 163, 165 Morais, C.A., 8 Morais, S., 261 Moreira, F.C., 89 Mostafa, H., 112 Motsi, T., 529 Moyo, M., 127 Mulligan, C.N., 181

N Naeem, M.A., 133 Naja, G., 137 Nanocomposites, 167, 198, 200, 226, 236, 306, 308, 357, 363, 374, 445, 450, 453, 455, 458–460 Nanofiltration, 152–154, 156, 157, 167, 168, 192, 222, 231–233, 236, 282, 317, 444, 496, 526, 527

577 Nanoparticles, 13, 15, 39, 40, 44–54, 56–65, 128, 131, 168, 213, 220, 222, 228, 236, 264, 284–287, 289–292, 294–296, 299–303, 305, 306, 308, 325, 329, 358, 359, 363, 372, 375, 377, 378, 380, 382, 385–388, 391, 393, 414, 446, 459, 477, 547 Nanotechnologies, 149, 166, 167, 179, 187, 196, 201, 282, 285, 292, 317, 343, 344, 358, 364, 366, 371, 387, 414 Nano zero-valent iron, 371–389, 414 Naoto Takeno, 326, 328, 330, 332, 335, 337, 339, 341, 376 Nasir, S., 527 Nasir, Z., 308 Naumov, V., 8 Navarro, J., 8 Neetesh, K.D., 488 Nefzi, H., 540 Nĕmeček, J., 343, 385 Ngah, W.S.W., 127 Nguyen, N.C., 127 Nguyen, T.A.H., 127 Nharingo, T., 127 Nicomel, N.R., 129 Nidheesh, P.V., 86–113 Nieboer, E., 348 Noli, F., 448 Nuhoglu, F., 444 Nurzhanova, A., 530

O Obiora, S.C., 181 O’Carroll, D., 45, 375, 380 Ociński, D., 128 O’Connell, D.W., 127, 343 O’Hannesin, S., 50 Oliveira, L.S., 127 Oninla, V.O., 447 Oral, R., 13 Oreggioni, D., 294 Organic and inorganic pollutants, 40, 44, 64, 216, 230, 389, 560 Ortiz, M.J., 444 Oturan, M.A., 95

P Pagano, G., 17, 19 Pagano, M., 444 Pandiyan, R., 308 Panic, V.V., 127 Parga, J.R., 486

578 Parkin, G.F., 181 Patel, N., 262 Patel, S., 127 Patil, S.B., 444 Paulick, H., 8 Payne, E., 247 Pearson, R.G., 348 Peduzzi, P., 19 Peng, W., 295 Pérez, G., 99 Perez, J.F., 110 Pérez Barthaburu, M.E., 282–307 Pešovski, B., 318, 343, 396 Peters, C.D., 168 Phenrat, T., 381 Photocatalysis, 21, 86, 230, 231, 233, 234, 236, 237, 284–289, 295, 296, 298, 303, 343, 361, 366, 368–371, 414 Photovoltaics-reverse osmosis, 162, 164–167, 170 Phukan, A., 308 Physicochemical process, 21, 416, 485, 501 Phytoremediation, 124, 179, 187, 193–196, 273, 343, 344, 352, 405–409, 412, 414–416, 473, 500–501, 503, 516, 529–531, 541 Polat, H., 250 Pollutants, 21, 38–40, 44, 47, 49–54, 57, 58, 64, 74, 75, 78, 86, 87, 89, 92–94, 96, 98–100, 102, 103, 105–107, 110, 112, 124–126, 128, 130, 134, 136, 137, 178, 180–182, 191–193, 196, 201, 212–215, 217, 219–224, 228, 229, 232, 237, 244, 245, 254, 255, 260, 262, 265, 271–274, 282–284, 286, 316, 317, 319, 326, 328, 341, 343–345, 351, 352, 358–361, 364, 366, 368, 370, 373, 375, 383, 389, 392, 394, 401–403, 406–409, 412, 414, 415, 445, 456, 460, 462, 479, 493, 498–501, 516, 520, 521, 525, 531, 540–542, 546, 552, 554–556, 558–560 Pouyfaucon, A.B., 161 Prabha, D., 308 Pradhan, S.K., 450 Prasad, M., 461 Precipitation, 10, 21, 23, 39, 44, 58–60, 106, 107, 124, 128, 133, 136, 137, 179, 187–189, 193–195, 200, 290, 295, 316, 320, 327, 330, 334, 336, 340–343, 347, 358, 359, 365, 373–375, 377, 382, 391, 406, 407, 444, 448, 452, 459, 473, 477, 479–482, 484–490, 501, 502, 516–521, 526, 528 Pyrzynska, K., 343

Index Q Qasim, M., 155 Qi, J., 128 Qian, L., 56 Qin, N., 299, 301 Qiu, F., 292 Qu, J., 301 Quiton, K.G., 447

R Radjenovic, J., 99 Raheim, A.R,M, 486, 487 Rahman, M.M., 124–138 Rai, D., 260–274 Rajoriya, S., 228 Rangabhashiyam, S., 127 Rangel-Mendez, J.R., 132 Ranjan, D., 129 Raouf, A., 486, 487 Rasaki, S.A., 451 Raskin, I., 409, 503 Rathinam, K., 525 Raval, N.P., 127 Raza, M.A., 472 Reddy, K.R., 251, 253 Remediation, 20, 25, 39, 40, 44, 45, 47, 49–53, 56–58, 60–65, 77, 78, 81, 124, 125, 128, 133, 134, 137, 179, 186–189, 192, 194–201, 272–273, 282, 284, 285, 294, 295, 304, 306, 307, 358, 372, 373, 375, 379, 381, 385, 386, 405, 406, 415, 444, 445, 455, 457–460, 474, 477, 500, 501, 529, 531, 541 Remediation approaches, 51–53 Removal mechanisms, 44, 45, 61, 98, 106, 364, 377, 378 Ren, J., 350, 355, 357 Rengaraj, S., 371 Rengarajan, R., 10 Resende, L.V., 8 Reverse osmosis, 21, 39, 60, 99–101, 151–157, 159, 162–164, 166–168, 170, 171, 192, 216, 217, 317, 401, 444, 477, 479–481, 483–488, 496, 498, 501, 503, 526, 527 Rezania, S., 407 Richardson, D.H.S., 348 Risk assessments, 19, 20, 64, 78–79 Robalds, A., 137 Rodrigo, M.A., 111 Rogowska, J., 4 Roncati, L., 17

Index S Saad, A.H.A., 448 Sabry, R., 444 Sadegh, H., 127 Sadrzadeh, M., 444, 503 Sahmoune, M.N., 127 Salamatinia, B., 252 Salesh, R., 228 Salman, M., 356 Samani, M.R., 134 Sankararamakrishnan, N., 132 Sari, A., 132 Saroha, A.K., 127 Sawyer, C.N., 181 Saxena, S., 461 Scaria, J., 86–113 Schiewer, S., 444 Schwartz, J., 263 Scott, T.B., 384 Seawater reverse osmosis, 153, 157, 164 Sebastian, A., 446, 448 Seepana, M.M., 212–238 Seghatoleslami, M.J., 503 Semiconductors, 221, 284, 285, 295, 296, 300, 324, 361, 366–368, 371 Sen, A., 127 Shah, B.A., 353 Shahane, S, 260–274 Shahmansouri, A., 444 Shammi, M., 124–138 Shang, J., 134 Sharma, A.K., 127 Sharma, P.R., 131 Sharma, S., 343, 367 Sheng, T., 132 Sholkovitz, E.R., 10 Shukla, A., 127 Shukla, S.S., 137 Siciliano, A., 56 Sikder, M.T., 133, 137 Sillanpää, M., 127 Siti, N.A.A., 343 Sonawane, S.H., 212–238 Spasojević, M., 392 Spasojević, P., 343, 392 Srasra, E., 549 Srivastava, S., 137 Srivind, J., 308 Stegen, K.S., 8 Stormwater management, 244, 245, 248, 249 Sudhaparimala, S., 308 Sulyman, M., 127 Sun, Q., 555

579 Suresh Kumar, M., 86–113 Suzuki, Y., 132 Swami, D., 127 Swami, D.N., 134 Syam Babu, D., 86–113 Szpyrkowicz, L., 102

T Taha, M., 165 Tahari, N., 540 Takasu, Y., 400 Talei, A., 244–255 Tan, Q., 8 Tang, J., 13 Tang, X., 295 Tassel, F., 444 Taufik, A., 228 Temkin, M.I., 556–558 Tessele, F., 444 Thaçi, B., 503 Thiruvenkatachari, R., 368, 370, 371 Thomas, P.J., 13, 14 Thomsen, H.S., 17 Tin compounds, 304, 306, 307 Tin (IV) oxide, 288, 291, 299, 304, 306 Tin (IV) sulfide, 286, 288, 290–292, 295, 300–302, 304 Tirry, N., 344 Toghraie, D., 134 Townley, H.E., 8 Toxic effects, 82, 195, 260, 270, 282, 325, 331, 443 Toxicities, 4, 14, 15, 18, 19, 25, 38–41, 43, 57, 61, 63, 64, 82, 99, 112, 124, 133, 182, 183, 186, 221, 260, 261, 266, 268, 269, 274, 282, 283, 285, 304, 305, 486, 488 Toxicity, 321–324, 331, 334, 371, 375, 387, 388, 405 Toxicity and removal of heavy metals, v, vi, 2–5, 14, 15, 18, 19, 37–65, 82, 99, 112, 124, 133, 182, 183, 186, 188, 190, 197, 221, 243–255, 260, 261, 266–269, 274, 282, 283, 285, 304, 305, 318, 320–340, 343, 344, 351, 355, 360, 364, 366, 368, 370–389, 405, 409, 410, 412, 415, 455, 462, 485–492, 494, 500, 502, 503, 513–531, 546 Tratnyek, P.G., 385 Trumić, B., 363 Tsantaki, E., 96 Tu, J.R., 296, 301

580 Tukaram Bai, M., 446 Tuzen, M., 132

U Ultrafiltration, 152–154, 156, 159, 163, 166, 187, 192, 317, 496–497, 503, 526, 527, 548 Umar, A., 287–289 Uzma, I., 181

V Vafaeifard, M., 451 Vaishnavi, M., 308 Vajedi, F., 450 Vakili, M., 127 Vandenbossche, M., 127 Van Hege, K., 100, 101 Vanhoudt, N., 460 Vargas, A.M.M., 346 Vengris, T., 356 Venkateswarlu, P., 446 Venu, D., 105, 106 Vieira, B.R., 130 Vijaya Bhaskar Reddy, A., 49 Vohla, C., 127 Volesky, B., 137 Vooradi, R., 212–238

W Wang, C., 16, 446, 448 Wang, C.B., 373 Wang, C.T., 97, 98 Wang, G., 444 Wang, H., 128, 135, 344, 446, 448 Wang, H.J., 502 Wang, J., 248, 255 Wang, L., 502 Wang, L.K., 502 Wang, M., 502 Wang, Q., 444, 456, 502 Wang, S., 128, 130, 135, 181 Wang, Y., 130 Wang, Z., 168 Warsinger, D.M., 156 Wastes, 4, 8, 9, 20–22, 24, 39, 40, 42, 53, 56, 74, 75, 77, 79–83, 102, 103, 108, 111, 124–128, 131, 133, 134, 154, 159, 164, 167, 171, 178, 180, 190, 198, 199, 201, 214, 216, 218, 219, 222, 253, 263, 267,

Index 272, 283, 317, 318, 320, 324, 328, 338–340, 342, 343, 345, 350–352, 355, 356, 387, 409, 414, 416, 446, 448, 451, 455–457, 474, 476, 478, 480, 481, 483, 486, 488, 489, 493, 499, 500, 503, 517, 518, 526, 554 Wastewaters, 4, 8–11, 20, 21, 24, 25, 39–41, 43, 44, 47, 50–53, 56–61, 65, 74, 75, 79, 90, 94–103, 106–113, 126, 130, 133, 149, 150, 153–157, 163, 164, 167, 168, 171, 180, 188–192, 194–197, 199–201, 212–220, 223–227, 229, 233, 236, 237, 262, 313–416, 442, 455–458, 473, 474, 478, 480–502, 514, 517, 518, 521, 527–531, 540–542, 549, 554, 555, 560 Wastewater treatments, 9, 11, 40, 74, 86, 87, 98, 99, 101, 102, 111, 112, 156, 164, 188, 192, 193, 212–219, 221, 222, 224, 227, 229, 316, 317, 341, 342, 352, 355, 358, 360, 361, 364–366, 371, 389, 392, 412–414, 416, 443, 473, 489, 498, 499, 525 Water-energy nexus, 150, 171 Waters, A., 444 Water security, 148–150, 159, 162, 167, 169–171 Water treatments, 24, 44, 78, 81, 82, 128, 149, 150, 152, 154, 158, 161, 166, 167, 191, 193, 224, 249, 251, 316, 482, 499, 501, 541 Wickramasinghe, S., 486 Woo, M.Y.C., 389 Wu, B., 165 Wu, W., 451 Wu, X., 488 Wu, Y., 392, 394, 403 Wu, Z., 287–289

X Xavier, P., 353 Xia, Q., 13 Xia, Y., 110 Xiao, S., 325 Xia Zhang, 302 Xie, F., 8 Xu, F., 343 Xu, H., 452 Xu, J., 295 Xu, M., 131 Xu, X., 343 Xue, G., 343 Xue, P.Y., 412

Index Y Yahya, N., 343, 369 Yamin, N., 472–501 Yanful, E.K., 329 Yang, Y., 168 Yanyan, L., 353 Yao, Z., 250 Yeddou, A.R., 127 Yeheyis, M.B., 250 Yin, K., 295 Yin, N., 453 Yoi Can Zhang, 287 Yong, S.K., 127 Yoon, I.H., 343, 371 Youssef, A.M., 450 Yu, S., 344 Yu, X., 396, 398, 449 Yuan, P., 251 Yuan, X., 290, 292 Yurekli, Y., 343 Yurlova, L., 444

Z Zaman, M.I., 359 Zeraatkar, A.K., 348, 349

581 Zerovalent iron, 40, 44–54, 56–65 Zhang, A., 178 Zhang, F., 450 Zhang, F.S., 368, 370, 377 Zhang, J., 130 Zhang, W.X., 373 Zhang, Xi., 299, 301 Zhang, Y., 13, 401 Zhang, Y.C., 288, 289, 291, 299, 301 Zhang, Y.-F., 344 Zhang, Z., 292, 327 Zhao, D., 381 Zhao, F., 8 Zhao, Y., 127, 370, 450 Zheng, C., 168 Zhou, N., 135 Zhou, Y., 127 Zhou, Z.Q., 127 Zhu, L., 460 Zhu, W., 17 Zhu, W.F., 17, 19 Zhu, X., 17 Zhuang, M., 17 Zou, J., 97 Zu, 288