Advanced Research in Nanosciences for Water Technology [1st ed.] 978-3-030-02380-5, 978-3-030-02381-2

Table of contents :
Front Matter ....Pages i-xviii
Economic Assessment of Nanomaterials in Bio-Electrical Water Treatment (Michael Siegert, Jayesh M. Sonawane, Chizoba I. Ezugwu, Ram Prasad)....Pages 1-23
Bioconjugated Quantum Dots in Rapid Detection of Water Microbial Load: An Emerging Technology (Indu Pal Kaur, Joga Singh, Jatinder V. Yakhmi, Gurpal Singh, Corinne Dejous, Alka Bhatia et al.)....Pages 25-38
Water Pollution Remediation Techniques with Special Focus on Adsorption (Sujata Mandal, C. Muralidharan, Asit Baran Mandal)....Pages 39-68
Effect of Nano-TiO2 Particles on Mechanical Properties of Hydrothermal Aged Glass Fiber Reinforced Polymer Composites (Ramesh Kumar Nayak)....Pages 69-93
Nanotechnology: An Innovative Way for Wastewater Treatment and Purification (Muhammad Rafique, Muhammad Bilal Tahir, Iqra Sadaf)....Pages 95-131
Immobilized Nanocatalysts for Degradation of Industrial Wastewater (Jayaseelan Arun, Marudai Joselyn Monica, Vargees Felix, Kannappan Panchamoorthy Gopinath)....Pages 133-145
New Technologies to Remove Halides from Water: An Overview (José Rivera-Utrilla, Manuel Sánchez-Polo, Ana M. S. Polo, Jesús J. López-Peñalver, María V. López-Ramón)....Pages 147-180
Nanotechnology Explored for Water Purification (A. Laha, D. Biswas, S. Basak)....Pages 181-193
Nanomaterials in the Development of Biosensor and Application in the Determination of Pollutants in Water (Germán A. Messina, Matías Regiart, Sirley V. Pereira, Franco A. Bertolino, Pedro R. Aranda, Julio Raba et al.)....Pages 195-215
Clay-Based Nanocomposites: Potential Materials for Water Treatment Applications (Faraan Fareed, M. Ibrar, Yaseen Ayub, Rabia Nazir, Lubna Tahir)....Pages 217-248
Application of Nano-Photocatalysts for Degradation and Disinfection of Wastewater (Jayaseelan Arun, Vargees Felix, Marudai Joselyn Monica, Kannappan Panchamoorthy Gopinath)....Pages 249-261
Degradation of Emerging Contaminants Using Fe-Doped TiO2 Under UV and Visible Radiation (Irwing M. Ramírez-Sánchez, Oscar D. Máynez-Navarro, Erick R. Bandala)....Pages 263-285
Oxide Nanomaterials for Efficient Water Treatment (Alagappan Subramaniyan)....Pages 287-297
Nanotechnology for Oil-Water Separation (Prakash M. Gore, Anukrishna Purushothaman, Minoo Naebe, Xungai Wang, Balasubramanian Kandasubramanian)....Pages 299-339
Nanotechnology for Wastewater Treatment and Bioenergy Generation in Microbial Fuel Cells (M. J. Salar-García, V. M. Ortiz-Martínez)....Pages 341-362
Nanocomposite Materials Based on TiO2/Clay for Wastewater Treatment (Soulaima Chkirida, Nadia Zari, Abou El Kacem Qaiss, Rachid Bouhfid)....Pages 363-380
Nanotechnology: The Technology for Efficient, Economic, and Ecological Treatment of Contaminated Water (S. Vijayakumar, M. Priya)....Pages 381-405
Silver Nanoparticles as a Biocide for Water Treatment Applications (Renat R. Khaydarov, Rashid A. Khaydarov, Olga Gapurova, Ilnur Garipov, M. Lutfi Firdaus)....Pages 407-419
Micro- and Nano-Hollow Spheres in Heavy Metal Removals from Water (Jayeeta Chattopadhyay)....Pages 421-441
Back Matter ....Pages 443-457

Citation preview

Nanotechnology in the Life Sciences

Ram Prasad Thirugnanasambandham Karchiyappan Editors

Advanced Research in Nanosciences for Water Technology

Nanotechnology in the Life Sciences Series Editor Ram Prasad School of Environmental Science and Engineering, Sun Yat-sen University, Guangzhou, China Amity Institute of Microbial Technology, Amity University Uttar Pradesh, Noida, UP, India

Nano and biotechnology are two of the 21st century’s most promising technologies. Nanotechnology is demarcated as the design, development, and application of materials and devices whose least functional make up is on a nanometer scale (1 to 100 nm). Meanwhile, biotechnology deals with metabolic and other physiological developments of biological subjects including microorganisms. These microbial processes have opened up new opportunities to explore novel applications, for example, the biosynthesis of metal nanomaterials, with the implication that these two technologies (i.e., thus nanobiotechnology) can play a vital role in developing and executing many valuable tools in the study of life. Nanotechnology is very diverse, ranging from extensions of conventional device physics to completely new approaches based upon molecular self-assembly, from developing new materials with dimensions on the nanoscale, to investigating whether we can directly control matters on/in the atomic scale level. This idea entails its application to diverse fields of science such as plant biology, organic chemistry, agriculture, the food industry, and more. Nanobiotechnology offers a wide range of uses in medicine, agriculture, and the environment. Many diseases that do not have cures today may be cured by nanotechnology in the future. Use of nanotechnology in medical therapeutics needs adequate evaluation of its risk and safety factors. Scientists who are against the use of nanotechnology also agree that advancement in nanotechnology should continue because this field promises great benefits, but testing should be carried out to ensure its safety in people. It is possible that nanomedicine in the future will play a crucial role in the treatment of human and plant diseases, and also in the enhancement of normal human physiology and plant systems, respectively. If everything proceeds as expected, nanobiotechnology will, one day, become an inevitable part of our everyday life and will help save many lives. More information about this series at http://www.springer.com/series/15921

Ram Prasad • Thirugnanasambandham Karchiyappan Editors

Advanced Research in Nanosciences for Water Technology

Editors Ram Prasad School of Environmental Science and Engineering Sun Yat-sen University Guangzhou, China

Thirugnanasambandham Karchiyappan State University of Maringa Paraná, Paraná, Brazil

Amity Institute of Microbial Technology Amity University Uttar Pradesh Noida, UP, India

ISSN 2523-8027 ISSN 2523-8035 (electronic) Nanotechnology in the Life Sciences ISBN 978-3-030-02380-5 ISBN 978-3-030-02381-2 (eBook) https://doi.org/10.1007/978-3-030-02381-2 Library of Congress Control Number: 2018963720 © Springer Nature Switzerland AG 2019 This work is subject to copyright. All rights are reserved by the Publisher, whether the whole or part of the material is concerned, specifically the rights of translation, reprinting, reuse of illustrations, recitation, broadcasting, reproduction on microfilms or in any other physical way, and transmission or information storage and retrieval, electronic adaptation, computer software, or by similar or dissimilar methodology now known or hereafter developed. The use of general descriptive names, registered names, trademarks, service marks, etc. in this publication does not imply, even in the absence of a specific statement, that such names are exempt from the relevant protective laws and regulations and therefore free for general use. The publisher, the authors, and the editors are safe to assume that the advice and information in this book are believed to be true and accurate at the date of publication. Neither the publisher nor the authors or the editors give a warranty, express or implied, with respect to the material contained herein or for any errors or omissions that may have been made. The publisher remains neutral with regard to jurisdictional claims in published maps and institutional affiliations. This Springer imprint is published by the registered company Springer Nature Switzerland AG The registered company address is: Gewerbestrasse 11, 6330 Cham, Switzerland

Preface

Establishment of clean, safe water is one of the major challenges faced around the world. The continued urbanization of human populations, the increasing manipulation of natural resources, and the resulting pollution are placing a remarkable burden on water resources. Increasing demands for food, energy, and natural resources are expected to continue to accelerate into the near future in response to the demands of changing human populations. In addition, the complexity of human activities is leading to a diversity of new chemical contaminants in the environment that present a major concern for water managements. This will create increased pressure on both water quantity and quality, making it increasingly difficult to provide a sustainable supply of water for human welfare and activities. Due to their unique physico-chemical properties, nanotechnologies are extensively used in antibacterials (medical products), membrane filters, electronics, photo catalysts, and biosensors. Nanoparticles can have distinctly different properties from their bulk counterparts, creating the opportunity for new materials with a diversity of applications. Recent developments related to water treatment include the potential use of quantum dots, nanocomposite, nanospheres, and nanowires for the removal of a diversity of chemical pollutants. By exploiting the assets and structure of these new materials, such as increased surface area, high reactivity, and photocatalytic action, it will be possible to create technologies that can be very efficient at removing and degrading environmental pollutants. Understanding and using these unique properties should lead to innovative, cost-effective applications for addressing the complexities of the emerging needs for wastewater treatment. The book Advanced Research in Nanosciences for Water Technology comprises 19 chapters written by the experts in this field, highlighting latest research on nanoscience such as applications of nanotechnology for wastewater treatment, bioenergy generation in microbial fuel cells, water purification, heavy metal removal, and oil–water treatment. With a fortune of information on different aspects of nanotechnology, this extensive volume is a valuable resource for researchers, academicians, policy makers, and students in various fields such as microbiology, nanotechnology, and environmental science. v

vi

Preface

We are highly delighted and thankful to all our contributing authors for their support and active cooperation in their writing of these valuable chapters. Ram Prasad extends sincere thanks to all our colleagues who helped us in the preparation of the generous volume. We thank Springer officials, especially Eric Stannard, Anthony Dunlap, Rahul Sharma, and Gomathi Mohanarangan, for their generous support and efforts to accomplish this volume. Noida, Uttar Pradesh, India Paraná, Paraná, Brazil

Ram Prasad Thirugnanasambandham Karchiyappan

Contents

1

2

3

4

5

Economic Assessment of Nanomaterials in Bio-Electrical Water Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Michael Siegert, Jayesh M. Sonawane, Chizoba I. Ezugwu, and Ram Prasad Bioconjugated Quantum Dots in Rapid Detection of Water Microbial Load: An Emerging Technology . . . . . . . . . . . Indu Pal Kaur, Joga Singh, Jatinder V. Yakhmi, Gurpal Singh, Corinne Dejous, Alka Bhatia, Ashish Sattee, and Udit Soni Water Pollution Remediation Techniques with Special Focus on Adsorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sujata Mandal, C. Muralidharan, and Asit Baran Mandal Effect of Nano-TiO2 Particles on Mechanical Properties of Hydrothermal Aged Glass Fiber Reinforced Polymer Composites . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Ramesh Kumar Nayak Nanotechnology: An Innovative Way for Wastewater Treatment and Purification . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Muhammad Rafique, Muhammad Bilal Tahir, and Iqra Sadaf

1

25

39

69

95

6

Immobilized Nanocatalysts for Degradation of Industrial Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 133 Jayaseelan Arun, Marudai Joselyn Monica, Vargees Felix, and Kannappan Panchamoorthy Gopinath

7

New Technologies to Remove Halides from Water: An Overview . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 147 José Rivera-Utrilla, Manuel Sánchez-Polo, Ana M. S. Polo, Jesús J. López-Peñalver, and María V. López-Ramón

vii

viii

Contents

8

Nanotechnology Explored for Water Purification . . . . . . . . . . . . . . 181 A. Laha, D. Biswas, and S. Basak

9

Nanomaterials in the Development of Biosensor and Application in the Determination of Pollutants in Water . . . . . . . . . . . . . . . . . . 195 Germán A. Messina, Matías Regiart, Sirley V. Pereira, Franco A. Bertolino, Pedro R. Aranda, Julio Raba, and Martín A. Fernández-Baldo

10

Clay-Based Nanocomposites: Potential Materials for Water Treatment Applications . . . . . . . . . . . . . . . . . . . . . . . . . 217 Faraan Fareed, M. Ibrar, Yaseen Ayub, Rabia Nazir, and Lubna Tahir

11

Application of Nano-Photocatalysts for Degradation and Disinfection of Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . 249 Jayaseelan Arun, Vargees Felix, Marudai Joselyn Monica, and Kannappan Panchamoorthy Gopinath

12

Degradation of Emerging Contaminants Using Fe-Doped TiO2 Under UV and Visible Radiation . . . . . . . . . . . . . . . . . . . . . . 263 Irwing M. Ramírez-Sánchez, Oscar D. Máynez-Navarro, and Erick R. Bandala

13

Oxide Nanomaterials for Efficient Water Treatment . . . . . . . . . . . . 287 Alagappan Subramaniyan

14

Nanotechnology for Oil-Water Separation . . . . . . . . . . . . . . . . . . . 299 Prakash M. Gore, Anukrishna Purushothaman, Minoo Naebe, Xungai Wang, and Balasubramanian Kandasubramanian

15

Nanotechnology for Wastewater Treatment and Bioenergy Generation in Microbial Fuel Cells . . . . . . . . . . . . . . . . . . . . . . . . . 341 M. J. Salar-García and V. M. Ortiz-Martínez

16

Nanocomposite Materials Based on TiO2/Clay for Wastewater Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 363 Soulaima Chkirida, Nadia Zari, Abou El Kacem Qaiss, and Rachid Bouhfid

17

Nanotechnology: The Technology for Efficient, Economic, and Ecological Treatment of Contaminated Water . . . . . . . . . . . . . 381 S. Vijayakumar and M. Priya

18

Silver Nanoparticles as a Biocide for Water Treatment Applications . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 407 Renat R. Khaydarov, Rashid A. Khaydarov, Olga Gapurova, Ilnur Garipov, and M. Lutfi Firdaus

Contents

19

ix

Micro- and Nano-Hollow Spheres in Heavy Metal Removals from Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 421 Jayeeta Chattopadhyay

Index . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 443

Contributors

Pedro R. Aranda Instituto de Química de San Luis (INQUISAL)—Consejo Nacional de Investigaciones Científicas y Técnicas (CONICET), Universidad Nacional de San Luis (UNSL), San Luis, Argentina Jayaseelan Arun Department of Chemical Engineering, SSN College of Engineering, Kalavakkam, Tamil Nadu, India Yaseen Ayub Government Islamia College Civil Lines, Lahore, Pakistan Erick R. Bandala Desert Research Institute (DRI), Las Vegas, NV, USA S. Basak Indian Council of Agricultural Research, Central Institute for Research on Cotton Technology (CIRCOT), Mumbai, Maharashtra, India Franco A. Bertolino Instituto de Química de San Luis (INQUISAL)—Consejo Nacional de Investigaciones Científicas y Técnicas (CONICET), Universidad Nacional de San Luis (UNSL), San Luis, Argentina Alka Bhatia Department of Experimental Medicine and Biotechnology, PGIMER, Chandigarh, India D. Biswas Indian Jute Industries Research Association, Kolkata, West Bengal, India Rachid Bouhfid Laboratory of Polymer Processing, Moroccan Foundation for Advanced Science, Innovation and Research (MAScIR), Institute of Nanomaterials and Nanotechnology (NANOTECH), Rabat, Morocco Jayeeta Chattopadhyay Department of Chemistry, Amity University Jharkhand, Ranchi, Jharkhand, India Soulaima Chkirida Laboratory of Polymer Processing, Moroccan Foundation for Advanced Science, Innovation and Research (MAScIR), Institute of Nanomaterials and Nanotechnology (NANOTECH), Rabat, Morocco

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Contributors

Corinne Dejous Universite de Bordeaux, IMS, ENSEIRB, CNRS UMR 5218, Talence, France Abou El Kacem Qaiss Laboratory of Polymer Processing, Moroccan Foundation for Advanced Science, Innovation and Research (MAScIR), Institute of Nanomaterials and Nanotechnology (NANOTECH), Rabat, Morocco Chizoba I. Ezugwu School of Environmental Science and Engineering, Sun Yat-sen University, Guangzhou, China Faraan Fareed Government College of Science, Lahore, Pakistan Vargees Felix Department of Chemical Engineering, SSN College of Engineering, Kalavakkam, Tamil Nadu, India Martín A. Fernández-Baldo Instituto de Química de San Luis (INQUISAL) – Consejo Nacional de Investigaciones Científicas y Técnicas (CONICET), Universidad Nacional de San Luis (UNSL), San Luis, Argentina Olga Gapurova Institute of Nuclear Physics, Uzbekistan Academy of Sciences, Tashkent, Uzbekistan Ilnur Garipov Institute of Nuclear Physics, Uzbekistan Academy of Sciences, Tashkent, Uzbekistan Kannappan Panchamoorthy Gopinath Department of Chemical Engineering, SSN College of Engineering, Kalavakkam, Tamil Nadu, India Prakash M. Gore Nano Surface Texturing Lab, Department of Metallurgical & Materials Engineering, DIAT (DU), Ministry of Defence, Pune, Girinagar, India Institute for Frontier Materials, Deakin University, Geelong, VIC, Australia M. Ibrar Lahore Garrison University, Lahore, Pakistan Balasubramanian Kandasubramanian Nano Surface Texturing Lab, Department of Metallurgical & Materials Engineering, DIAT (DU), Ministry of Defence, Pune, Girinagar, India Indu Pal Kaur University Institute of Pharmaceutical Sciences, Panjab University, Chandigarh, India Rashid A. Khaydarov Institute of Nuclear Physics, Uzbekistan Academy of Sciences, Tashkent, Uzbekistan Renat R. Khaydarov Institute of Nuclear Physics, Uzbekistan Academy of Sciences, Tashkent, Uzbekistan A. Laha Reliance Industry Ltd., Navi Mumbai, Maharashtra, India Jesús J. Lopez-Peñalver Faculty of Science, Department of Inorganic Chemistry, University of Granada, Granada, Spain

Contributors

xiii

María V. López-Ramón Faculty of Experimental Science, Department of Inorganic and Organic Chemistry, University of Jaén, Jaén, Spain M. Lutfi Firdaus Graduate School of Science Education, University of Bengkulu, Bengkulu, Indonesia Asit Baran Mandal CSIR-Central Leather Research Institute, Chennai, Tamil Nadu, India Sujata Mandal CLRI-Centre for Analysis, Testing, Evaluation and Reporting Services (CATERS), CSIR-Central Leather Research Institute, Chennai, Tamil Nadu, India Oscar D. Máynez-Navarro Universidad de las Américas Puebla (UDLAP), Ex-Hacienda Santa Catarina Mártir, Cholula, Puebla, Mexico Germán A. Messina Instituto de Química de San Luis (INQUISAL) – Consejo Nacional de Investigaciones Científicas y Técnicas (CONICET), Universidad Nacional de San Luis (UNSL), San Luis, Argentina Marudai Joselyn Monica Department of Chemical Engineering, SSN College of Engineering, Kalavakkam, Tamil Nadu, India C. Muralidharan Leather Processing Department, CSIR-Central Leather Research Institute, Chennai, Tamil Nadu, India Minoo Naebe Institute for Frontier Materials, Deakin University, Geelong, VIC, Australia Ramesh Kumar Nayak School of Mechanical Engineering, KIIT, Deemed to be University, Bhubaneswar, Odisha, India Rabia Nazir Pakistan Council of Scientific and Industrial Research Labs Complex, Applied Chemistry Research Centre, Lahore, Pakistan V. M. Ortiz-Martínez Department of Chemical and Environmental Engineering, Campus Muralla del Mar, Technical University of Cartagena, Cartagena, Spain Department of Chemical Engineering, Campus Espinardo, University of Murcia, Murcia, Spain Sirley V. Pereira Instituto de Química de San Luis (INQUISAL) – Consejo Nacional de Investigaciones Científicas y Técnicas (CONICET), Universidad Nacional de San Luis (UNSL), San Luis, Argentina Ana M. S. Polo Faculty of Science, Department of Inorganic Chemistry, University of Granada, Granada, Spain Ram Prasad School of Environmental Science and Engineering, Sun Yat-sen University, Guangzhou, China

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Contributors

M. Priya Department of Science and Humanities, Sri Ramakrishna Institute of Technology, Coimbatore, Tamil Nadu, India Anukrishna Purushothaman Centre for Biopolymer Science and Technology, Central Institute of Plastics Engineering and Technology, Eloor, Udyogmandal, Kochi, India Julio Raba Instituto de Química de San Luis (INQUISAL) – Consejo Nacional de Investigaciones Científicas y Técnicas (CONICET), Universidad Nacional de San Luis (UNSL), San Luis, Argentina Muhammad Rafique Department of Physics, University of Gujrat, Gujrat, Pakistan Irwing M. Ramírez-Sánchez Department of Civil, Architectural and Environmental Engineering, The University of Texas at Austin, Austin, TX, USA Matías Regiart Instituto de Química de San Luis (INQUISAL) – Consejo Nacional de Investigaciones Científicas y Técnicas (CONICET), Universidad Nacional de San Luis (UNSL), San Luis, Argentina José Rivera-Utrilla Faculty of Science, Department of Inorganic Chemistry, University of Granada, Granada, Spain Iqra Sadaf Department of Physics, University of Gujrat, Gujrat, Pakistan M. J. Salar-García Department of Chemical and Environmental Engineering, Campus Muralla del Mar, Technical University of Cartagena, Cartagena, Spain Department of Chemical Engineering, Campus Espinardo, University of Murcia, Murcia, Spain Manuel Sánchez-Polo Faculty of Science, Department of Inorganic Chemistry, University of Granada, Granada, Spain Ashish Sattee Faculty of Applied Medical Sciences, Department of Pharmacognosy and Phytochemistry, LPU, Phagwara, Punjab, India Michael Siegert Independent Investigator, Chicago, IL, USA Gurpal Singh University Institute of Pharmaceutical Sciences, Panjab University, Chandigarh, India Joga Singh University Institute of Pharmaceutical Sciences, Panjab University, Chandigarh, India Jayesh M. Sonawane IITB-Monash Research Academy, Indian Institute of Technology Bombay, Mumbai, India Udit Soni Department of Biotechnology, Teri School of Advanced Studies, New Delhi, India

Contributors

xv

Alagappan Subramaniyan Department of Physics, Thiagarajar College of Engineering, Madurai, Tamil Nadu, India Lubna Tahir Government Islamia College Civil Lines, Lahore, Pakistan Muhammad Bilal Tahir Department of Physics, University of Gujrat, Gujrat, Pakistan S. Vijayakumar Department of Science and Humanities, Sri Ramakrishna Institute of Technology, Coimbatore, Tamil Nadu, India Xungai Wang Institute for Frontier Materials, Deakin University, Geelong, VIC, Australia Jatinder V. Yakhmi Formerly at Bhabha Atomic Research Centre, Mumbai, India Nadia Zari Laboratory of Polymer Processing, Moroccan Foundation for Advanced Science, Innovation and Research (MAScIR), Institute of Nanomaterials and Nanotechnology (NANOTECH), Rabat, Morocco

About the Editors

Ram Prasad PhD has been associated with the Amity Institute of Microbial Technology, Amity University, Uttar Pradesh, India, since 2005. His research interests include microbiology, plant-microbe interactions, sustainable agriculture, and microbial nanobiotechnology. Dr. Prasad has more than a hundred publications to his credit, including research papers, review articles, book chapters, five patents issued or pending and several authored or edited books. Dr. Prasad has 12 years of teaching experience, and has been awarded the Young Scientist Award (2007) and Prof. J. S. Datta Munshi Gold Medal (2009) by the International Society for Ecological Communications; FSAB Fellowship (2010) by the Society for Applied Biotechnology; the American Cancer Society UICC International Fellowship for Beginning Investigators, USA (2014); Outstanding Scientist Award (2015) in the field of Microbiology by Venus International Foundation; BRICPL Science Investigator Award (ICAABT-2017); and Research Excellence Award (2018). He has been serving as an editorial board member on Frontiers in Microbiology, Frontiers in Nutrition, and Academia Journal of Biotechnology and is the series editor of Nanotechnology in the Life Sciences, Springer Nature, USA. Previously, Dr. Prasad served as visiting assistant professor at the Whiting School of Engineering, Department of Mechanical Engineering, Johns Hopkins University, USA, and presently serves as research associate professor at the School of Environmental Science and Engineering, Sun Yat-sen University, Guangzhou, China, and Amity Institute of Microbial Technology, Amity University, Noida, Uttar Pradesh, India Thirugnanasambandham Karchiyappan PhD is pursuing a career that contributes valuable teaching and research in the area of Chemistry and Engineering. He completed his Bachelor’s and Master’s degree in Chemistry from Bharathiar University, Tamil Nadu, India, and completed his PhD in Chemistry (Anna University, Tamil Nadu, India) by studying industrial wastewater treatment. He has focused on the research area of wastewater and drinking water purification, biogas generation,

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About the Editors

CO2 capture, biodegradable polymers, and fermentation technology. He has published 25 research articles in international peer-reviewed journals and attended more than 10 seminars or conferences. Presently, he is working as a postdoctoral research fellow at the UEM, Brazil. State University of Maringá, Paraná, Brazil.

Chapter 1

Economic Assessment of Nanomaterials in Bio-Electrical Water Treatment Michael Siegert, Jayesh M. Sonawane, Chizoba I. Ezugwu, and Ram Prasad

Contents 1.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2 Economic Assessment of BES Materials . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.3 Nanomaterials Used BES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.3.1 Carbon-Based Nanomaterials . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.3.2 Metal and Metal Composite Materials . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.3.3 Minerals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.3.4 Anodes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.3.5 Cathodes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.4 Membranes and Electrolyte . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.5 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

1.1

1 4 8 8 11 11 12 13 14 16 17

Introduction

Three percent of the electricity consumption in fully industrialized countries is accounted for by wastewater treatment (McCarty et al. 2011; Rothausen and Conway 2011). In China and other developing countries, this value is roughly 1.5%

M. Siegert (*) Independent Investigator, Chicago, IL, USA e-mail: [email protected] J. M. Sonawane IITB-Monash Research Academy, Indian Institute of Technology Bombay, Mumbai, India C. I. Ezugwu · R. Prasad School of Environmental Science and Engineering, Sun Yat-sen University, Guangzhou, China © Springer Nature Switzerland AG 2019 R. Prasad, K. Thirugnanasambandham (eds.), Advanced Research in Nanosciences for Water Technology, Nanotechnology in the Life Sciences, https://doi.org/10.1007/978-3-030-02381-2_1

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M. Siegert et al.

WWTP 200 m3/d 20 g/L COD 5,800 kWh/d With electrolysis (normal plant) all in kWh/d (1,100) Grit filter Pump

Surplus 3,600 kWh/d

440 (2,200)

Clarifier 1

Aeration

(200) Clarifier 2

UV

Discharge

Sludge removal Sludge removal 240 (1,200) Anaerobic digester

Total (100%) −2,200 kWh/d (savings) +3,600 kWh/d (surplus)

20 (100) Drying 200 (1,000)

+1,400 kWh/d $90/d

Fig. 1.1 Possible cost savings and power surplus from microbial electrolysis for an ideal industrial wastewater treatment plant (WWTP) scenario, using 5.8 MWh d−1 for treating 200 m3 of high strength wastewater. Savings in bold font are based on 80% COD removal

(China: 100 TWh; Wang 2013) for the treatment of 6000 TWh (Central Intelligence Agency 2017) total production. Much of the wastewater produced in developing countries is still discharged untreated (Liang et al. 2018). At the same time, the reuse of wastewater resources can turn treatment facilities into net energy producers (Shizas and Bagley 2004). Several technologies are able to accomplish this with anaerobic membrane bed reactors (AnMBR) being the most mature (van Lier et al. 2016). Another promising future technology is bio-electrical systems (BES). Depending on the quality of the wastewater, all of these technologies routinely remove more than 80% chemical oxygen demand (COD) and recover the respective energy amount. COD reduction and energy recovery result in substantial savings during the construction and operation of wastewater treatment plants (WWTP; Fig. 1.1). Throughout this article, we use market prices of $0.06 kWh
1 and $3 MMBtu
1 as baseline for our technoeconomic assessment. Material costs were extracted from the respective publications, or inquired from manufacturers and retailers either through quotations or Internet offers. Unlike in most previous cost assessments of BES technologies, we used a scaling coefficient of 0.6 to account for capital expenditure discounts in large-scale applications but admit that this is somewhat arbitrary (Tribe and Alpine 1986). We also include savings of 80% incurred from reduced COD loading after BES treatment. Two different types of BES are possible (Fig. 1.2), microbial fuel cells (MFC) and microbial electrolysis cells (MEC). MFCs directly harvest electrical power using wastewater organics as fuel, with oxygen as terminal electron acceptor. The

1 Economic Assessment of Nanomaterials in Bio-Electrical Water Treatment

3

Fig. 1.2 Schematic of a bio-electrical system (BES) in the form of a so-called H-cell. The BES can be operated in electrolysis mode with a power supply, or in fuel cell mode with a resistor connecting both electrodes. The electrodes may be of the same material and shape but they are usually different as they carry out different reactions, shown on the bottom. Each electrode may be poised to an electrochemical potential relative to a reference electrode. The reference electrode is optional and may be used for monitoring purposes only, e.g., in fuel cell mode. A membrane may separate both chambers of the H-cell. Ox, oxidant, Red, reductant

technology has been known for more than a century and was heavily investigated during the last two decades (Potter 1911; He et al. 2005; Logan and Rabaey 2012). Despite these efforts, power densities remain low compared with conventional hydrogen fuel cells. This situation might be improved using better materials for electrodes and membranes. MECs experienced two stages of discovery, first cathodic nitrate reduction was reported by Sakakibara and Kuroda (1993). Fifteen years later, organic matter degradation coupled to hydrogen production was discovered by Rozendal and Buisman (2010). The difference of MECs to MFCs is that a small voltage is applied, which improves reaction kinetics and makes it easier to control the treatment process. As a result, hydrogen gas is produced which is then converted into methane (Kuroda and Watanabe 1995). The process is naturally exergonic and called acetoclastic methanogenesis, assuming that acetate is the substrate: þ Rxn 1 ! H3 C
COO
þ Hþ þ H2 O ! CH4 þ HCO
3 þH ; 0

ΔG ¼
31 kJ mol
1 ðCH4 Þ The application of a small electrical potential to the reaction, for example, 0.7 V, drastically improves the energy balance for acetoclastic microbes by additional
540 kJ mol−1 (CH4), and hence also microbial reaction kinetics. Ultimately, this

4

M. Siegert et al.

leads to shorter hydraulic retention times (HRTs) and about 20 times smaller reactors, occupying four times less area. While size reduction makes MECs more attractive compared with MFCs, downstream combustion of produced gases, usually methane, in turbines occurs at notoriously low energy efficiencies between 30 and 40%. In contrast, the direct conversion of wastewater energy into electricity in MFCs eliminates the need for inefficient combustion steps. Despite intensive research and a long history, BES technology is still in its infancy. While there have been some attempts to commercialize BES for wastewater treatment, there is still room for improvement. For MFCs, this is arguably power density and for MECs it is gas production, which ultimately translates into power density as well. We therefore suggest to use dollar per peak power capacity ($Wp
1) as a benchmark for cost assessment. Dollar per power production ($[Ws]
1) would still be better but such data are usually not reported. Nanomaterials are promising leads to solve the problem of power density in wastewater BES. There are three different types of nanomaterials that can potentially be used in BES: carbon-based, metal-based, and mineral nanomaterials. These materials may be in suspension, electrode deposits, or in membranes. Depending on the location, they serve different purposes. On electrodes they may improve attachment of microorganisms, increase the catalytic surface area, reduce over potentials, and so forth.

1.2

Economic Assessment of BES Materials

Both MEC and MFC technologies face similar economic challenges when treating wastewater. They work better with high strength wastewater and are therefore better applicable for industrial waste streams where COD concentrations exceed 5 g L
1. To some extent, lower COD concentrations can be compensated for by larger volumes, which, however, also drive electrode costs as more electrode surface is required. Additionally, a potential can be applied in MECs to improve COD removal rates. This is accomplished at the expense of anodic corrosion. Anodes are therefore the bottleneck in MECs. In MFCs, the cost driver is the cathode. Cathodic oxygen reduction reaction (ORR) are limited by oxygen diffusion and catalyst wear. An excellent review on BES cathodes and their workings was published by Liu et al. (2014). As further detailed in the section on cathodes, some power outputs of MFCs using nanomaterials reach 12 Wp m
2 (Wen et al. 2012, 2014) as opposed to hydrogen fuels, which routinely exceed 5 kWp m
2 (Al-Baghdadi 2005). For our estimates we used a typical fuel cell voltage of 0.7 V. At this voltage, others suggested that a power output of 10–20 Wp m
2 would be required to operate MFCs profitably, but scaling factors were not taken into account and expensive precious metal cathodes were used for previous evaluations (Sleutels et al. 2012). With such low power output, neither the use of precious metals nor mixed metal oxide cathodes appears to be justified and cheap carbon should be considered, perhaps in combination with iron or manganese minerals (Table 1.1). As in conventional fuel cells, platinum on carbon (Pt/C) is often used in MFCs as well. Due to its high power densities (for MFCs), frequently surpassing 2 Wp m
2, it is also cheap on a

Carbon brush

Graphite plate

Graphene–steel Carbon felt

Carbon brush

N-doped graphene Carbon cloth

Goethite–steel C–cloth C–cloth C–cloth

C–cloth

Nanocomposite

Nanocomposite

Carbon

Graphene Graphene

Carbon

Graphene

Precious metal Precious metal Nanocomposite Graphene

Carbon

Graphene

Anode Carbon brush Graphene/ PEDOT Carbon felt

Category Precious metal Carbon

20

12.6 7 7 5

9

1.5

12.5

3.2 10

5

NA

78

Anode area in cm
2 NA 5

C–cloth

C–mesh C–cloth C–cloth C–cloth

C–paper

C–cloth

GDL

C–felt C–paper

Steel mesh

C–cloth

GDL

Cathode support C–cloth C–paper

1

NA 7 7 5

9

1.5

10

2 NA

1

1

10

Cathode area in cm
2 1 12

N-doped graphene Pt/C Pt/C CoTMPP Bio-modified graphene Carbon nanofibers

5

MnO2–graphene nanosheet N-doped graphene/CNT/ Co-Ni N-doped carbon nanofibers NA Fe/N– functionalized graphene N-doped nanosheet on graphene NA

14

0.5 NA 0.6 NA

2

NA

5

NA NA

NA

5

Catalyst load in mg cm
2 5 NA

Cathode catalyst Pt/C NA

Table 1.1 Ranking of MFC cathode nanomaterials in terms of areal performance

Single

Single Single Single Dual

Dual

Dual

Single

Dual Single

Dual

Dual

Single

Chambers Dual Dual

0.2

0.7 0.5 0.4 0.3

0.8

1.0

1.1

1.3 1.2

1.9

2.0

2.1

P in Wp m
2 (cathode) 2.8 2.1

(continued)

Peng et al. (2013b) Cheng et al. (2006) Cheng et al. (2006) Zhuang et al. (2012) Santoro et al. (2013)

Kirubaharan et al. (2015) Liu et al. (2013b)

Wen et al. (2014)

Hou et al. (2014) Li et al. (2012)

Chen et al. (2012)

Hou et al. (2016)

Wen et al. (2012)

Reference Hou et al. (2016) Wang et al. (2013b)

1 Economic Assessment of Nanomaterials in Bio-Electrical Water Treatment 5

Graphite granules/felt C–felt

C–felt

Mo–steel CNT/Au/TiO2 C–paste C–felt C–paper

Magnetite/CNT CNT/TiO2 on C–cloth C–cloth Graphene/ PANI on C– cloth Magnetite–steel

Nanocomposite

Precious metal

Metal Precious metal Carbon Nanocomposite Nanocomposite

Nanocomposite Nanocomposite

NA

16 NA

NA 4.5

1200 9 0.07 197 12

197

197

16

78.5

Anode area in cm
2 97 20

C–paper

C–cloth C–cloth

NA C–brush

Steel C–paper C–paste C–felt C–paper

C–felt

Si–wafer/ nanowires C–felt

C–felt

Cathode support RVC C–cloth

NA

NA 4

NA NA

1200 7 0.07 74 12

69

69

6

1

Cathode area in cm
2 194 20.25

Pt/C

NA NA

NA NA

Steel Pt–TiO2 CNT Fe(CO)5 Pt/C

Pt/C

Fe/Ca–PANI

PANI–graphene nanosheets MoS3

Cathode catalyst NA Pt/C

0.5

NA NA

NA NA

NA 0.5 5% NA 0.5

NA

0.6

NA

NA

Catalyst load in mg cm
2 NA 0.5

Single

Single Dual

Single Dual

Single Single Single Single Dual

Single

Single

Dual

Single

Chambers Single Single

NA

NA NA

NA NA

0.02 1.9 10
3 2.4 10
4 2.0 10
6 NA

0.04

0.04

0.1

0.1

P in Wp m
2 (cathode) 0.2 0.2

Peng et al. (2012)

Li et al. (2014) Hou et al. (2013)

Ledezma et al. (2015) Ledezma et al. (2015) Dumas et al. (2007) Wu et al. (2013) Reid et al. (2016) Wang et al. (2013a) Ghasemi et al. (2012) Park et al. (2014) Wen et al. (2013)

Zang et al. (2014)

Reference He et al. (2005) Thepsuparungsikul et al. (2014) Ren et al. (2013)

Wp peak power in W, NA data not available, PEDOT poly(3,4-ethylenedioxythiophene), GDL gas diffusion layer, CNT carbon nanotube, RVC reticulated vitreous carbon, CoTMPP Co–tetramethylphenylporphyrin, PANI polyaniline

Precious metal

Carbon Graphene

Carbon

Graphene

Anode RVC Single–walled carboxy–CNT Graphite

Category Carbon Precious metal

Table 1.1 (continued) 6 M. Siegert et al.

1 Economic Assessment of Nanomaterials in Bio-Electrical Water Treatment

7

dollar per power basis (~$500 Wp
1) compared to other materials (Hou et al. 2016). The runners-up are a potpourri carbon nanofibers and graphene composites. Nitrogen or manganese can be used to improve their performance, for example, N-doped nanofibers cost ~$500 Wp
1 (Chen et al. 2012) or MnO2-graphene costs ~$2400 Wp
1 (Wen et al. 2012). For graphene, single layer material ($200 g
1) was used for our estimate but other forms may further reduce costs. Plain reticulated vitreous carbon performed slightly better than graphene with $2100 Wp
1 (He et al. 2005). Other types of functionalized graphene were N-doped (Wen et al. 2014) or Fe/N-doped nanosheets (Li et al. 2012), both of which cost ~$4300 Wp
1. There were several extreme outliers, orders of magnitude above or below these values. Some of them used ferricyanide as an electron acceptor. One worth noting was N-doped graphene with only $3 Wp
1 (Hou et al. 2016). All of the abovementioned MFC cathodes use carbon materials as catalyst support. While moving closer to become economically feasible, the costs are still far away from what is necessary to operate MFC profitably. MFC cathodes need to be large, which is not only an engineering challenge but also drives material and also real estate costs. MECs, in contrast, can be operated completely without the use of precious metal catalysts, for example, with stainless steel (Call et al. 2009). Steel is among the cheapest materials (0.4 mm steel plate, type A286, ~$430 m
2; type 304 ~$80 m
2) and can also serve as structural element in MECs, further reducing cost. As explained above, in MECs, anodes are the critical component. Escapa et al. (2012) estimated that an MEC WWTP producing hydrogen gas can be operated economically when costs of the anode compartment are below $1500 m
3 (at 2012 exchange rates). This agrees well with our own, independent calculation (Fig. 1.3), except that we use methane gas as product with a scaling coefficient of 0.6 (Tribe and Alpine 1986). While the scaling coefficient is somewhat arbitrary, existing commercial

107 Total electrode cost in log10 USD

Fig. 1.3 Relationship between MEC reactor capacity and total electrode cost including anode and cathode. Errors are standard deviations of four different tubular reactor designs. Anodes are graphite granules and cathodes are steel pipes

106

105

104

103

102 100

101

102

103

Total reactor capacity in log10 kW

104

8

M. Siegert et al.

MEC WWTP are living evidence for this technology’s economic viability. For our calculation we used graphite granule anodes in titanium mesh drums inserted into steel pipe cathodes which also serve as MEC housing. With these assumptions, we found that electrodes make up 5  2% of total reactor costs. Future scaling experiments will show how these costs contribute in reality. Unfortunately, the data reported for MECs are, as for MFCs, not standardized. Patil et al. (2015) suggested a framework for reporting MEC performance and we refer future investigators to this excellent article in order to standardize their reports.

1.3 1.3.1

Nanomaterials Used BES Carbon-Based Nanomaterials

Nanomaterials in BES can, as in nonbiological applications, improve electron transfer processes (Cai and Chen 2004) and catalytic activity. Carbon nanoparticles, nanofibers, nanotubes (CNTs), and graphene are the most intensively studied nanomaterials in BES. Their synthesis is relatively simple (Box 1.1) and carbon as support material is inexpensive and widely available. It is therefore not surprising that CNTs have been used in BES where cheap electricity or gas production are the primary goals. In Shewanella, one of the model microbes in BES, CNTs redirected the electron flow between the electron donor lactate and nitrobenzene as an electron acceptor (Yan et al. 2014). The altered electron flow occurred mostly outside Shewanella cells whereas nitrobenzene had been reduced inside without CNTs, suggesting that CNTs can also improve electron flow between cells and electrodes. Carbon nanoparticles also supported hydrogenase-dependent hydrogen oxidation on indium tin oxide (ITO) electrodes (Szot et al. 2013). The ITO electrodes were coated with carbon nanoparticles films which were then used to immobilize the hydrogenase enzymes. Box 1.1: Preparation of Carbon Nanotubes (CNTs) Ab initio, CNTs are prepared by high temperature techniques like arch discharge and laser ablation. Due to the high energy consumption, these methods have been replaced by low temperature ( ZACl-3 > MACl-3 The adsorption data presented in Fig. 3.2 are fitted to the Langmuir and Freundlich isotherm models. The experimental isotherm data along with predictions using the Langmuir and Freundlich isotherm models (Eqs. 3.2 and 3.1) have also been presented in Fig. 3.2. Figure 3.2 reveals that the Langmuir model gives a better fit for the isotherm data as compared to the Freundlich model. Although both the Langmuir and Freundlich isotherm models fitted well with the experimental data at lower selenite concentrations, the Freundlich model gets deviated more and more as the concentration of selenite ions increased. The Langmuir model fitted well with the experimental data throughout the concentration range. Consequently, the Langmuir isotherm model can be considered to estimate and evaluate the adsorption of selenite ions on synthetic anionic clay adsorbents in aqueous medium at 25  C under simulated environment. In real-life scenario, influence of the parameters like solution pH, presence of other ions and concentration of the contaminant (in the present case, the selenite ions) in water have to be taken into consideration.

3 Water Pollution Remediation Techniques with Special Focus on Adsorption

45

Fig. 3.2 Langmuir and Freundlich adsorption isotherms for selenite uptake on the LDHs at 25  C. The symbols indicate experimental data. Solid/dotted lines indicate model predictions (Initial selenite concentration: 0–125 mg/L, adsorbent dose: 0.5 g/L, contact time: 4 h) (reproduced with permission from Mandal et al. 2009a)

3.2.3

Adsorbents

Adsorbent plays the key role towards the success of an adsorption process. Exploitation of large variety of natural and synthetic adsorbents with varying surface characteristics, which are responsible for the selective adsorption of specific contaminant from water, has led to the rapid growth of research interests in this field. The efficiency of an adsorbent to remove a specific species of contaminant depends on the following factors: • Surface characteristics of the adsorbent, viz. surface charge (pHpzc), surface area and pore structure. • Nature of the adsorbate such as its pKa, functional groups present, polarity, ionic/molecular weight and size. • Water pH and the adsorbate concentration. The commonly used adsorbents in water treatment are activated carbon obtained from various sources, activated alumina, natural and synthetic clays, metal oxides and hydroxides and non-conventional low-cost adsorbents (biopolymeric adsorbents, agricultural and industrial byproducts). In addition to the conventional adsorbent materials, advanced materials like magnetic and other inorganic nanomaterials, carbon nanotubes (CNTs), organic–inorganic (nano) composites, etc. have attracted considerable research interests including their application in water purification.

46

S. Mandal et al.

These nanomaterials and their application as adsorbent in water and wastewater treatment have also been discussed in this chapter.

3.2.4

Activated Carbon

Use of carbon in water purification was originated in ancient era. Charcoal was used for drinking water treatment by ancient Hindus in India, and carbonized wood was used as a medical adsorbent and purifying agent in Egypt during 1500 B.C. (Cheremisinoff and Angelo 1980). Even in this highly advanced modern era, it is probably the most widely used adsorbent with high reliability to control micropollutants and organic contaminants in water. Carbon absorption is a widely used method of water treatment at home because of its ability to improve water quality by removing micropollutants, disagreeable tastes and odours, including objectionable chlorine. Organic contaminants like phenols, benzene, toluene, aniline, pyridine, organic dyes, pesticides, etc. are removed from water expeditiously by adsorption on activated carbon. Activated carbon can be made from variety of sources such as wood, coal, agricultural wastes, peat and petroleum residues. They are highly porous materials that usually possess exceptionally high surface area, ranging between 500 and 1500 m2/g (Kandasamy et al. 2009). The porous nature of carbon helps to trap microscopic particles and large organic molecules, while the activated surface areas facilitate adsorption of small organic molecules. The four different forms of activated carbon that are known for water treatment applications are powder (PAC), granular (GAC), fibrous (ACF) and cloth (ACC). All these four forms of activated carbon obtained from various sources have been investigated for drinking water as well as wastewater purification (Sontheimer et al. 1988; Phan et al. 2006). It is observed that surface acidity of the GAC plays an important role for the adsorption of hydrophobic synthetic organic contaminants from water. An increase in surface acidity increases the polarity of the surface, and subsequently the adsorption capacity of GAC decreased towards hydrophobic synthetic organic contaminants (Karanfil and Kilduff 1999). The performance of activated carbon fibers (ACF) is significantly higher than that of granular activated carbon (GAC) in terms of adsorption rate and selectivity for organic micropollutants such as nitrophenol, chlorophenol and benzoic acid (Le Cloirec et al. 1997). The process of activated carbon adsorption is highly effective and well established for removing microorganisms and organic contaminants from water; however, the adsorbent is not effective in removing toxic inorganic species including heavy metals from water. Regeneration is an important consideration in the use of activated carbon for water and wastewater treatment. The economy of the adsorption process greatly depends on reuse of the activated carbon as they are expensive. Various regeneration techniques such as thermal regeneration, chemical regeneration, electrochemical regeneration and wet air oxidation have been attempted; however, thermal regeneration process is regarded as the most widely accepted technique. During thermal regeneration process,

3 Water Pollution Remediation Techniques with Special Focus on Adsorption

47

the organic contaminants are destroyed by heating at 800  C with concomitant regeneration of the activated carbon. It is feasible to regenerate granular carbon by conventional thermal technique for at least 15 cycles of successive saturation and regeneration (Weber 1974).

3.2.5

Metal Oxides and Hydroxides

Adsorption of metals and metalloids from aqueous solution onto the surfaces of various metal oxides (Table 3.1) and oxy-hydroxides are well known and the process is considered to be an important means of trace metal transport in many natural systems. Leaching of metals and minerals from underneath clay and rocks, interactions between sediments and the water column in natural aquatic systems, stripping of toxic metals in landfills and ponded fly ash, and the use of adsorption for removal or recovery of trace metals/metalloids in wastewater and water treatment operations are some of the examples of such process (Benjamin and Leckie 1981). In recent years, there have been numerous experimental investigations of adsorption including studies of both complex natural systems and well-characterized model solutions. The mechanism of adsorption on metal oxides and oxyhydroxide surfaces are studied using surface complexation modelling. In aqueous systems, the surfaces of oxides are covered with hydroxyl groups. The acid-base equilibrium of a hydroxylated oxide surface is commonly represented as:  S
OHþ 2 $

 S
OH þ Hþ Ka1

ð3:3Þ

 S
OH

 S
O
þ Hþ

ð3:4Þ

$

Ka2

where S
OH2+, S
OH and S
O
represent positively charged, neutral and negatively charged surface hydroxyl groups, respectively, and Ka1 and Ka2 are the acidity constants (Dzombak and Morel 1990). The adsorption of a metal ion on the oxide surface involves the formation of bonds of the metal ion with the surface oxygen atoms and the release of protons from the surface.  S
OH þ M2þ

$

 S
OMþ þ Hþ

KM

ð3:5Þ

where M2+ represents a divalent cation and KM is surface complexation constant. At near-neutral pH, the theoretical affinity for anion sorption on metal oxides follows the order (Manning and Goldberg 1996): PO4 > SeO3 > AsO4 > AsO3 >> SiO4 > SO2 > F > BðOHÞ3 The most exploited oxides and oxy-hydroxides used for water purification applications by adsorption are of aluminium, iron and manganese (Jeong et al. 2007).

48

S. Mandal et al.

Table 3.1 Surface characteristics and selectivity of commonly used adsorbents in water treatment Sl. no.

Surface area (m2/g)

pHpzc

Contaminant selectivity

Adsorbent

Types

1.

Activated carbon and related adsorbents

GAC, PAC, ACF, ACC

500–1500

7.5–9.0

2.

Metals oxides and hydroxides

Activated alumina

150–260

6.2–9.6

Iron oxide and hydroxides Fe-Mn binary oxides and hydroxides Montmorillonite Kaolinite Sea nodule

4–700

6.5–8.8

265

5.9

Arsenic, selenium, phosphates Arsenic

18.6 9.1 –

8.16 8.42 7.4

Arsenic and heavy metals Arsenic

Synthetic clay

50–250

8–12

Rice husk, bagasse, coconut coir, coconut shell Chitosan Alginate Untreated cellulose fibers Microfibrillated cellulose

–

–

Anionic species, viz. fluoride, As(III), As (IV), selenium, Cr (VI), azo dyes, surfactants and pesticides Arsenic, heavy metal ions

4–40 – 0.2–1

6.8–7.3 6.5 3.6–7.1

Heavy metal ions Arsenic

–

3.9

Ni(II), Cu(II) and Cd (II)

3.

4.

5.

Natural and synthetic clay, ores and minerals

Non-conventional low-cost adsorbents Biopolymeric adsorbents

Organic pollutants, viz. phenols, benzene, toluene, aniline, pyridine, dyes and pesticides Fluoride, arsenic, selenium

References Dabrowski et al. (2005), Al-Degs et al. (2008), Mandal and Kulkarni (2011), and Menya et al. (2018) Choi and Chen (1979), USEPA (1980, 1984), Ku and Chiou (2002), Mohan and Pittman Jr. (2007), Su et al. (2008), and Craig et al. (2017) Cornell and Schwertmann (2003) Zhang et al. (2007a, b)

Manning and Goldberg (1996) Bhattacharjee et al. (2003) Goh et al. (2008), Mandal and Mayadevi (2009), and Mandal et al. (2009a, b)

Daifullah et al. (2003), Amin et al. (2006), and Ahmaruzzaman and Gupta (2011) Kamari et al. (2009) Escudero et al. (2009) Fiol and Villaescusa (2009) Hokkanen et al. (2014)

Application of inorganic nanoparticles in water treatment has gained significant importance in the recent decade (Konstantinos et al. 2016), which has been discussed in detail under the heading ‘nanosorbents’.

3 Water Pollution Remediation Techniques with Special Focus on Adsorption

3.2.6

49

Activated and Impregnated Alumina

Activated alumina, aluminium hydroxides and alumina impregnated with various metal oxides and hydroxides (manganese, magnesium, ferric hydroxides, etc.) have found enormous applications as adsorbent in water treatment as well as in industrial separations. Activated alumina is a calcined granular form of aluminium oxide (Al2O3). The high area to weight ratio, excellent pore structure and high thermal stability of activated alumina adsorbent make them ideal choice for environmental and industrial researchers. Activated alumina has been used for defluoridation of water since 1934. The most practical community level technology for defluoridation of drinking water is adsorption with activated alumina (γ-aluminium trioxide), which shows a high affinity and selectivity towards fluoride ions (Sorg 1978; USEPA 1980; Bhatnagar et al. 2011). Fluoride concentration in the effluent can be reduced below the level of 1.0 mg/L using this treatment process. This defluoridation technique is largely affected by the factors such as pH, influent fluoride concentration, media particle size and the presence of competing ions (arsenic, selenium, silica, hardness ions) (USEPA 1978; Choi and Chen 1979; Schoeman and Macleod 1987). The pHpzc (point of zero charge) of the adsorbent is an important parameter in fluoride sorption. pHpzc of an adsorbent is the pH at which the adsorbent has an overall neutral charge. At a pH, above and below the pHpzc, the adsorbent surface becomes negatively and positively charged, respectively. Below the pHpzc of activated alumina (8.2, typical zero point charge), the alumina surface has a net positive charge, and it has a strong tendency to adsorb anionic species in aqueous medium. The pHpzc value is different for different forms or types of the alumina and hence their defluoridation efficiency. A number of studies have shown that the maximum fluoride removal efficiency of activated alumina is achieved at the optimum pH between 5.5 and 6.0 (USEPA 1978; Choi and Chen 1979). At this pH, the attraction of fluoride ions towards the activated alumina surface is most favourable and the interference by competing ions and silicate is minimal. An adjustment of the water pH may be required prior to the treatment process in order to achieve the maximum removal efficiency. Also it is important to choose the proper grade of activated alumina for its effective reuse in multiple defluoridation cycles. Yang et al. (2007) concluded that differences observed in surface morphology of activated alumina can be attributed to the ratio of singly and doubly coordinated Al–O bonds, the greater bonding yields higher surface acidity. This higher surface acidity is typically correlated to an enhanced positive surface charge and thus larger fluoride sorption due to increased electrostatic interactions. According to Yang et al. (2007), the surface acidity of three different types of aluminium (hydr)oxides (gibbsite, γ-alumina and α-alumina), with different ionic strengths, are in the order of α-Al2O3 > γ-Al(OH)3 > γ-Al2O3. The difference in surface acidity apparently arises from the difference of the surface chemical composition of various alumina minerals. It is practical that different intermediate forms of aluminol groups [AlOH, Al2OH, Al(OH)2] may exist at the hydrated alumina surfaces. However, studies have failed to specify whether interactions

50

S. Mandal et al.

between fluoride and alumina surfaces form inner or outer-sphere complexes. Possible surface-exchange reactions can be expressed as follows:
AlOHþ 2surface þ F

AlOHsurface þ F




$ $

AlFsurface þ H2 O

ð3:6Þ




ð3:7Þ

AlFsurface þ OH

Desorption of fluoride from the activated alumina surface depends on the pH of the environment. At a neutral pH, the most dominant species that formed during desorption are AlF3 and F
. In developed countries, activated alumina adsorption has been the method of choice for defluoridation of drinking water. Around 500–1500 L of safe water could be produced with 3 kg of activated alumina at natural water pH (7.8–8.2), when the raw water fluoride concentration is between 11 and 4 mg/L. Generally, it is implemented on a large scale at point-of-source for community water treatment plants. A few pointof-use defluoridation units have also been developed, which can be directly attached to the water tap. During recent years, this technology is gaining wide attention even in developing countries. In India, domestic defluoridation units have been developed using indigenously manufactured activated alumina. Because of the factors like high affinity between fluorine and aluminium, their high porosity, availability of large number of surface active sites, aluminium based adsorbents such as poly aluminium chloride, poly aluminium hydroxyl sulphate, alum sludge (aluminium hydroxides), activated alumina and alumina impregnated with various oxides (manganese, magnesium, ferric hydroxides, etc.) are also found to be suitable for defluoridation of water as compared to other defluoridation methods. Activated alumina adsorbent is also used for the removal of arsenic and selenium from water. Arsenic removal from groundwater by activated alumina adsorption is classified as one of the best available technologies for arsenic remediation by United States Environmental Protection Agency (USEPA 1984, 1986; Mohan and Pittman 2007). Adsorption of oxyanionic species of arsenic on the alumina surface takes place in the pH range between 6 and 8, i.e. below the pHpzc of alumina. Activated Al2O3 has been effectively used for arsenic removal from drinking water at the Fallon, Nevada, Naval Air Station (Hathaway and Rubel 1987). Selenium adsorption on activated alumina surface has been studied by several researchers (USEPA 1980; Su et al. 2008). It is demonstrated that sulphate and bicarbonate ions has no effect on Se(IV) adsorption but greatly affected Se(VI) adsorption by activated alumina. Most fluoride removal applications are long term, and hence necessitate regeneration of the alumina. The regeneration of alumina is generally performed by chemical method. There are three known methods that are usually followed for the chemical regeneration of activated alumina adsorbents: (1) NaOH/H2SO4, (2) Al2(SO4)3 and (3) H2SO4. An efficient regeneration was achieved with a combination of 1% NaOH and 0.4 N H2SO4. A simple regeneration process named as ‘dip regeneration procedure’ was developed, which is appropriate for a rural set up. This requires the transfer of activated alumina from domestic units to a nylon bag (mesh size 0.109 mm), dipping the bag in 1% NaOH solution for 8 h (or overnight) with intermittent mixing.

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The material was then washed with water to remove the excess alkali followed by dipping the bag in 0.4 N H2SO4 solution for another 8 h. Finally, the activated alumina is washed with raw water to raise the pH between 6 and 7 followed by drying at 120  C to get the adsorbent ready for next cycle. The only limitation of ‘dip regeneration procedure’ is that the process takes long time. Presently, the ‘bucket regeneration procedure’ has been developed, which addresses the disadvantage of the ‘dip regeneration procedure’. In ‘bucket regeneration procedure’, a plastic bucket with flow control device is used to continuously pass the regenerant (alkali and acid solutions) over the exhausted activated alumina bed (Daw 2004). Studies indicated that the time required for regeneration decreases substantially. Also the method is user-friendly and can be easily adopted in rural setups.

3.2.7

Iron Oxides and Oxy-Hydroxides

Iron oxides and oxy-hydroxides are among the most important transition metal compounds of technological importance that are readily available in nature and can be easily synthesized in the laboratory. They may present in a number of different mineralogical forms or phases. Sixteen pure phases of iron oxides, hydroxides or oxy-hydroxides are known till date (Mohapatra and Anand 2010). Among all, goethite (αFeOOH), ferritehydrite (Fe5HO8. 4H2O) and hematite (αFe2O3) are the most commonly used compounds for water treatment applications. In most of its phases, iron is present in oxidation state of three. Except Schwertmannite and ferrihydrite, all iron oxides and oxy-hydroxides are crystalline in nature. Iron oxide and oxy-hydroxides forms one of the most powerful groups of adsorbents used for water treatment applications. Various forms or phases of iron oxide and oxy-hydroxides are used for the removal of many toxic cations (Co, Zn, Pb, Cd, Cs, U, Sr, etc.) and anions like AsO43
,CrO42
,PO43
,CO32
, etc. The characteristics of iron oxide or oxy-hydroxides that govern the process of adsorption and retention of cationic or anionic species on its surface are specific surface area, surface charge, pHpzc and surface ionization. Significant advances have been made in recent years in the development of phenomenological models, in particular, the concept of surface ionization and complexation, to describe trace metal adsorption on hydrous iron oxide surfaces (Davis and Leckie 1978). Benjamin and Leckie (1981) have reported multiple-site adsorption of Cd, Cu, Zn and Pb on amorphous iron oxy-hydroxide (Benjamin and Leckie 1981). According to Benjamin and Leckie, for each metal, there is a narrow pH band where fractional adsorption increases from near nil to near 100%. Oxy-hydroxides of iron are extensively used in removing oxyanionic species of arsenic (arsenate and arsenite), selenium (selenate, selenite), molybdenum (molybdate), antimony and anionic contaminant like phosphate from water (Peak and Sparks 2002; Fujita et al. 2006; Mohan and Pittman 2007). In the viewpoint of drinking water treatment, iron oxy-hydroxides have found tremendous application in removing arsenic from groundwater.

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Extensive studies are carried out to understand the adsorption mechanism of oxyanionic species of arsenic and selenium on iron oxy-hydroxides. Though the concept of formation of inner-sphere and outer-sphere surface complexes exists, there are views on formation of both types of surface complexes of iron oxy-hydroxides with oxyanionic species of arsenic and selenium (Goldberg and Johnston 2001; Peak and Sparks 2002). Both the surface area and surface charge of the adsorbent particles strongly influence the binding process. Unlike oxyanionic species of arsenic and selenium, the sorption of phosphates by ferrihydrite takes place through a complex mechanism. The sorption of phosphates by ferrihydrite includes both sorption and precipitation. Sorption of phosphates takes place on the surface of ferrihydrite particles through the following reaction:  Fe
OH þ H2 PO4


!

 Fe
O
H2 PO3 þ OH


ð3:8Þ

Because of its strong affinity for metal oxides and its similarity to the arsenate ion, phosphate is the anion considered most likely to compete with arsenic for adsorption sites. It is possible to increase the selectivity of iron oxyhydroxides towards a specific contaminant by tailoring its composition during synthesis. Fe-Mn binary oxide is one such material, which is found to be an excellent adsorbent for both arsenite [As (III)] and arsenate [As(V)] species in water (Zhang et al. 2007a). Manganese in the adsorbent act as oxidizing agent to oxidize As(III) to As(V), which simultaneously gets adsorbed on the Fe-Mn oxide (adsorbent) surface (Zhang et al. 2007b). The immobilization and practical application of Fe-Mn composite materials may be a major research avenue in the arsenic removal. Theoretically, the regeneration of hydrous ferric oxide is possible; however, disposal of the sludge saturated with the contaminant is more preferred. There are some reports on regeneration of this class of adsorbent by alkali treatment. For example, regeneration of fluoride-rich hydrous ferric oxide using 1.0 M NaOH solution is reported by Dey et al. (2004) and a maximum of 75% regeneration was achieved (Dey et al. 2004).

3.2.8

Natural and Synthetic Clay, Ores and Minerals

Natural clay, ores and minerals cover a broad range of adsorbents that are used in water treatment studies. Laterite, gibbsite, calcite, goethite, kaolinite, bentonite, natural zeolite (s), sea nodule, raw and chemically treated soil and clay, iron oxide coated sand, China clay and Fuller’s earth are among the naturally occurring clays and minerals that have been explored for water purification by adsorption (Dube et al. 2001; Bhattacharjee et al. 2003; Maity et al. 2005; Karageorgiou et al. 2007; Wang and Peng 2010). In general, the clay, ore or minerals themselves do have some adsorption capacity for specific contaminant, which can be increased through ‘activation’ by chemical treatment, calcination or air-drying. None of these natural minerals can be considered as universal adsorbent, nevertheless, they can be used in the local area and thus will

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have low or no cost. It may be considered as the medium of choice for the particular area, where the mineral is available. For example, Chakravarty et al. (2002) have reported the use of iron-manganese bearing lateritic ore for adsorptive removal of arsenic from groundwater of Jharkhand state in India, where plenty of this mineral is available (Chakravarty et al. 2002). The adsorbent is reported to have excellent adsorption property and successfully removes both As(III) and As(V) from water in household filtering system. However, it may not be feasible to use the same lateritic ore in other corner of the world; nevertheless, the information can be used to explore minerals having similar composition for arsenic remediation application. Adsorbent regeneration is impractical for this class of adsorbent as the process cost may exceed the adsorbent cost and hence may not be economical. Among synthetic clays, hydrotalcite type anionic clay adsorbents have attracted considerable attention due to their easy synthesis procedure, tailored properties and regenerability. They also possess good catalytic properties. The pH at point of zero charge (pHpzc) of this class of clay usually lies between 8 and 12 (Bergaya et al. 2006; Mandal and Mayadevi 2009) depending on the composition and nature of interlayer anion. The high pHpzc value indicates that they have positive surface charge at the natural water pH, and hence are good adsorbent for removing anionic contaminants from water near neutral pH. They are highly effective even in very low adsorbate concentrations. They have shown great adsorption potential towards anionic and oxyanionic species (Goh et al. 2008) that includes fluoride (Mandal and Mayadevi 2008a, 2009), arsenic (Turk et al. 2009), phosphates (He et al. 2010), selenium (Mandal et al. 2009a), organic dyes (Mandal et al. 2009b), anionic surfactants (Zhang et al. 2012) and pesticides (Li et al. 2005) in aqueous medium. The sorption mechanism of these synthetic anionic clays involves both surface adsorption and ion exchange. Because of this dual characteristic, hydrotalcites form a promising adsorbent in water purification applications. Wang et al. (2006) have reported a novel hydrotalcite-supported Pd-Cu material that utilizes both adsorptive and catalytic property of hydrotalcites for nitrate adsorption and reduction from water. Though these synthetic clay adsorbents have many advantages, they suffer from the limitation of influence of other co-existing ions present in water. These synthetic clay adsorbents have not been implemented so far in household or community level water treatment. The potential of these adsorbents need to be studied in detail for their prospective field application.

3.2.9

Non-conventional Low-Cost Adsorbents

In the recent decade, development and exploitation of non-conventional low-cost adsorbents has led to the rapid growth of research interests in the field of adsorption (Babel and Kurniawan 2003; Crini 2006; Sud et al. 2008; Bhatnagar and Sillanpaa 2010). This group of adsorbents mostly includes agricultural byproducts (sugarcane bagasse, rice husk, wheat straw, coconut coir, coconut shell, sugar beet pulp, fruit

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pulp, saw dust, etc.) and their chemically modified forms, biopolymeric materials (Chitin/chitosan, alginate, guar-gum, cellulose) and industrial bi-products (red mud, fly ash, papermill sludge, etc.).

3.2.10 Agricultural Waste and Byproducts Agricultural waste and byproducts such as rice husks are used as adsorbent for the treatment of various contaminants from water and wastewater (Daifullah et al. 2003; Ahmaruzzaman and Gupta 2011; Sangeetha et al. 2017). Rice husk possesses a granular structure with reasonably high chemical stability and high mechanical strength that makes it a good adsorbent material. Amin et al. (2006) have used rice husk for arsenic removal from groundwater (Amin et al. 2006). It is reported to adsorb both As(III) and As(V) and the uptake of arsenic increased with increasing temperature (Nasir et al. 1998). Complete removal of arsenic was also achieved in a fixed bed column packed with rice husk. Sugarcane bagasse is another agro-industry waste that has been explored extensively in water treatment applications in its untreated and chemically treated forms. Sugarcane bagasse ash has been tested for the adsorption of heavy metal and organic dyes in aqueous medium (Isa et al. 2010). The role of sawdust collected from timber working shop is evaluated for copper removal and is reported to be a good adsorbent for copper removal in aqueous medium (Ajmal et al. 1998).

3.2.11 Biopolymeric Adsorbents Numerous approaches have been studied for the development of natural biopolymeric adsorbents for effective and selective removal of contaminants from water. Biopolymers derived from microorganisms and plants such as alginate, chitosan, guar-gum and cellulose show very active chemical behaviour with the heavy metal ions. These natural biopolymers have some common typical properties, such as non-toxicity, biocompatibility, biodegradability, poly-functionality, high chemical reactivity, chirality, chelation and adsorption capacities, which make them a favourable choice for environmental scientists. Alginate is a water-soluble linear polysaccharide extracted from brown seaweeds and it is composed of alternating blocks of 1–4 linked α-L-guluronic and β-Dmannuronic acid fragments. It has a typical gel-forming property by ion-exchange with alkali metal ions such as calcium, which makes it an adsorbent of choice to be used alone or as composite, for heavy metal removal from water. Composite alginate beads are studied by several researchers for the adsorptive removal of arsenic and some heavy metals like Cu, Ni, Cd, Cr, Pb, etc. from water (Al-Rub et al. 2004; Park and Chae 2004).

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Chitosan is a naturally abundant polysaccharide-based biopolymer, which is derived from chitin, a major component of crustacean shells and fungal biomass (Prasad et al. 2017). It is readily available from seafood processing wastes, and hence chitosan may be produced commercially at low cost. Chitosan is hydrophilic and it has the ability to form complexes with metals. It is also a nontoxic, biodegradable and biocompatible material. Like alginate, chitosan also has a tendency to agglomerate or form a gel in aqueous media due to which the active binding sites are not readily available for sorption when it is in a gel or in its natural form. To increase the accessibility of the active surface sites, chitosan composites are formed with inorganic, organic or polymeric materials and studied for their adsorption efficiency. Several attempts have been made on chemical modification of the chitosan structure and their performances have been evaluated through the adsorption of heavy-metal ions and various organic pollutants from aqueous solutions. Among all natural biopolymers, cellulose is the most abundant in nature. Various researches are focused on the use of raw or modified cellulose and its derivatives with good sorption properties for water purification applications. Studies on Fe(III) loaded cotton cellulose for arsenic removal (Zhao et al. 2009) and surface modified cellulose for removing fluoride and arsenic ions from water have been reported (Tian et al. 2011). It has been found that surface modified cellulose fibres are efficient adsorbent for removing both fluoride and arsenic ions from water. Though the possibility of using biopolymeric adsorbents for water treatment has been widely demonstrated, however, the low mechanical and chemical resistance of most of these biopolymers is a disadvantage, which restricts their application in real scale adsorption processes. Hence, these materials are mostly used in their supported or composite forms. Various methods such as entrapment, encapsulation, adhesion, crosslinking and grafting have been used for the preparation of composites from these biopolymeric materials in combination with other inorganic clay or polymers and studied for removing organic and inorganic contaminants from water (Hokkanen et al. 2016). Crosslinked chitosan beads formed with various crosslinking agents like glutaraldehyde, carboxylic acids, sodium trimetaphosphate, sodium tripolyphosphate, etc. have shown several characteristic properties like stability in acidic, alkaline as well as organic solvents, mechanical stability, which is advantageous in terms of operation, faster kinetics and diffusion properties (Crini 2006). Cellulose supported hydrotalcites (Fig. 3.2) for fluoride removal (Mandal and Mayadevi 2008b), fluoride removal by lanthanum alginate beads (Huo et al. 2011), chitosan-coated perlite beads for cadmium removal (Hasan et al. 2006), chitosan/PVA hydrogel beads for lead removal (Jin and Bai 2002), copper removal by calcium alginate enclosed organic resins (Jodra and Mijangos 2003), arsenic removal by iron oxide loaded alginate beads (Zouboulis and Katsoyiannis 2002), removal of hexavalent and trivalent chromium ions by composite alginate beads (Kim et al. 2008), vinyl monomer glycidyl methacrylate (GMA) grafted cellulose for copper adsorption (O’Connell et al. 2006), and crosslinked chitosan beads for nitrate removal (Chatterjee et al. 2009) are some of the examples of biopolymeric composites that are explored for water purification studies.

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Use of biopolymeric adsorbents in their pure form or in composites is an attractive option for water treatment applications. However, the technology has not reached its stage of maturity and more detailed studies are required for its practical use. Advancement is required towards the field of process development, adsorbent regenerability and environment friendly disposal of the sludge.

3.2.12 Industrial Byproducts and Wastes Industrial byproducts such as chars, coals along with red mud, fly ash, paper mill sludge and their chemically modified forms are studied for their possible application in wastewater treatment. Use of peat in the control of industrial wastewater pollution is well documented. Oak bark, pine bark, oak wood and pine wood chars obtained from fast pyrolysis are investigated as adsorbents for the removal of the heavy metal ions, viz. As3+, Cd2+ and Pb2+, from water. Red mud, the waste material formed during the extraction of alumina from bauxite by Bayer process, has been explored as an alternate adsorbent for arsenic removal from water. It has been observed that the As(III) adsorption on red mud is favoured by an alkaline aqueous medium (pH 9.5), whereas an acidic pH (1.1–3.2) was effective for As(V) removal (Altundogan et al. 2000) by the red mud. Thermal and acid treatment of the red mud resulted into an increase in arsenic adsorption capacity (Altundogan et al. 2002). Fuhrman et al. (2004) reported activated seawater-neutralized red mud (Bauxsol) for removing inorganic arsenic (As) from water (Fuhrman et al. 2004). Red mud is also studied for the adsorption of organic dyes, Chromium, lead and zinc in aqueous solutions. Like red mud, fly ash is another industrial byproduct that is generated from the combustion of coal. Presently, its application is limited only for making cement, bricks and roads. Environmental scientists are continuously exploring the possibilities of using these materials as adsorbents for water and wastewater treatment, and consequently reducing the solid waste of the environment (Chaturvedi et al. 1990; Albanis et al. 2000). Industrial sludge such as paper mill sludge is studied for some selected heavy metal adsorption in water by Calace et al. (2002) and the adsorption affinity of the paper mill sludge is reported to follow the order: Cu(II) > Pb(II) > Cd (II) ffi Ag(I) (Calace et al. 2002).

3.2.13 Nano-adsorbents With the advancement of materials science and technology, new advanced materials with improved characteristics and hence better adsorption properties are being synthesized and tested for water and wastewater treatment. In the past decade, nanoscale materials have become very important because they exhibit a series of unique physical and chemical properties. Among all, one very important property of this class of

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Table 3.2 Surface characteristics and selectivity of nano-adsorbents in water treatment Sl. no. 1. 2.

3. 4. 5.

Adsorbent Magnetic MCM-41 Nanomaterials (single/multi walled CNTs) Alumina supported on CNTs Humic acid coated Fe3O4 Amine functionalized Fe3O4

Surface area (m2/g) 800 150–3000

pHpzc 3.5 2.2–11.8

Contaminant selectivity As(V), Cr(VI) Heavy metal ions

165

–

Fluoride ion

64

2.3

–

–

Heavy metal ions Heavy metal ions, Pb, Cd, Cu Organic dyes and drug molecules Nickel

6.

Fe3O4/SiO2/HPGCOOH

–

2.1

7.

Fulvic acid decorated Fe3O4 magnetic nanocomposite Ce impregnated fibrous protein

51.5

2.2

–

9.0

8.

Fluoride, arsenate and phosphate ions

References Chen et al. (2009) Rao et al. (2007), Xin et al. (2012), and Khajeh et al. (2013) Wang et al. (2002) Liu et al. (2008) Xin et al. (2012)

Zhou et al. (2010)

Fu et al. (2015)

Deng and Yu (2012)

materials is their adsorption capacity to several heavy metal ions and many organic compounds (Khajeh et al. 2013). The unique properties of nanoparticles, such as small size, high reactivity, large surface area and large number of active sites for interaction with different contaminants, make them ideal adsorbent materials for the treatment of water and wastewater (Ali 2012). Nano-adsorbents (Table 3.2) can be classified into carbonaceous nanomaterials [carbon nanotubes (CNTs) and carbon nano-fibers (CNFs)], inorganic nanoparticles (inorganic metal, metal oxides, metal hydroxides, magnetic materials, silicon nanomaterials) and organic–inorganic nanocomposites (containing biopolymeric templates). Magnetic nanoparticles, carbon nanotubes (CNTs), graphene/graphene oxide, silicon nanomaterials, inorganic metal/-oxide/-hydroxide nanoparticles and organic–inorganic nanocomposites are the most extensively studied nanomaterials that are synthesized and tested as adsorbents for water and wastewater treatment (Chen et al. 2009; Xin et al. 2012; Khajeh et al. 2013; Sankar et al. 2013; Konstantinos et al. 2016). In terms of regeneration and reusability of the adsorbent, magnetic nanoparticles have been preferred over many other adsorbents (Ali et al. 2017). However being inorganic oxide material, in aqueous systems and under oxidative atmosphere, magnetite nanoparticles are highly prone to oxidation and aggregated, resulting into a decrease in adsorption capacity and reduced saturation magnetization. To overcome these limitations, magnetic nanoparticles coated with various organic compounds have been developed and improvement in the adsorption

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Fig. 3.3 Scheme of the removal of heavy metals with the humic acid coated Fe3O4 magnetic nanoparticles (Liu et al. 2008)

capacity as well as dispersity of the magnetic nanoparticles (Fig. 3.3) was observed (Liu et al. 2008; Fu et al. 2015). Moreover, the organic coating helps to retain the saturation magnetization of the magnetic nanoparticles by protecting the nanoparticles from direct contact of the oxidative environment. Several different organic compounds coated magnetic Fe3O4@SiO2 nanoparticles have also been designed, developed and tested for adsorption of heavy metals, dyes and drug molecules from water (Zhou et al. 2010). The advantage of easy regenerability for reusability and good dispersity of the magnetic nanoparticles have been retained by coating them with organic molecules. Being nanoparticles, both magnetic materials and CNTs possess very high surface area that resulted into high adsorption efficiency for both the adsorbents. Moreover, an important characteristic of CNTs is that its adsorption capacity and selectivity can be tailored by alteration of polarity and hydrogen bonding potential through introduction of different functional groups on their surface. The utilization of CNTs for the treatment of water and wastewater containing divalent metal ions is reviewed and future research work on development of cost-effective ways of CNT production is recommended (Rao et al. 2007). Though few studies report the recovery of adsorbed metal ions from the surface for reuse of the CNT, detailed information regarding regeneration and reuse of CNT in water and wastewater purification is limited. In a recent study by Vijwani et al. (2015), carbon nanotube attached porous carbon foams have been synthesized for adsorption of dyes from water, and the reusability of the adsorbent materials without loss of activity has also been demonstrated (Fig. 3.4). In addition to magnetic materials and CNTs, many hybrid materials such as alumina supported carbon nanotubes (Al2O3/CNT), Ce(IV) doped iron oxide, Ce impregnated

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Fig. 3.4 CNT attached porous carbon foam as reusable adsorbent (Vijwani et al. 2015)

Fig. 3.5 (a) Schematic diagram of the synthesis mechanism of silver nanoparticle loaded nanocomposites and (b) actual photograph of the water treatment device (Sankar et al. 2013)

fibrous protein and zirconium (IV) oxide-ethanolamine (ZrO-EA) are tested for water purification by adsorption (Wang et al. 2002, 2017; Deng and Yu 2012). Inorganic metal nanoparticles like silver and gold nanoparticle have attracted considerable attraction in the field of water treatment because of their characteristic antimicrobial property. Sankar et al. (2013) reported facile synthesis of a unique family of silver nanoparticle loaded nanocrystalline metal oxyhydroxide-chitosan composite materials and their application for effective removal of bacteria, arsenic,

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iron and lead from water. Successful application of the nanocomposites (Fig. 3.5) is also demonstrated by Sankar et al. (2013). To evaluate the efficiency of nanomaterials, one of the main limitations is the absence of a unified procedure that enables direct comparison of results. Moreover, these advanced materials are usually expensive, and hence find limited application in real-life water purification systems. Cost-effective synthesis methods and regeneration techniques for reuse of these advanced materials should be developed and tested for their real-life application.

3.3

Modes of Operation of Adsorption Process

Mode of operation of any water purification process is important in terms of their application at the point of use, in household purification system or in community level water treatment plants. Purification of water by adsorption is usually performed in two different modes.

3.3.1

Batch Adsorption Process

The efficiency of an adsorbent towards a specific contaminant is screened by batch process. It is the intuitive way of testing the potential of an adsorbent for a specific ion or molecule. Here a fixed amount of the adsorbent is contacted with a fixed volume of the adsorbate solution of known concentration at constant temperature. The change in concentration of the adsorbate after a fixed time interval gives the adsorption capacity of the adsorbent for the selected adsorbate ion or molecule. This is the first step to design the process of purification by adsorption.

3.3.2

Continuous Flow Adsorption Process

In this process, the feed water is percolated through a fixed bed column packed with the adsorbent in a continuous flow mode. The column is removed and regenerated on near-saturation of the adsorbent, as evidenced by breakthrough curve. Though the fixed bed continuous flow process can be operated either up-flow or down-flow direction, the up-flow mode has many advantages over the down-flow mode. The up-flow mode minimizes pressure drop, channelling and fouling of the adsorbent. In addition, adsorbent with relatively smaller particle size can be used in the up-flow mode of operation that led to the increase in adsorption rate and hence, a decrease in adsorber size. This fixed bed continuous flow adsorption process may be adopted for water purification at the point of use or in community level water treatment plants.

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61

Adsorption-Based Techniques Available for Drinking Water Purification

Iron oxy-hydroxides and zero-valent iron (raw iron in the form of filings, nails, balls, etc.) are the most commonly used adsorbent for arsenic removal from water, whereas activated alumina technology is one of the widely used methods for defluoridation of water. Reports are available on design and installation of arsenic removal and defluoridation units for purification of contaminated water in various levels of requirements (household level/community-level/large-scale). Sono filter and Kanchan™ filter are two such filters developed for arsenic remediation in household level and installed, respectively, in Bangladesh and Nepal. The Sono filter is a two-stage filtering system in which the arsenic is removed in the first stage by adsorption on iron hydroxides (Hussam et al. 2008). The residual iron, pathogen and other impurities are removed in the second stage of filtration. The Kanchan™ filter was developed by Massachusetts Institute of Technology (MIT), Environment and Public Health Organisation (ENPHO) and the Rural Water Supply and Sanitation Program (RWSSSP) of Nepal (Ngai et al. 2006). It is based on the principle of slow sand filtration and adsorption on iron hydroxide. The Kanchan™ filter is reported to be efficient in removing arsenic, pathogens, iron, turbidity, odour and some other contaminants from drinking water. An innovative idea of treatment of arsenic contaminated water by adding adsorbent in the form of tablets was developed by Chakraborti and his research group, School of Environmental Studies, Jadavpur University, Kolkata, India (SOES Report n.d.). In this process, the adsorbent tablet (made from Fe3+ salt, an oxidizing agent and activated charcoal) is added to the arsenic contaminated water and allowed to settle for few hours. Afterwards, the water is filtered through a filter candle indigenously made from fly ash, clay and charcoal, to get arsenic-free clean water. The above process is reported to remove 93–100% of arsenic present in water. Domestic water filter units for defluoridation of water are also developed and commissioned by different research groups at various places worldwide. Most of these filter units are based on the principle of sorption on activated alumina (Venkobachar et al. 1997). Investigations on activated alumina based domestic defluoridation units (DDU) have been reported by Chauhan et al. (2007). The exhausted activated alumina was reported to be regenerated and reused in DDU for five cycles without much loss in efficiency (Chauhan et al. 2007). The domestic filter units described above are reported to perform well at the removal viewpoint; however, regeneration or disposal of the adsorbent saturated with toxic contaminant remained a critical issue. Bone char based domestic and community defluoridizer units are tested in Rift Valley of Kenya, Africa (Mavura and Bailey 2002). The filter bed is made with locally made charred, crushes bones and is reported to be highly efficient (between 97.4 and 99.8%) in removing fluoride from water. The regeneration of the filter bed by eluting the filter with a mild acid or alkali solution has been reported by Mavura and Bailey (2002). Though the bone char based filter is low-cost and performs well,

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there is no report on its wide scale implementation. The method is not universally accepted. The major objection to the use of bone charcoal defluoridation is related to religious belief of some societies and communities. Community level water filter units are designed by attaching the filter bed with hand pump. AMAL arsenic filter is such a community level filter where activated alumina adsorption is used for on-spot arsenic removal from groundwater. It is developed by Bengal Engineering and Science University, Shibpur (BESUS), Kolkata, India (Lambert 2008), for on-spot arsenic removal from groundwater. About 125 filter units have been installed and running successfully in West Bengal, India. Each hand pump (nonelectric) is claimed to produce about 6000 litres of arsenic safe water per day. A similar community water filter unit (Hand pump attached model) based on activated alumina adsorption is developed for on-spot defluoridation of groundwater by a research group from Indian Institute of Technology (IIT), Kanpur (Iyenger 2005), with the support of UNICEF. Many government and non-government organizations are working worldwide for drinking water remediation. Large numbers of technologies are developed and being developed for drinking water treatment in household level, community-level and in large-scale plants; nevertheless, more work ought to be done to meet the requirement of safe drinking water specifically in rural areas.

3.5

Summary and Conclusion

Purification of water is essential to make the water suitable for human consumption. A number of water purification technologies exist; however, the selection of appropriate water treatment technique primarily depends on the nature of contaminant to be removed and cost-effectiveness of the technique. Since each purification technology removes a specific type of contaminant, it is impossible to get any universal technology for all types of water treatment. Many of the water purification plants use combination of treatment technologies to achieve desired water quality. The number of treatment techniques and combinations of techniques developed is expected to increase with time as more and more complex contaminants are discovered and come under regulation. An enormous number of research articles from all over the world on purification of water by adsorption using various types or groups of adsorbents indicate the necessity of development in this field of research from the environmental viewpoint and for human sustainability. The adsorption technology is growing at a tremendous pace and will continue to grow because of its environmental and industrial importance. Development of adsorption technology depends on the development of promising adsorbent materials for future applications. An adsorbent having the characteristics of high sorption capacity, easy separation from aqueous solution, low cost and regenerability for reuse may be considered as a promising material. The high selectivity of the adsorbent for the interested contaminant is most important in the removal of special ion or molecule from large volumes of aqueous solutions.

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Though non-conventional low-cost adsorbents are gaining interest, the traditional adsorbents such as activated carbon, activated alumina, metal oxides and hydroxides, natural ores and minerals still find tremendous applications in water purification. Biomaterial such as chitosan, alginate, cellulose and some agricultural byproducts (rice husk, bagasse, fruit peels/pulps) are among natural low-cost adsorbents that have attracted considerable attention of the researchers for drinking water purification. With the evolution of nanotechnology since past two decades, use of nanoadsorbents in water treatment is considered as an important achievement since the small size, high surface area and high effective surface contact/reactivity leads to high uptake capacity of this class of materials. Though the nano-adsorbents are highly efficient in removing pollutants from water however, the foremost drawbacks for nanoparticle use in water treatment are economical and safety constrains which complicate the replacement of conventional adsorbents. Furthermore, there is still requirement of technology advancement in the field of process development for practical application and usability of nano-adsorbents in water treatment.

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Chapter 4

Effect of Nano-TiO2 Particles on Mechanical Properties of Hydrothermal Aged Glass Fiber Reinforced Polymer Composites Ramesh Kumar Nayak

Contents 4.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2 Nano-TiO2 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3 Materials and Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.4 Void Content . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.5 Water Diffusion Kinetics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.6 Flexural Strength . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.7 Weibull Model Validation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.8 Interlaminar Shear Strength (ILSS) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.9 Post-failure Analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.10 Glass Transition Temperature . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.11 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

4.1

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Fibrous reinforced polymer composites are most assuring material due to its high specific strength, stiffness, and lightweight as compared to traditional metallic materials. Therefore, these materials are gradually replacing metallic material not only in high-end engineering applications like aerospace, navy, defense, and automotive but also in civil infrastructure and home appliances (Mangalgiri 1999). Due to the highly corrosive nature of metals used in oil and gas pipelines and failure rate of conventional materials, the importance of nanocomposites cannot be overemphasized; this is

R. K. Nayak (*) School of Mechanical Engineering, KIIT, Deemed to be University, Bhubaneswar, Odisha, India © Springer Nature Switzerland AG 2019 R. Prasad, K. Thirugnanasambandham (eds.), Advanced Research in Nanosciences for Water Technology, Nanotechnology in the Life Sciences, https://doi.org/10.1007/978-3-030-02381-2_4

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because nanocomposites are materials with a nanoscale structure that improve the macroscopic properties of products. In mechanical terms, nanocomposites differ from conventional composite materials due to the exceptionally high surface to volume ratio of the reinforcing phase and/or its exceptionally high aspect ratio. The reinforcing material can be made up of particles (e.g., minerals), sheets (e.g., exfoliated clay stacks), or fibers (e.g., carbon nanotubes or electrospun fibers) (Okpala 2014). However, fiber is stronger than polymer matrix, and it plays an important role to bear the load along the fiber direction resulting in FRP composites having superior mechanical properties in an in-plane direction. However, epoxy is a thermosetting polymer and brittle, resulting in poor resistance to initiation of crack and its propagation, which limits the application where high strength and fracture toughness is required. Nevertheless, in out-of-plane directions, mechanical properties depend on the fiber/matrix interface/interphase adhesive bond (Godara et al. 2010). Therefore, mechanical properties of FRP composites in out-of-plane direction are weaker than in-plane direction. This is attributed to the formation of the crack in the matrix at low strain and unable to transfer the load to the fiber. Thus, several attempts have been made towards the enhancement of mechanical properties by modifying the epoxy matrix; as a result matrix toughness and interface strength improved (Gojny et al. 2005; Iqbal et al. 2009; Böger et al. 2010; Kinloch et al. 2007). In the last decades, the majority of the research focused on the development of high-performance composites material based on inorganic fillers with organic epoxy, and key emphasizes on better integration between inorganic particles with organic polymers. Researchers have reported that the addition of inorganic nanofillers like Al2O3, SiO2, CaCO3, and nano-TiO2 into the epoxy matrix enhances the physical, mechanical, thermal, and electrical properties (Zhai et al. 2006; Bauer et al. 2006; Zhang et al. 2006; Zunjarrao and Singh 2006; Yu et al. 2006; Li et al. 2007). The enhancement of mechanical properties is fundamentally determined by the nanoparticle size, shape, distribution, and interaction with the organic polymer. The blending of inorganic nanofillers/particles into the epoxy polymer matrix demonstrates remarkable physical and mechanical properties due to the large surface area per unit volume of nanoparticles. The enhancement of properties is attributed to the interaction of nanoparticles with polymer matrix at the interface/interphase, and the physiochemical interaction governed by the nanoparticles surfaces (Luo and Daniel 2003; Pavlidou and Papaspyrides 2008). Scientists and researchers have observed that nano inorganic fillers are most promising fillers, especially metal oxides, to enhance the mechanical and thermal properties of the polymer matrix (Hayashi et al. 2005; Hong et al. 2005). This is because high specific surface area per unit volume of nanofillers helps for better interfacial interaction between polymer and nanoparticles, resulting in better mechanical and thermal properties of the composites (Carballeira and Haupert 2010; Ghosh and Nukala 2008). However, the high specific surface area of nanoparticle attracts each other at a higher content of it. This is due to electrostatic van der Waals forces between nanoparticles resulting in segregation of nanoparticles. Mechanical, thermal, and corrosion properties of fiber reinforced polymer composites are deteriorated under different environmental conditions. This is because, in

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Fig. 4.1 Illustration of damage development in the multiscale glass fabric/epoxy/CNT–Al2O3 composites: (a) matrix-dominated cracking; (b) transverse cracks and fiber/matrix debonding, and (c) delamination

hydrothermal condition, polymer matrix composites absorb water through diffusion mechanism. Diffused water in the GFRP composites ends up either in the matrix or at the fiber matrix interphase. The amount of water present in the matrix must be different than the interface which results in the mismatch of volumetric expansion between matrix and interface/interphase in hydrothermal condition. In the above said condition, the formation of localized stress and strain is favorable in the composites. Further, absorbed water is transmitted to other location of the composites through capillary action at the fiber/matrix interface. The absorbed moisture/water molecules bonded with the hydroxyl group of epoxy and free water clustered in the free volume/voids present in the matrix or at the fiber/matrix interface (Gonon et al. 2001; Maggana and Pissis 1999). Absorbed moisture helps the epoxy matrix to expand/swell and plasticize; hydrolysis and formation of cracks enhance the diffusion of water into the composites resulting in deterioration of mechanical and thermal properties. The differential thermal expansion and/or contraction of polymer matrix and fiber may also degrade the interface strength resulting in the decrease in mechanical properties. Hence, these materials are facing challenges and threats in different environments like high and low temperature, water, hydrothermal, alkaline, corrosive, and UV light exposure. One of the probable methods to improve the interface strength is by adding nanofillers into GFRP composites (Fan et al. 2009; Dikobe and Luyt 2010; Jongsomjit et al. 2004). Figure 4.1 shows the illustration of damage developed in the multiscale glass fabric/epoxy/CNT-Al2O3 nanocomposites (Li et al. 2014). The microcrack propagation can be hindered through nanofillers. This is due to the presence of high toughness of nanofillers in the polymer matrix. The crack propagation hinders, bypass the nanoparticles, or stop at the nanoparticles due to the high toughness of the nanofillers resulting in improvement of overall strength of the composites. The addition of inorganic or metal oxide nanofillers into the epoxy matrix is the promising method to enhance the thermal and mechanical properties in dry and

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hydrothermal conditions of GFRP composites. Although water absorption study is not surprising for GFRP composites, investigation of residual mechanical properties of nano-Al2O3/TiO2/(Al2O3 + TiO2) filled glass fiber reinforced epoxy polymer composites subjected to accelerated hydrothermal aging is quite exciting and necessary. Hence the following topic has emerged. 1. Fabrication of nanocomposites by adding nano-TiO2 filler at different concentrations into the epoxy polymer matrix. 2. Evaluation of water absorption kinetics, flexural strength, interlaminar shear strength, and glass transition temperature of nano-TiO2 enhanced GFRP composites subjected to hydrothermal conditioning. 3. In-service assessment of durability regarding water absorption kinetics, mechanical and glass transition temperature of nano-TiO2 enhanced glass fiber reinforced polymer composites (GFRP). Primarily it has been focused on the addition of nano-TiO2 into the epoxy polymer matrix on water absorption, flexural strength, interlaminar shear properties, and glass transition temperature of GFRP composites. Effect of accelerating hydrothermal aging of these nanocomposites on residual mechanical and thermal properties has been investigated. Weibull design parameters were determined for dry and hydrothermally conditioned composites. Experimental and Weibull simulated stress-strain results are compared. Nano and microscale strengthening mechanism and failure modes were analyzed through the fractographic study of FESEM images. Overall, the potential of nano-TiO2 reinforcement into the epoxy polymer matrix on environmental durability and reliability is analyzed and discussed.

4.2

Nano-TiO2

Nano-TiO2 particle is one of the promising inorganic nanofillers used in polymer matrix composites to enhance the mechanical properties. However, the reliability of this type of nanocomposites is yet to be ensured in the hydrothermal environment. The present work investigates the addition of nano-TiO2 filler on water sorption, residual strength, and thermal properties of glass fiber reinforced polymer (GFRP) composites. The results revealed that addition of 0.1 wt% TiO2 has reduced water diffusion coefficient by 9%, and improved residual flexural strength by 19% and residual interlaminar shear strength by 18% among all the nano-TiO2 modified composites. The improvement of mechanical properties in the hydrothermal environment creates opportunity and reliability to be used in different engineering applications. Weibull design parameters are evaluated and found a good agreement between Weibull stress-strain curves and experimental one. The fractographic analysis confirmed the various failures and strengthening mechanisms of nanocomposites in the dry and hydrothermal environment.

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4.3

73

Materials and Methods

Glass fiber (GF) reinforced polymer composites (GFRP) are fabricated using Diglycidyl ether of Bisphenol A (DGEBA) type of epoxy, Triethylene tetra amine (TETA) as hardener supplied by Atul Industries, India, and woven roving E-glass fiber procured from Owens Corning, India. Nano GF composites were made of nanoTiO2 fillers, epoxy, hardener, and woven fabric glass fiber. The TiO2 nanofillers were supplied by SRL Industries limited, India. Some of the essential properties of epoxy, E-glass fiber, and nanoparticle are reported in Table 4.1 (Nayak et al. 2016). It is observed that nano-TiO2 particle has superior mechanical properties compared to neat epoxy and it is expected that the nanofiller may enhance mechanical/thermal properties of the epoxy matrix and matrix/fiber interface. Nano-TiO2 is dried at 100
C before it is mixed with epoxy to remove moisture in it. Control GF composites were fabricated without, and nano GF composites are fabricated with different wt% of nano-TiO2 fillers. The fraction of fiber and epoxy is maintained at 60:40 ratios by weight for both control GF and nano GF composites during fabrication. As per the suppliers’ instruction 10% of epoxy of hardener is used for the curing process. Researchers have reported that mechanical stirring followed by sonication is an excellent method to disperse nanofillers in the polymer matrix (Chang and Chow 2010; Soundararajah et al. 2009; Barbezat et al. 2009). In this work instead of mechanical stirring rapid magnetic stirring is adopted and followed by sonication shown in Fig. 4.2. This is because in mechanical stirring the probability of formation of bubbles is more and increases the void content. It was expected that magnetic stirring might reduce bubbles formation and reduce the void content. The magnetic stirrer stirs the epoxy, nano-TiO2 at different wt% for 1 h followed by sonication at 60
C for another 45 min to disperse nanoparticles in the epoxy matrix. It is expected that the shear force developed between the nanoparticles helps the process of de-agglomeration of nanoparticles. Composite laminates were fabricated with 16 layers of woven fabric glass fiber by hand lay-up techniques followed by temperature assisted compression molding (pressure 10 kg/cm2) and at a temperature of 60
C for 20 min. Further curing of composites is done at 140
C for 6 h before characterization. Figure 4.3 shows the schematic diagram of the fabrication method. Samples of different sizes are cut as per ASTM standard using diamond coated tipped cutter for further characterization.

Table 4.1 Properties of raw materials

Properties Density (g/cm3) Tensile strength (MPa) Tensile modulus (GPa) Poisson’s ratio

Epoxy 1.15 70 3.6 0.30

TiO2 (rutile) 4.00 51.6 228 0.27

Glass fiber 2.58 3800 78 0.20

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Fig. 4.2 Equipment used for fabrication of nano GFRP composites

4.4

Void Content

Voids are nothing but closed pores present in the composites, which plays the initial absorption of water into the composites and resulting in reduction in mechanical properties. Void content of the GFRP composites is determined by resin burn off test. The fiber weight fraction and void content of each laminate were determined as per ASTM D 3171-99. As per the standard, there are six no. of samples each of size 25 mm  25 mm of surface area considered for this analysis. Initial weight and dimensions of the samples are measured through high accuracy weighing balance and digital vernier caliper, respectively. Samples are put inside the muffle furnace at 575  10
C for 5 h to burn off the epoxy. The mathematical expression used to determine the void content is given in Eq. (4.1) (Chau et al. 2008).

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Fig. 4.3 Fabrication method of nano-TiO2 enhanced glass fiber reinforced polymer composites

Fig. 4.4 (a) Fiber weight percentage (wt%) and (b) void content versus TiO2 content (wt%)


 w f wm V v ¼ 1  ρc þ ρ f ρm

ð4:1Þ

where Vv is the volume fraction of void, ρc is the density of composite, ρf is the density of fiber, ρm is the density of epoxy, wf weight fraction of fiber, and wm weight fraction of epoxy. Figure 4.4a, b shows the effect of wt% of nano-TiO2 on fiber wt% and void content, respectively. It is observed that fiber weight (%) is around 60–62% for all types of composites. Void content increases with increase in nano-TiO2 as

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compared to control GF composites. This is because the tendency to form bubbles are more with the increase in nanofillers content and the entrapped gases bubbles are unable to come out from viscous epoxy/TiO2 suspension during the fabrication process. However, further reduction of entrapped bubbles might be possible by high vacuum desiccators.

4.5

Water Diffusion Kinetics

Absorption of moisture has the significant impact on degradation of mechanical properties of FRP composites. Water diffuses into the composites ends up either in the matrix or at the fiber matrix interphase. The amount of water present in the matrix must be different than the interphase. As a result, a mismatch of volumetric expansion between matrix and fiber interphase leads to the formation of localized stress and strain in the composites. Further, absorbed water is transmitted to other locations of the composites through capillary action at the fiber/matrix interface. Absorbed moisture helps the epoxy matrix to expand/swell and plasticize; hydrolysis and formation of cracks enhance the diffusion of water in the composites. Hence, moisture reduces mechanical performance, lowering Tg, viscosity, and elasticity of polymer (Park et al. 2007; Won et al. 2012). Control and nano GF composite samples were correctly dried at 75
C for 5 h and weighed in a high precision weighing machine (“Sartorius “having an accuracy of 0.01 mg) before hydrothermal conditioning. Nanocomposites samples were immersed in a temperature controlled water bath at 70
C for 30 days shown in Fig. 4.5. The pH of water is measured and found around 5.65. The experiments were conducted as per the ASTM D 570-98 standard. Specimens were removed from the water bath at regular interval of time. The adhered water on the surface of composites is wiped off using dry cotton cloth and weighed with high precision weighing balance. The time required for

Fig. 4.5 Temperature controlled water bath used for hydrothermal aging

Moisture content (%)

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Fickian Diffusion

Mm M2

M1 √t1

√t2

Square root of time (min)

Fig. 4.6 Typical water absorption behavior of FRP composite

removal of samples from the water bath and until finishing weight measurement is very short and assuming no desorption of water takes place during this time period. This procedure was repeated in a regular interval of time for 30 days. The amount of water absorbed by the composite (Mt) is calculated using the given Eq. (4.2). Mt ¼

m  m0  100% m0

ð4:2Þ

where Mt is the percentage of moisture content at time t, m0 is the weight of the specimen at its dry state, and m is the specimen weight at time t. A typical water absorption behavior of GFRP composite is shown in Fig. 4.6. It is observed that initially the water absorbed by the polymer is linear and follows Fick’s law of diffusion. Then absorption nearly saturated and became ceased. Further, it increases because of degradation of epoxy and moisture penetrated through the microcracks and capillary action. The water/ moisture diffusion coefficient is calculated in Fickian diffusion range using Eq. (4.3) (Silva et al. 2014; Pervin et al. 2005).  Dz ¼ π

h 4  ð% M m Þ



2 

 ð% M 2  % M 1 Þ 2 pffiffiffiffi pffiffiffiffi t2  t1

ð4:3Þ

where M1 and M2 are the percentages of water absorption at time t1 and t2. Mm is the maximum moisture content in Fickian reason, and Dz is the diffusion coefficient. Dz is one dimensional, which does not account for diffusion taking place through the edge. Rao et al. (1984) suggested the corrected diffusion constant D, which is expressed in Eq. (4.4).

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D¼

Dz 1 þ hl þ wh

2

ð4:4Þ

where h is the thickness, w is the width, and l is the length of the sample. Figure 4.7a shows the water absorption behavior of nanocomposites at different wt% TiO2 (0.1–0.7). It is observed that impingement of nano-TiO2 into the epoxy matrix reduces water absorption tendency as compared to control GF composites. However, with the increase in wt% of TiO2, absorption of water increases and this may be because of more no. of entrapped voids present in the composites. The decrease in water absorption of nano GF composites compared to control GF is attributed to the formation of good interface bond between matrix and nanofillers that reduces the tendency to absorb moisture through capillary action. Figure 4.7b shows the linear portion of the water diffusion behavior. It is reasonable to assume that in the linear portion of absorbed water wt% versus square toot of time plot follows Fickian diffusion. Water diffusion coefficient is calculated for control GF and nano GF composites using the Eqs. (4.3) and (4.4). It is observed that diffusion coefficient reduces about 9% with nanofillers impingement as compared to control GF composites shown in Fig. 4.7c.

Fig. 4.7 Water absorption curves for control GF and nanocomposites, (a) effect of wt% of TiO2, (b) Fickian diffusion region, and (c) diffusion coefficient versus TiO2 content (wt%)

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79

Flexural Strength

Durability and reliability of nano GFRP composites are very important in different engineering applications. Retainability of its interface strength and toughness is necessary for hydrothermally conditioned GFRP composites. Interface and matrix strength of the GFRP composites can be tailored through the flexural test. A Universal Testing Machine (UTM) of INSTRON 5967 is used to evaluate the flexural strength as per ASTM D7264 standard at room temperature for dry and hydrothermally conditioned control and nano GF composites shown in Fig. 4.8. The flexural strength (σ F), modulus (EF), and strain to failure (εF) are calculated as per the equation given below (Dong and Davies 2012): σF ¼

3Pmax L 2wt 2

when

mL3 4wt 3 6dt εF ¼ 2 L

EF ¼

Fig. 4.8 INSTRON 5967 used for flexural strength

L  16 t

ð4:5Þ ð4:6Þ ð4:7Þ

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Fig. 4.9 Effect of TiO2 content (wt%) on (a) flexural strength, (b) strain at peak load, and (c) modulus of dry and hydrothermally conditioned samples

where L is the span length of the sample, w is the width and t is the thickness of the sample, Pmax is the maximum load applied before failure, m is slop of the initial portion of load versus displacement curve, d is the maximum bending before failure. Effect of nano-TiO2 content on flexural strength, strain, and modulus has been evaluated and shown in Fig. 4.9. For dry samples, it is observed that maximum strength is achieved at 0.1 wt% of TiO2 and the further increase in TiO2 content, strength and strain decreases. However, modulus of the nanocomposites is increased with increase in wt% of TiO2. Therefore, in dry condition improvement of flexural strength is 18% maximum at 0.1 wt% TiO2 and modulus by 19% at 0.7 wt% of TiO2 as compared to control GF composite. Initial improvement of strength is attributed to the good interaction between epoxy and nanoparticles that produces better interface bond with the fiber and enhances the strength and toughness of matrix (Chatterjee and Islam 2008). With the increase in TiO2 content also increases the van der Waals’ force between the nanoparticles, resulting in decrease of the dispersion of nanoparticles in the epoxy matrix. The decrease in strength may be attributed to agglomeration of nanoparticles resulting in the reduction of the active surface area of nanoparticles to interact with

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epoxy matrix. Further, it reduces the load transfer from matrix to fiber, resulting in the decrease in flexural strength. In hydrothermally conditioned samples, flexural strength, stain, and modulus follow the similar trend like dry samples. However, there is a reduction in strength and modulus of hydrothermally conditioned nanocomposites samples. The decrease in strength of hydrothermally treated composites are because of (i) interfacial debonding developed by differential swelling between matrix, nanofillers, and glass fiber at 70
C and (ii) osmotic cracking at the interphase and hydrogen bond breakup (hydrolysis) in the presence of water which leads to decrease in adhesion (Gautier et al. 1999; Hodzic et al. 2004; Ellyin and Maser 2004). In summary, GFRP composite life period is highly dependent on the ability of the epoxy matrix or interface to absorb moisture in different service conditions. Figure 4.10 shows the residual strength, strain, and modulus versus TiO2 content (wt%). The maximum improvement of residual flexural strength is about 18% and modulus by 22% as compared to control GF composites. The increase in strength, strain, and modulus of nano GF composites compared to control GF composites in hydrothermally conditioned samples is attributed to reduction in water absorption, better interface/interphase strength between matrix and glass fiber, and improvement

Fig. 4.10 Effect of TiO2 content (wt%) on (a) residual flexural strength, (b) strain, and (c) modulus

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of matrix strength, and propagation of microcrack is possibly hindered by nano-TiO2 particles resulting in increase in strength and modulus of the composites.

4.7

Weibull Model Validation

Fiber reinforced polymer composites are used in different structural applications. Keeping in view for reliable structural design, prediction of the mechanical properties with high accuracy is necessary. In general, FRP composites are anisotropic, and the mode of failure depends on the types of reinforced fiber, matrix and fiber/matrix interface strength. During deformation, each constituent of the composite behaves differently either independently or combinedly depending upon the flow behavior of matrix. Statistical variation of mechanical performance can be addressed through Weibull probability distribution function to predict the mechanical properties of nano-TiO2 filled glass fiber reinforced polymer composites. Weibull simulated stress (σ) ~ strain (ε) relationship can be expressed as per Eq. (4.8) (Pervin et al. 2005; Lau et al. 2013). "
 # Eε β σ ¼ Eεexp  σ0

ð4:8Þ

where E is the elastic modulus as per the applied loading direction and σ 0 and β are the Weibull design parameters. The physical significance of the Weibull design parameters of σ 0 and β are nominal strength and extent of randomness in the performance of the material, respectively. It means, with the increase in the value of σ 0, increases the nominal strength and with the increase in β, reduces the scatter of the performance. The design parameters can be determined by taking double logarithm on both sides and rearrangement of Eq. (4.8) and expressed in Eq. (4.9). By invoking the experimental data of E, σ, and ε, a straight line can be drawn, where the y-axis becomes ln[ln(Eε/σ)], and x-axis become ln(Eε). The slope of the straight line becomes β and intercept β ln(σ 0). From the value of β and intercept β ln(σ 0), σ 0 can be calculated. 
 Eε ¼ β ln ðEεÞ  β ln ðσ 0 Þ ln ln σ

ð4:9Þ

In this section, Weibull design parameters are evaluated for dry and hydrothermally conditioned control GF and nanocomposites using flexural data. Further, the comparison between the experimental stress-strain curve and Weibull simulated one is carried out. Figure 4.11 shows the linear fitting of experimental results to determine the Weibull design parameters for dry and hydrothermally conditioned samples. The design parameters are reported in Table 4.2. In dry and hydrothermally conditioned samples, the σ 0 and β values follow the trend of flexural strength and

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Fig. 4.11 Weibull linear fittings of experimental data for (a), (b) dry samples, (c) and (d) hydrothermally conditioned samples Table 4.2 Weibull scale (σ 0) and shape (β) parameters for control and nano GF composite TiO2 content (wt%) 0 0.1 0.3 0.7

Dry condition σ0 756.6  14.0 1177.3  52.4 732.7  32.2 678.6  25.1

β 1.7  0.20 1.85  0.15 1.67  0.03 1.59  0.04

Hydrothermally conditioned σ0 Β 571.3  19.5 2.36  0.03 703.3  36.5 2.37  0.05 548.67  19.0 2.51  0.09 440.3  9.5 3.48  0.07

strain except for shape parameters (β) increases with increase in wt% TiO2 in hydrothermally conditioned samples. Using Weibull design parameters, simulated stress and strain values are calculated. Figure 4.12 shows the comparison between experimental and simulated stress-strain curves for dry and hydrothermally conditioned samples. A close agreement between experimental and simulated has been observed. Hence it may be concluded that the experimental data is reliable.

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Fig. 4.12 Comparison between experimental and Weibull predicted simulated stress-strain curve for dry and hydrothermally conditioned samples, (a) control GF(0.0 wt% TiO2), (b) 0.1 wt% TiO2, (c) 0.3 wt% TiO2, and (d) 0.7 wt% TiO2

4.8

Interlaminar Shear Strength (ILSS)

Interlaminar shear strength indicates the degree of adhesiveness between fiber and matrix of nanocomposites. Generally, nanofillers are added to increase the surface area at the interface of fiber and matrix to enhance the interface strength. The ILSS tailored through short beam shear test. Short beam shear (SBS) test of the composites was evaluated as per the ASTM: D2344-13 standard. The sample size used for the test is 37  9  4.5 mm3 and span length 27 mm. ILSS is evaluated using Instron5967 UTM machine at different cross head speed by adjusting the crosshead velocity. In SBS test, the sample placed in between two supporting rollers of 3 mm diameter and forces were applied at the middle of the sample through another 6 mm diameter roller shown in Fig. 4.13. The crosshead speed varied through machine crosshead speed. The specimens were tested at room temperature and different crosshead speeds. The ILSS values were calculated as per Eq. (4.10).

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Fig. 4.13 Shows mode II in-plane shear during short beam shear (SBS) test of the composites

ILSS ¼

3Pmax 4wt

ð4:10Þ

Figure 4.14a, b indicates the effect of nano-TiO2 on ILSS between dry and hydrothermally conditioned samples and residual ILSS. At 0.1 wt% TiO2, ILSS has improved by about 19% in dry and about 18% in hydrothermally conditioned samples. The residual strength improved with the increase in wt% of TiO2 and the maximum value was observed at 0.3 wt% TiO2. However, ILSS reduces in hydrothermally conditioned samples compared to dry samples. The decrease in ILSS of hydrothermally conditioned samples is attributed to microcrack formation in the matrix, at the interphase and interfacial debonding. The microcrack formation at the interphase is due to differential swelling of matrix and fiber because of water absorption and thermal gradient during hydrothermal conditioning. However, nanoparticles reduce the water absorption tendency compared to control GF composites resulting in improvement of ILSS. Therefore, the self-healing capacity of nanocomposites has improved as compared to control GF composites.

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Fig. 4.14 (a) Interlaminar shear strength in dry and hydrothermally conditioned samples (b) Residual interlaminar shear strength after hydrothermally conditioned samples

4.9

Post-failure Analysis

Fractured surface features are investigated through field emission scanning electron microscope to understand the various micro and nanoscale failure mechanisms that occurred during flexural and interlaminar shear test. Effect of nano-TiO2 on strengthening mechanism has also been investigated. Figure 4.15 shows the fracture surface features of control GF and nanocomposites up to 0.3 wt% TiO2 in dry condition. The FESEM images revealed that there is a crack bridging by nano-TiO2 particles by which the flexural and ILSS increase in nanocomposites as compared to control GF composites. Figure 4.16 shows fracture surface features of nanocomposites (0.7 wt% TiO2) in dry condition. The strengthening mechanism of nanocomposites is the combination of highly oriented shear cusps, matrix deformation, and crack bridging through nano-TiO2 particles. Figure 4.17 shows the comparison of fiber imprints of dry and hydrothermally conditioned control and nano GF composites. The fiber imprints along with matrix deformation morphology of nanocomposites compared to control GF. It implies that nano-TiO2 particles improved the matrix as well as the interface strength between fiber and matrix resulting in improvement of ILSS, Fexural strength, and modulus in hydrothermally conditioned nanocomposites. However, in dry samples, fiber imprints have matrix deformation, river line marks, hackles (Kim et al. 2008), and highly oriented shear cusps resulting in improvement of mechanical strength. The overall failure mode is consisting of fiber delamination, interface debonding, and brittle failure of fiber.

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Fig. 4.15 FESEM fracture surface features of dry samples, (a) interfacial debonding and matrix drain out in 0.0 wt% TiO2, (b) good interfacial bonding in 0.1 wt% TiO2, (c) nano-TiO2 pull out, and (d) crack bridging in 0.3 wt% TiO2

Fig. 4.16 FESEM fracture surface features of highly oriented shear cusps (a) matrix deformation (b) and crack bridging (c) in 0.7 wt% TiO2 content composites in dry condition

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Fig. 4.17 FESEM fracture surface features of fiber imprints in dry (a) 0.0 wt% TiO2, (b) 0.1 wt% TiO2, (c) 0.3 wt% TiO2, (d) 0.7 wt% TiO2 and hydrothermally conditioned (e) 0.0 wt% TiO2, (f) 0.1 wt% TiO2, (g) 0.3 wt% TiO2, (h) 0.7 wt% TiO2 composites

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4.10

89

Glass Transition Temperature

Thermal properties of GFRP composites are very critical regarding the sustainability and reliability of mechanical properties in in-service temperature. Differential scanning calorimetry (DSC-822, Mettler Toledo) is used to evaluate the glass transition temperature between control GF and GF nanocomposites shown in Fig. 4.18. The test is done under nitrogen atmosphere and at a heating rate of 10
C/min. Figure 4.19 shows heat flow versus temperature of control GF and nano GF composites at (a) dry and (b) hydrothermal condition. The remarkable change in glass transition temperature between control and nano GF composites in dry and hydrothermal condition. Figure 4.20a, b shows glass transition temperature (Tg) of the nanocomposites before and after hydrothermal treatment, respectively. Table 4.3 shows the glass transition temperature of control GF and nano GFRP composites at the dry and hydrothermal condition. In hydrothermal condition, it is t there is an improvement of Tg around 4
C at 0.7 wt% TiO2 filled nanocomposite as compared to control GF composite. The increase in Tg in hydrothermal condition attributed to the loss in mobility of chain segments of the epoxy system because of nanoparticles and epoxy interaction. The adhesion of nanoparticles with polymer also restricts the molecular motion up to some extent

Fig. 4.18 TMDSC equipment (Mettler Toledo) used for glass transition temperature

Fig. 4.19 Heat flow versus temperature of control GF and nano GF composite, (a) dry, (b) hydrothermally conditioned

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Fig. 4.20 Glass transition temperature of (a) dry and hydrothermally conditioned samples and (b) residual Tg of hydrothermally conditioned nanocomposites Table 4.3 Glass transition temperature (Tg) of control and nano GFRP composites at different environmental conditions Wt% TiO2 0.0 0.1 0.3 0.7

Dry condition Tg (
C) 122.27 123.50 123.71 123.94

Hydrothermal condition Tg (
C) 95.87 95.45 94.80 99.16

resulting enhancement of Tg. Hard nanoparticles are also acting as a virtual network node for the enhancement of Tg (Chatterjee and Islam 2008). However, the overall Tg of hydrothermally treated nanocomposites decreases around 25–30
C as compared to the dry condition. It means the thermal resistance of the nanocomposites has reduced in hydrothermal conditioning and its application beyond 95
C is not advisable. The decrease in glass transition temperature (Tg) is attributed to the hydrolysis of epoxy takes place during hydrothermal treatment lead to decrease the cross-linking density, resulting in the decrease in Tg.

4.11

Conclusions

The present investigation emphasizes the effect of TiO2 nanofiller content on water absorption and residual mechanical and thermal properties of nano GFRP composites. The following conclusions are drawn. • Temperature assisted magnetic stirring followed by sonication reasonably disperses the nano-TiO2 particles in the epoxy matrix. • The results revealed that the effect of the addition of 0.1 wt% nano-TiO2 reduces moisture diffusion coefficient about 9%, increases residual flexural strength by

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19%, interlaminar shear strength by 18% as compared to control GF composites. Interestingly, nanocomposites having 0.7 wt% TiO2 improves its residual flexural modulus about 22% in comparison to control GF composites. • The improvement of mechanical properties has been correlated with the fracture surface features of the nanocomposites. The fractured surface revealed the enhancement of mechanical properties is due to microcrack bridging through nano-TiO2, matrix toughening, and the good interfacial bond between matrix and fiber. However, the mode of failure is the combination of interface debonding, fiber pulls out, matrix cracking, matrix deformation, and fiber breakage. • Interestingly, glass transition temperature has not been improved substantially in dry and hydrothermally conditioned samples in comparison to control GF composites. • The reduction in water absorption and improvement of the residual mechanical properties of 0.1 wt% nano-TiO2 filled GFRP composites creates an opportunity to be used in the hydrothermal environment as compared to control GF composites.

References Barbezat M, Brunner AJ, Necola A, Rees M, Gasser P, Terrasi G (2009) Fracture behavior of GFRP laminates with nanocomposite epoxy resin matrix. J Compos Mater 43:959–976 Bauer F, Decker U, Ernst H, Findeisen M, Langguth H, Mehnert R (2006) Functionalized inorganic/organic nanocomposites as new basic raw materials for adhesives and sealants Part 2. Int J Adhes Adhes 26:567–570 Böger L, Sumfleth J, Hedemann H, Schulte K (2010) Improvement of fatigue life by incorporation of nanoparticles in glass fibre reinforced epoxy. Compos Part Appl Sci Manuf 41:1419–1424 Carballeira P, Haupert F (2010) Toughening effects of titanium dioxide nanoparticles on TiO2/ epoxy resin nanocomposites. Polym Compos 31:1241–1246 Chang LN, Chow WS (2010) Accelerated weathering on glass fiber/epoxy/organo-montmorillonite nanocomposites. J Compos Mater 44:1421–1434 Chatterjee A, Islam MS (2008) Fabrication and characterization of TiO2–epoxy nanocomposite. Mater Sci Eng A 487:574–585 Chau JLH, Tung C-T, Lin Y-M, Li A-K (2008) Preparation and optical properties of titania/epoxy nanocomposite coatings. Mater Lett 62:3416–3418 Dikobe DG, Luyt AS (2010) Comparative study of the morphology and properties of PP/LLDPE/ wood powder and MAPP/LLDPE/wood powder polymer blend composites. Express Polym Lett 4:729–741 Dong C, Davies IJ (2012) Optimal design for the flexural behaviour of glass and carbon fibre reinforced polymer hybrid composites. Mater Des 37:450–457 Ellyin F, Maser R (2004) Environmental effects on the mechanical properties of glass-fiber epoxy composite tubular specimens. Compos Sci Technol 64:1863–1874 Fan XJ, Lee SWR, Han Q (2009) Experimental investigations and model study of moisture behaviors in polymeric materials. Microelectron Reliab 49:861–871 Gautier L, Mortaigne B, Bellenger V (1999) Interface damage study of hydrothermally aged glassfibre-reinforced polyester composites. Compos Sci Technol 59:2329–2337 Ghosh PK, Nukala SK (2008) Properties of adhesive joint of inorganic nano-filler composite adhesive. Indian J Eng Mater Sci 15:68

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Godara A, Gorbatikh L, Kalinka G, Warrier A, Rochez O, Mezzo L (2010) Interfacial shear strength of a glass fiber/epoxy bonding in composites modified with carbon nanotubes. Compos Sci Technol 70:1346–1352 Gojny FH, Wichmann MH, Fiedler B, Bauhofer W, Schulte K (2005) Influence of nanomodification on the mechanical and electrical properties of conventional fiber-reinforced composites. Compos Part Appl Sci Manuf 36:1525–1535 Gonon P, Sylvestre A, Teysseyre J, Prior C (2001) Combined effects of humidity and thermal stress on the dielectric properties of epoxy-silica composites. Mater Sci Eng B 83:158–164 Hayashi K, Kurosaka Y, Osako Y, Ha J, Vacha M, Sato H (2005) Electrical properties of composites of TiO 2-triphenylamine derivatives. Thin Solid Films 474:337–340 Hodzic A, Kim JK, Lowe AE, Stachurski ZH (2004) The effects of water aging on the interphase region and interlaminar fracture toughness in polymer–glass composites. Compos Sci Technol 64:2185–2195 Hong JI, Winberg P, Schadler LS, Siegel RW (2005) Dielectric properties of zinc oxide/low density polyethylene nanocomposites. Mater Lett 59:473–476 Iqbal K, Khan S-U, Munir A, Kim J-K (2009) Impact damage resistance of CFRP with the nanoclay-filled epoxy matrix. Compos Sci Technol 69:1949–1957 Jongsomjit B, Chaichana E, Praserthdam P (2004) LLDPE/nano-silica composites synthesized via in situ polymerization of ethylene/1-hexene with MAO/metallocene catalyst. J Mater Sci 40:2043–2045 Kim H-Y, Park Y-H, You Y-J, Moon C-K (2008) Short-term durability test for GFRP rods under various environmental conditions. Compos Struct 83:37–47 Kinloch AJ, Masania K, Taylor AC, Sprenger S, Egan D (2007) The fracture of glass-fibrereinforced epoxy composites using nanoparticle-modified matrices. J Mater Sci 43:1151–1154 Lau K, Wong T, Leng J, Hui D, Rhee KY (2013) Property enhancement of polymer-based composites at cryogenic environment by using tailored carbon nanotubes. Compos Part B Eng 54:41–43 Li H, Zhang Z, Ma X, Hu M, Wang X, Fan P (2007) Synthesis and characterization of epoxy resin modified with nano-SiO2 and γ-glycidoxypropyltrimethoxy silane. Surf Coat Technol 201:5269–5272 Li W, He D, Dang Z, Bai J (2014) In situ damage sensing in the glass fabric reinforced epoxy composites containing CNT–Al2O3 hybrids. Compos Sci Technol 99:8–14 Luo J-J, Daniel IM (2003) Characterization and modeling of mechanical behavior of polymer/clay nanocomposites. Compos Sci Technol 63:1607–1616 Maggana C, Pissis P (1999) Water sorption and diffusion studies in an epoxy resin system. J Polym Sci Part B Polym Phys 37:1165–1182 Mangalgiri PD (1999) Composite materials for aerospace applications. Bull Mater Sci 22:657–664 Nayak RK, Mahato KK, Ray BC (2016) Water absorption behavior, mechanical and thermal properties of nano TiO2 enhanced glass fiber reinforced polymer composites. Compos Part A 90:736–747 Okpala CC (2014) The benefits and applications of nanocomposites. Int J Adv Engg Tech 6 (4):12–18 Park H-K, Su-Jin L, Yoon-Jeong K, Jang C-I, Won J-P (2007) Mechanical properties and microstructures of GFRP Rebar after long-term exposure to chemical environments. Polym Polym Compos 15:403–408 Pavlidou S, Papaspyrides CD (2008) A review on polymer-layered silicate nanocomposites. Prog Polym Sci 33:1119–1198 Pervin F, Zhou Y, Rangari VK, Jeelani S (2005) Testing and evaluation on the thermal and mechanical properties of carbon nano fiber reinforced SC-15 epoxy. Mater Sci Eng A 405:246–253 Rao RMVGK, Chanda M, Balasubramanian N (1984) Factors affecting moisture absorption in polymer composites part II: influence of external factors. J Reinf Plast Compos 3:246–253

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Chapter 5

Nanotechnology: An Innovative Way for Wastewater Treatment and Purification Muhammad Rafique, Muhammad Bilal Tahir, and Iqra Sadaf

Contents 5.1 Introduction: Water and Water Technology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2 Major Sources of Water Pollution . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2.1 Domestic Sewage . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2.2 Industrial Water Pollution . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2.3 Population Growth . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2.4 Pesticides and Fertilizers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2.5 Plastics and Polythene Bags . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2.6 Urbanization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2.7 Ground Pollution . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2.8 Agricultural Water Pollution . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.3 Wastewater Treatments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.3.1 Primary Treatment of Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.3.2 Secondary Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.3.3 Tertiary Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.4 Wastewater Remediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.4.1 Thermal Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.4.2 Hybrid Methods for the Treatment of Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.5 Wastewater Purification Techniques . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.5.1 Conventional Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.5.2 Advanced Methods Based on Nanotechnology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.6 Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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M. Rafique (*) · M. B. Tahir · I. Sadaf Department of Physics, University of Gujrat, Gujrat, Pakistan e-mail: muhammad.rafi[email protected] © Springer Nature Switzerland AG 2019 R. Prasad, K. Thirugnanasambandham (eds.), Advanced Research in Nanosciences for Water Technology, Nanotechnology in the Life Sciences, https://doi.org/10.1007/978-3-030-02381-2_5

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Introduction: Water and Water Technology

Clean water is a fundamental prerequisite for the development and maintenance of human activities, valuable food via irrigation and aquatic life. Solid and liquid wastes produced by industrial and settlement activities of human pollute most of the water sources throughout the world. Current modern civilization era is causing water contamination due to the development of fast growing economic systems and emerging technologies. In twenty-first century, water will become one of the scarcest resources because of the enormous rise in worldwide population and over five billion global population is accommodated in urban areas in 2015 (Black and Sutila 1994; Day 1996; Economic 2007; Mahadik 2017). Currently, a large proportion of the population is living or trying to live in urban areas to access the existing facilities easily. Urbanization marvels are associated with problems related to water sanitation and municipal services for provision of both fresh water and sanitation. The provision of health care, housing, social services, and basic human needs like clean water and disposal of effluents are the major challenges for engineers, planners, and politicians. As population increases, greater strains will be placed on available resources which pose greater threat to environmental sources. A report by the Secretary General of the United Nations Commission on Sustainable Development showed that there is no sustainability in the utilization and flow employments of fresh water both by developed and developing countries (Economic 2007). Globally, the water use is increasing three times the total population increment, subsequently, prompting across the board general medical issues, limiting economic sources and agricultural improvement and harmful effects on biological systems. The increasing global population and urbanization will limit the resources and create a threat to environmental sources (Goel 2006; Mance 2012). In developing countries, the situation is more adverse, especially the wastewater pollution is increasing rapidly. This water pollution is due to direct discharge of industrial wastewater into water bodies (Dhote et al. 2012; Kanu and Achi 2011). In many developing countries, the bulk industrial and domestic wastewater is ejected directly without any treatment or only after primary treatment (Dhote et al. 2012). In Latin America, 97% of the organic and inorganic sewage raw is ejected directly into environment and only 15% of collected wastewater passes through the plant treatment. The industry rich countries (i.e., China) are ejecting about 55% of sewage without any treatment directly into environment. In Middle Eastern countries like Iran the absolutely untreated sewage is discharged into groundwater (Tajrishy and Abrishamchi 2005). In South Africa, some level of wastewater treatment is examined where an inadequate maintenance of most sewage and poor operational state prompting the contamination of different water bodies and causing serious socioeconomic and health threats. Further, the sub-Saharan Africa discharges wastewater without any treatment. Water pollution is a worldwide problem which needs amendment and assessment of policies for water resources at all levels. The physicochemical processes incorporate coagulation and flocculation using different compound reagents, i.e., aluminum chloride, ferric chloride, and polyelectrolytes, as well as produce a lot of sludge.

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Therefore, increasing demands for drastic change on wastewater regulations and water quality sensing need exceptional development and emerging wastewater treatment like ultrafiltration, ion exchange, reverse osmosis, electrochemical technologies and chemical precipitation. Water resources administration practices perpetually squeezing demands on wastewater treatment innovations to decrease modern adverse effects on natural water sources (Greenlee et al. 2009; Gupta et al. 2012; Shon et al. 2013; Zheng 2011). Therefore, the new procedures and discharge limits should be enforced. The industrial activities are required to pursue new techniques and advancements for the effective control on discharge of heavy metals and hence reduction in volume of wastewater. The recycling and reuse of wastewater to complete the water cycle can help in this regards. Advanced technologies for wastewater treatment are vital in order to eradicate pollution as well as to enhance separation processes. These advancements can be applied effectively as part of ordinary pollution discharge techniques to remove biodegradable organic materials, colloidal substances, heavy metals, nitrogen and phosphorus compound mixtures, suspended solids (SS), dissolved compounds and microorganisms. Moreover, nanotechnology advancements/methods are gaining attraction due to flexibility, security, selectivity, plausibility of mechanization, ecologically inviting and low speculation costs of the techniques. Therefore, to address the wastewater issues, nanotechnology-based techniques for water and wastewater treatment are gaining attention globally. The inspiring properties of nanomaterials and their combination with conventional treatment technologies represent great opportunities to revolutionize water and wastewater treatment. Although the challenges and solutions provided by the nanotechnologybased methods are fascinating, however, many of these challenges are perhaps only temporary, including technical hurdles, high cost, and potential environmental and human risk. The advancement in nanotechnology by routing its direction while avoiding inadvertent consequences can continuously provide robust solutions to our water/wastewater treatment challenges (Savage and Diallo 2005; Aziz et al. 2015; Bhuyan et al. 2015).

5.2

Major Sources of Water Pollution

Sources of water pollution can be broadly classified into two types as shown in Fig. 5.1. (a) Point Sources The pollutants that enter in water through detectable sources and discharge directly into aquatic environment or water are termed as point source pollutions, for example, domestic sewage and oil spill. (b) Non-point Sources When pollutants enter in aquatic environment through numerous-distinct sources it is termed as non-point source of water pollution. The examples are pesticides, industrial wastage, and fertilizer which can be indirectly discharged

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POINT SOURCE Direct discharge into water Sewage Treatment Plants

NON-POINT SOURCE Indirect discharge into water

Dumping Toxins Industrial Facilities

Urban Runoff Agricultural Runoff

Fig. 5.1 Description of point and non-point pollution (sources Burns 2014)

from wastewater. The extents of these impurities are durable to analyze because it arises from variant sources. The primary and important basis of water pollution throughout the world particularly in the USA are non-point source pollution (Schaffner et al. 2009). The point and non-point pollution sources are sub-categorized as follows.

5.2.1

Domestic Sewage

The domestic sewage comes from the water ejected from the apartments and houses called as sanitary sewage. It contains only 0.1% impurities or pollutants and more than 99.9% water. The major pollutants include plant nutrients, organic materials, and many microbes. It is estimated that two million tons of sewage are discharged into the water every day. In developing countries, the situation is worse where more than 90% of untreated industrial wastewater and 70% of untreated industrial waste are dumped into surface water sources. According to US approximation every year the volume of pollutant water from all over the world is 1500 km3. About 18% of people from all over the world use open area to excrete their waste which contaminate soil and water. Domestic sewage has become a main source of water contamination due to high levels of inorganic and organic substances in wastewater. The main causes of domestic sewage are: (i) Laundry, kitchen, and bathroom sources and (ii) Waste from food preparation, dishwashing, litter, toilets, bathrooms, showers, and sinks (Ma Ding et al. 2009).

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Industrial Water Pollution

Different industrial wastewater discharge into the river without treatment is the main source of water pollution. A pollutant discharge depends on the nature of the industry from hazardous and toxic materials discharge and cause surface water and ground water contamination. About 25% of pollution is caused by the industrial water pollution and causing alarming situation. In daily routine, approximately 2 billion heaps corresponding to load of 6.8 billion of whole population is discharged into water from industry. Under developed countries, the industrial effluents are discharged into nearby water bodies which pollute their content. For example, Colorado mines are polluting the water bodies in various contries. The major industries incorporating to water pollution are fertilizer factories, oil mills, food industries, leather tanning, ceramics, petrochemicals, and sugar industries. These industries produce several hundred thousands of wastewater incorporating huge amount of pollutants such as toxic metals magnesium, copper, cadmium, mercury, iron, lead, arsenic, chromium as well as nitrites, nitrates anions and cations. For example, River Kabul in Khyber Pakhtunkhwa, Pakistan, collects an estimated amount about 80,000 m3 of industrial pollutants each day (Greenlee et al. 2009). Indeed, even in the capital city Islamabad (Pakistan) there is no appropriate association of effluents in its two industrial estates and wastes are directly depleted into the Swan River. It has been ordinary that about 1% of wastewater of industries in Pakistan is treated before being released. A several number of industries are sited in surrounding areas or real urban communities. These industries drain out their pollutants directly into ditches, streams, ponds, streams, and agricultural or open lands (Greenlee et al. 2009). It is investigated that huge amount of toxic substances are poured into water bodies which makes hazards effects on environments. In Pakistan, an estimated quantity of 40
109 L of industrial pollutants is directly discharged into water bodies through different industries on daily basis (Scheren et al. 2000).

5.2.3

Population Growth

Increasing population leads to increase in solid waste production and hence causes increase in pollution. In general, solid and liquid wastes are discharged into the rivers. A large number of pathogens from human manure have also been found in wastewater which is harmful to human health. The governments are becoming unable in providing basic needs to their citizens due to rapid increase in population. The increasing population also demands more food (cultivated by water) which is affected by the water quality. Internationally, substantial form of water pollution is an inadequate cleanliness and 2.5 billion of whole population is deprived of proper hygiene. In Asia, more than 70% of population faces hygiene problems; however, in Africa, the situation is worst. Therefore, development of new water resources is needed to meet the increasing population requirements. On the other hand, the water consumption reduction can also help to maintain the existing resources (Clark et al. 1989).

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Pesticides and Fertilizers

Pesticides have been utilized to eliminate pest, bacteria, pesticides, and different germs. But pesticides are directly polluting the water, which affects the quality of water. When amount of pesticides become excessive or poorly managed, will cause harm to the agricultural ecosystem. It is generally observed that only 60% of fertilizer is effectively utilized in the soil and the remaining is absorbed in soil and water as a pollutant. The green sulfur bacteria (chlorobiaceae) are rich in contaminated water. Excess phosphate can cause eutrophication. The various fertilizers applied are not completely absorbed or used by the crops; therefore, a large amount of the fertilizers remains there. The remaining fertilizers are in the form of phosphates, sulfates, nitrates, and ammonia, and hence cause the water contamination. Furthermore, the fertilizers produce the algae into surface water which causes eutrophication and a threat to environment. Many of the algae produce toxicity in water resources. Moreover, the fertilizers also contain the heavy metals, leave the heavy metals in water resources and soil and produce the contamination (Greenlee et al. 2009). Due to floods, heavy rainfall, excessive irrigation, and access to the food chain, chemical residues are mixed with river water. The mixing of these chemicals is fatal to organisms. Many vegetables and fruits are contaminated by these chemicals. The use of pesticides around the world has increased significantly. It aims to protect crops from pests and achieve higher crop yields with higher quality. The world uses an estimated 205 million tons of pesticide every year and continues to increase. The pesticides were introduced firstly in Pakistan in 1954 and have 254 metric tons formulation. Since then, in the late 1960s and early 1970s from Europe and the United States, many fertilizers in thousands of tons have been imported, whereas some other pesticides such as dichlorodiphenyltrichloroethane (DDT) and hexachlorobenzene (BHC) were produced locally. At present, Pakistan is using about 70,000 tons of pesticides annually with 6% annual growth rate. A major part of the pesticides, about 75%, are used for cotton crops and the remaining to other crops such as maize, tobacco, paddy rice, sugarcane, fruits, and vegetables. It is expected that only 0.1% of the applied pesticides is reaching the target organ and the remaining 99.9% are dispersed in air, soil, and water which lead to the pollution of the natural ecosystem and affect the environment. In addition, the manufacturing, processing, and transportation of the pesticides and fertilizers also contribute to the pollution. In nitrate and its isotopes polluted the surrounding it is observed that the ratio of defilement by nitrate has become doubled in eastern Mediterranean and Africa. According to survey in India and Africa, water reservoir consists of more than 15 mg/L of nitrate. These waste pollutants do not remain restricted to surface water but their percolation to the soil results in contamination of ground water aquifers. In developing countries, this issue is more threatening due to improper storage, inappropriate handling of pesticide containers, and the use of expired pesticides (Greenlee et al. 2009). This is mainly due to the lack of understanding of the harmful

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effects to agricultural communities dealing with pesticides. For example, in Pakistan, pesticide residues were reported for the first time in cattle drinking water in Karachi. In Multan (a city of Pakistan), 12 ground water samples were taken from 6 different sites and were found impure with pesticides whereas 33% samples exceeding maximum residues in four cotton growing districts of Punjab Provinces, namely Bahawalnagar, Muzafargarh, Dera Ghazi Khan, and Rajanpur, as well as 90% cities in chains are polluted due to absence of ground water supplies.

5.2.5

Plastics and Polythene Bags

Another main source of water pollution is polythene bags and commodity plastic waste. The garbage is thrown away by putting the plastic bags in the trash. The plastic and polythene bags discharge un-decomposable and dangerous chemicals that are mostly heavy metals, which ultimately pollute the air and water bodies. Incinerator is used to recycle the plastic which further discharge noxious gases and products into environment and enters the food resources and further contaminate the whole food cycle (Mart 1979).

5.2.6

Urbanization

Urbanization is another factor for increase in pollution. Overcrowding, unhygienic conditions, and unsafe dirking water are main health issues in urban areas leading to many infectious diseases. It is observed that one-quarter of the city’s population is susceptible to diseases. Major pollutants found in urban runoff are sediments, nutrients, aerobic Substances, road salts, heavy metals, petroleum hydrocarbons, pathogens, and viruses. Suspended sediments are the largest pollutant load from urban areas water. Construction is a major source of sediment erosion. The sources of contaminated nutrition and bacteria are the use of fertilizers, pet waste, leaves, grass clippings, and septic tanks. In addition, petroleum hydrocarbons are mainly from transport sources.

5.2.7

Ground Pollution

Groundwater pollution is caused by many different industries and locations. Chemicals, detergents, human and animal wastes, and road salts contribute in ground water pollution. Groundwater is one of the most important sources of irrigation water, unfortunately, it is easily contaminated when man-made products such as gasoline, petroleum, pavement salts, and chemicals enter groundwater and make it unsafe and unfit for human use. In addition, most of these pollutants flow into the

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oceans carrying agricultural fertilizers and pesticides. Therefore, 80% of the marine environmental pollution comes from land. Moreover, one of the largest sources of pollution is called non-point source pollution which is caused by runoff. Since there are many different pollutants that can cause groundwater problems, it is a good idea to understand the source of these harmful substances. Residents all over the United States have been causing great damage to the nearby groundwater but they have not realized it because they are pouring detergents and cleaners into their yards or sewers, using garbage disposal instead of composting, and let animals garbage sitting on their yard, i.e., did not clean up for a long time. The construction sites also produce toxic runoff which infiltrate the groundwater and cause pollution. If you do not pay attention to keeping these runoffs and prevent them from escaping the construction site, the water in the entire area is at risk. The construction site will produce toxic runoff, infiltrate the groundwater, and cause pollution. If you do not pay attention to keeping these runoffs and prevent them from escaping the construction site, the water in the entire area is at risk (Gorelick et al. 1983).

5.2.8

Agricultural Water Pollution

Water pollution is caused due to the widespread use of agrochemicals in agriculture. The agricultural chemicals used in crop-fields such as fertilizers and pesticides become mixed in irrigation water and leaches through the soil and eventually reaches natural water resources. Agricultural drainage leads to overall pollution of water resources but it is lower than industrial and household waste. In agriculture, different fertilizers are used which are the sources of nitrogen and phosphorus. The drainage of these fertilizers reduce oxygen level and become dangerous for many species disturbing the ecosystem. Water pollution with agricultural chemicals has been reported in residential countries like China and the USA. However, conditions in Pakistan are not different from the world (Greenlee et al. 2009). A comprehensive description of water pollution is given in Fig. 5.2.

5.3

Wastewater Treatments

Organic water pollutants include ordinary hydrocarbons, polynuclear hydrocarbons, ketones, proteins, greases, and detergents. The heavy metals are inorganic pollutants which are very hazardous and highly toxic. The biological pollutants are including bacteria, fungi, amoeba, algae, and viruses. That’s why, all over the world ground water is polluted due to these types of pollutants and cannot be used for drinking purposes. Therefore, the major problem is wastewater treatment and its reuse. In last decades, a lot of work has been done to improve and preserve the quality of water.

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Fig. 5.2 Schematic representation of water pollutants

Nowadays, more efforts have been made by researchers in improving the wastewater treatment processes by appropriate technologies. There are three basics stages of wastewater treatment: (i) primary treatment, (ii) secondary treatment, and (iii) tertiary treatment. Primary treatment is also called preliminary stage of purification which includes grit removal, screening, grinding, and sedimentation. The secondary treatment treated wastewater via biological methods, whereas in tertiary treatment wastewater is treated by advance methods (Fig. 5.3). After primary and secondary treatment, wastewater is converted into clean water by processing in tertiary wastewater treatment. After tertiary treatment, good quality and highly purified water can be used for different purposes such as in drinking, medical, and industry. After passing through tertiary process, 99% of contaminants detached or removed and water is converted into high quality. The details of water treatment methods are as follows

5.3.1

Primary Treatment of Wastewater

The primary treatment is applied on the heavily polluted water. In this treatment, numbers of techniques are used for the removal of sand, suspended solid materials,

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Fig. 5.3 Schematic representation of classification of wastewater treatment

and floatable matter from the wastewater. This treatment is applicable for the polluted water which can include pharmaceutical produces and personal care products (PPCPs), grease, oil on water surface. The objective of preliminary treatment is to remove suspended solid (SS) and other large materials present in wastewater whereas in primary treatment organic and inorganic materials are removed from the water. Primary treatment decreases primary burden of wastewater for next stage. A best system operated removes up to 90% of settable solids and 40–60% of suspended materials (Ødegaard 1992).

5.3.2

Secondary Treatment

Biological contents of the sewage such as driven from food waste, human waste, soaps, and detergents are degraded by using secondary treatment. This treatment is normally performed via indigenous, aquatic bacterial in natural environment. Separation procedure is required for the removal of the microorganisms from already treated water preceding the discharge or the tertiary treatment. In secondary treatment, aerobic biological process is used to treat the settled sewage liquor by municipal plants. The microorganisms and protozoa use biodegradable solvent natural contaminants, i.e., sugars, fats, and natural short-chain carbon atoms, and tie a great part of the less solvent fractions into assembled. The secondary treatment can be classified on the basis of biomass as the fixed film and the suspended development. In suspended-development system, i.e., actuated sludge, the biomass is very much blended with the sewage and the required working space is smaller than

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settled film system. But, settled film systems are more capable to work with exceptional discharge rates of organic materials and higher changes in biological materials as compared to suspended-development system. Roughing filters are applied in treatment of the variable or solid organic waste from industries under secondary treatment. Comparatively, a high rate wastewater can be applied to open synthetic filter incorporating the circular filters. A high air progression as well as high pressure driven stacking are intended to permit. However, in large assemblies, the air progression is controlled by the blowers. The treated wastewater are needed for further treatment, as advance secondary treatment, to settle down the filtered material, remove the suspended and organic material (Gupta et al. 2012).

5.3.3

Tertiary Treatment

After preliminary, primary, and secondary treatments the chemical properties of wastewater are not similar to freshwater used for drinking purposes and many domestic purposes like cooking. For example, the pH of water is either less or greater than neural water, i.e., water is either acidic or basic in nature. To overcome all these problems water is further treated using tertiary treatment and water is proficient to use for domestic, agricultural, and industrial purposes. Hence, we have recycled the wastewater and in this way the shortage of water on our planet can be reduced to some extent. The tenacity of tertiary treatment is to give a last treatment, i.e., to raise the effluent quality before it is discharged to ocean, waterway, lake, ground, etc. In addition, one tertiary treatment process must be used in any treatment plant leads to the final process. This is known as effluent polishing which incorporates UV treatment, micron filtration, reverse osmosis, and ozonation (Gupta et al. 2012). The technologies of tertiary treatment are classified as follows.

5.4

Wastewater Remediation

5.4.1

Thermal Methods

Many thermal technologies are used for treatment of wastewater and are classified into following three categories.

5.4.1.1

Pyrolysis and Gasification

Pyrolysis is the process of decomposition of different dyes in particular anaerobic environment under very high temperature in the range of 300–900  C. In response to pyrolysis, following three main products are obtained:

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1. A fraction of gas: this gas consists of H2, CH4, CO2, and CO, and small fraction of some other gases is also present. 2. A fraction of solid: heavy metals and small amount of inert elements. 3. A liquid: water, oils, tar, and some carbon-based compounds. Different obtained products depend upon temperature and pressure during experimentation, disorder in the reactor and number of other factors as well as oil is obtained due to pyrolysis process. This process provides concurrently aeration of wastewater before its ignition or gasification and used for mixing of raw materials into oil wastage. The solid waste and liquid phases are separated through the centrifuge process and this water used as fuel. In the process of gasification, the solid is treated into a chamber that is collected after centrifuge and made an interaction with specific gas like oxygen and air. The gasification of sewage sludge results in flammable gas and this gas can be used in many useful resources like production of electricity. There are many applications of gasification process (Fytili and Zabaniotou 2008; Nadziakiewicz et al. 2007).

5.4.1.2

Rapid Thermal Conditioning

The rapid thermal conditioning is a useful technique for the treatment of sludge. In this process, the sludge is heated up to reaction temperature and continuous quenching of few seconds is done. The necessary conditions for this process are: 1. Unstable adjourned solids are hydrolyzed, which reduces solid output from wastewater. 2. The dewater ability characteristics of the slush are enhanced, hence clotheshorse cake is found by centrifugation even deprived of polymer addition. 3. The subsequent solvable carbon-based materials disturb biotic action and improve gas production in succeeding anaerobic absorption as well as expense of solids discarding is further reduced. The CH4 may be used as fuel for the transmission control protocol (TCP), creating the complete procedure thermally self-reliant. 4. The HT of RTC terminates pathogens, making it able to accomplish the requirements of the slush as distinct in the US EPA Sewage Sludge Use and Removal Regulations. This permits extensive terrestrial application of the remaining bio solids (Nadziakiewicz et al. 2007). 5.4.1.3

Microwave Based Pyrolysis

The removal of urban and industrial slush seeks a great attention and treatment of such slush has become a major problem. The treatment of slush is considered a difficult task as smaller or larger amount of the contamination remains in the treated solution. The further common substitutes of action and dumping of dirt mud are sludge burning and ocean clearance, none of them is problem free. For treatment of

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such kind of slush many reports suggested that most appropriate technique is the pyrolysis. This method appears to be less pollutant than burning as it distillates the metals of higher density in a solid carbonaceous filtrate. This treatment is such useful that its by-products can also be used as the fuel. In addition, the microwaves treatment also provides an alternative method of pyrolysis of different plant material and wastage of animals that can be used as fuel. In particular, it can be used to apply this process on different types of carbon-based compounds. Being deprived receiver of microwave type energy, so as usually to achieve the high pyrolysis, these are not provided with very high thermal energy. However, if the raw material is mixed up with material whose efficiency to absorb the microwave energy is higher, the method can become more effective. The best materials to absorb this energy are the carbon, different types of semiconductors, and certain metal oxides (MO).

5.4.2

Hybrid Methods for the Treatment of Wastewater

Most of hybrid-methods are the combination of unconventional oxidation processes. The important and effective hybrid-methods are: Ultrasound/Ozone, Ultraviolet/ Ozone, Ozone/H2O2, and Photo Fenton Process. In these methods, free radicals are produced which subsequently attack on pollutant molecules. These are mainly effective to bio refractory particles having an objective to entirely mineralize the impurities and it can also transform the impurities into less toxic, less harmful, and mediocre chain composites which can be treated biologically. The effectiveness of all these procedures depends on production of degree of free radicals. Moreover, the degree of interaction of produced radicals with pollutant particles and well-organized scheme should be capable to maximize these capacities (Gogate and Pandit 2004).

5.4.2.1

Ultrasound/Ozone Process

The main driving mechanism in degradation of pollutants utilizing ultrasound is production and succeeding attack of free radicals but few of the reactions are described properly on the basis of hotspot theory, i.e., localized generation of extreme conditions of pressure and temperature. In regulatory mechanism, the free radical attack, the rate of degradation is enhanced through utilization of ozone or hydrogen-peroxide due to production of additional free radicals. The reversible reaction of dissociation of hydrogen-peroxide provides free radicals, however, the combination of hydrogen-peroxide with ultrasound becomes a proficient route to enhance generation of free radicals. In combined process, amount of enhancement in generation of free radicles is determined by the concentration of hydrogen peroxide. The scavengers of the generated free radicals are represented in Eq. (5.1):

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OH þ H2 O2 ! H2 O þ HO∗ 2

ð5:1Þ

The significance of combined process depends upon the usage of free radicals by pollutant particles and also depends upon effective interaction of free radicals with pollutants. The selection of combination of hydrogen peroxide and ultrasounds as oxidation treatment depends upon turbulence intensity within reactor, operating pH, composition of sewage concentration, and nature of pollutants.

5.4.2.2

Ultraviolet/Ozone

The combination techniques such as combination of ultraviolet light with ozone or hydrogen peroxide produce higher rate of free radicals than the individual process. The combination techniques are similar with ultrasound/ozone or ultrasound/H2O2. The key points for wastewater treatment by combination techniques such as UV photolysis with H2O2 are: 1. The continuous bubbling and higher partial pressure of ozone should be used. 2. The suitable initial concentration of pollutants should be low and sewage wastewater requires dilution. 3. The rate of degradation and ozone stability are attained at optimized higher temperatures. 4. The operating pH must be neutral or in the range of slightly alkaline, i.e., 7–8. 5. The degradation process are inhibits due to existence of radical scavengers in higher concentration such as control mechanism of free radical attack. 5.4.2.3

Ozone/H2O2

The ability of O3 in oxidizing several pollutants through direct attack on bonds such as C¼C and aromatic rings is superior due to existence of hydrogen peroxide for generation of highly reactive hydroxyl (OH) radicals. Hydro-peroxide ions are produced by the dissociation of hydrogen peroxide which attack on ozone molecules resulting in the production of hydroxyl radicals. The advance oxidation technology based on combination of ozone and hydrogen peroxide is summarized as: • The advance oxidation technology was explored for systems having no chain initiators, i.e., presence of humic acid in low concentration as well as the pollutants having low reactivity towards the molecular ozone attack. • The optimized operating conditions are neutral pH, lower concentration of initial pollutants; optimum concentration of hydrogen peroxide and ozone as well as lower radicals scavenges concentration and neutralized lower organic solutes concentrations. • The structure of reactor for growing ozone is a vital factor. Spray columns are favorable than stirred or bubble columns and their efficiency in treating huge amount of pollutants should be tested. Utilization of inert mixers has been accounted to give peculiar rates of ozone ingestion and furthermore incredible

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Fig. 5.4 Schematic of photon Fenton reactions pathways (Cheng et al. 2016)

gas fluid blending. Utilization of ultrasonic atomizers or so far as that concerned even ultrasonic cascade type reactors should yield valuable products. • Multistage distillation systems, another current invention, provide high yield as compared to conventional operations. By utilizing the reactors in series, the addition of ozone or hydrogen-peroxide can be maintained in proceeding steps depending upon the remaining concentration of oxidizing products and pollutants.

5.4.2.4

Photo Fenton Process

A combination of UV radiation and hydrogen-peroxide with oxalate ion Fe(III) or Fe (II), i.e., the photo Fenton process, produces more hydroxyl radicals as compared to traditional Fenton method. This stimulates the degradation rate of organic pollutants. The exact concentration and mechanism of Fe(III) and Fe(II) ions are very complicated. The chemical reaction are initiated by primary photo reduction of dissolved species (Fe(III) and Fe(II)) through oxidation of organic compounds and Fenton’s reaction (as shown in Fig. 5.4). Furthermore, hydroxyl radicals produced in the first step are also involved in oxidation reaction. The OH radical becomes unavailable for the purpose of destruction through recombination of free radicals and scavenging via excess hydrogen peroxide. The primary photon Fenton oxidation reaction by UV and visible light is described by the relation in Eq. (5.2): FeðOHÞ2þ þ hυ ! ðFeÞ2þ þ OH∗

ð5:2Þ

Photo reactive oxalate complexes Fe(III) providing more quantum yield of Fe (II) than Fe (III) and absorption ranges are increased up to 570 nm and leading to rise in coefficient of absorption. The reaction is written as Eqs. (5.3) and (5.4):

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ð5:3Þ

3 C2 O∗ ! FeðIIIÞ þ 2C2 O2 4 þ FeðC2 O4 Þ 4 þ CO2

ð5:4Þ

The Fe (II) ions are generated by reaction of Fe(III) oxalate complex with intermediate oxalate radicals. The free radicals are generated by the reaction of hydrogen peroxide with generated Fe(II) ions. The Fe(II) ions are regenerated by the reaction of Fe ions with hydrogen peroxide in 3q process by the Eqs. (5.5) and (5.6):  3þ Fe2þ þ H2 O2 ! HO∗ 2 þ HO þ Fe

ð5:5Þ

þ 2þ H2 O2 þ Fe3þ ! HO∗ 2 þ H þ Fe

ð5:6Þ

The mineralization of organic pollutants is achieved by oxidative destruction of pollutants in water through highly reactive hydroxyl radicals and the relations are represented in (5.7) and (5.8) equation (Gogate and Pandit 2004). OH∗ þ RH ! R∗ þ H2 O ∗

R þ O2 !

5.5

RO∗ 2

ð5:7Þ ð5:8Þ

Wastewater Purification Techniques

5.5.1

Conventional Methods

Wastewater is a mixture of various organic and inorganic pollutants. This wastewater is treated via conventional methods to reduce offensive properties, which can cause harmful effects on ecosystem and humans. The conventional methods are utilized to decrease the amount of floatable and suspended materials as well as treatment of biodegradable organic matters present in wastewater. But toxic substances, dissolved solid, suspended solids, pathogens (strong oocyte composition), and all nutrients were not discharged by conventional methods. Nowadays, advance methods are used to discharge all organic and inorganic pollutants present in it.

5.5.1.1

Sedimentation

Sedimentation is simplest route for treatment of wastewater under the influence of gravity where the pollutants with density greater than water are separated. Physicochemical methods are attracting the attention of researchers like floatation, sedimentation, adsorption, coagulation, etc. in this regards. Sedimentation is one and

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important part of industrial wastewater purification. The pollutants are separated from the wastewater in the form of solid material known as sludge. Flick’s law, a mass transport law, determines that the removal of the suspended particles in the form of sludge depends upon the process time and the height of the tank used for the process. For example, a pilot study was carried out which showed the relation of the tank design and the wastewater treatment. It was concluded that the plant (including different part as tank and others) plays a vital role in defining the efficiency of the process and hence quality of the treated water. Currently, gravitation sedimentation is the most common process for treatment of industrial wastewater. Gravity coagulation and the clarification are the main process in the sedimentation. Higher polluted wastewater is treated under gravity coagulation whereas lower polluted water is treated under clarification.

5.5.1.2

Sand Filtration

Sand filtration is known as chemical-physical process for extrication of colloidal and suspended impurities from wastewater through passage bed of granular material. In filtration process, impurities were trapped in the openings or absorbed on the surface of grains and water fills in the filter medium pores. Sand filtration is categorized as: (i) Slow sand filtration and (ii) Rapid sand filtration.

Slow Sand Filtration The slow sand filtration uses biological, physical, and chemical methods for purification of wastewater. Within the same filter bed both physical and biological process works parallel to each other. The discharge of pollutants occurs at upper layer of filter via biological methods, which is followed by processes of degradation, adsorption, and mechanical filtration. The removal of pollutants depends upon physical factors such as fine sand and slow filtration rate, i.e., 0.1–0.3 m/h. Bacterial immobilization processes are followed by predation via protozoa and adsorption onto sand particle surface. After reduction of sand filtration viruses, extracellular production of microbial products, i.e., higher microorganism on virus particles and grazing of bacteria are suitable mechanism. The cysts of Cryptosporidium enteroparasites and Giardia are also potentially discharged using slow sand filtration. The supernatant water over the sand bed is around 100–150 cm deep, which depends upon the design of filters. The average time, which the sample residues over the sand bed depends on filtration rate ranges from 3 to 12 h. The lighter particles of suspended matters are drawn into the pores between the sand grains and discharged by straining on the top of millimeters, whereas the heavier particles begin to settle. A layer of organic matter and inactive grain systems are on the top layer of sand bed during the process of filtration. This layer is stated as Schmutzdecke as shown in Fig. 5.5. Furthermore, biological growth also takes place within the sand bed and gravel support which have enormous effect in the purification system (Verma et al. 2017).

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Fig. 5.5 Representation of slow sand filtration mechanism

Rapid Sand Filtration The physical adsorption and mechanical straining are two principles of operating rapid sand filtration as shown in Fig. 5.6. The media utilized significantly grainier with grain size of 0.6–2.0 mm in rapid filtration process. The interstices between the grains are bigger, giving less resistance from the descending flow, and allowing higher speeds in the range of 5–15 m3/m2/h. Rapid sand filtration can attain usually hydraulic high flow rates and automatically clean by backwash the systems to banish aggregated solids. The mechanism of particles through filtration depends upon adsorption and some straining, while does not depend upon biological filtration. The need for cleaning is to reestablish the quality of pollutants and capacity of the filter that takes place at regular intervals. The high-pressure water is driven upwards through the mechanical agitation and whole bed is utilized to scour the individual grains with the goal that the collected pollutants can be flushed away in order to clean throughout its whole depth (Nassar and Hajjaj 2013; Yao et al. 1971).

5.5.1.3

Micro Membrane Processes/Filtration

Micro filters consist of porous fibers and very smaller pores (smaller than 0.1 μm) are used in this process to filter the wastewater. Pathogens such as bacteria, protozoan cysts including sand, silt, clays, giardia lamblia and cryptosporidium cysts, algae, and some bacterial species (with size greater than micron) are bigger than pores and are hindered behind the membrane. By applying pressure, the water is pushed through the pores of the hollow fiber membranes. Since the water molecules are small enough to pass through and some useful impurities also pass down, clean water collects on the other side of the membrane. The hollow fibers are arranged in a special U-shape design which leads to the clean water towards the open ends of

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Fig. 5.6 Schematic illustration of principles of rapid sand filtration (WHO 1996)

Mechanical Straining

Physical Adsorption

individual fibers into mouthpiece of bottle. The pollutants are left on the outside of the fibers and can be washed away easily. Micro filtration is not an absolute barrier to viruses, but in combination with disinfection, micro filtration appears to regulate these microorganisms in water. Therefore, this method can be used for the removal of chemicals, pathogens, synthetic organic matter, etc.

5.5.2

Advanced Methods Based on Nanotechnology

Advanced treatment methods based on nanotechnology being developed and utilized in order to boost desalination of brackish and sea water, decontamination and disinfection and safe reuse of wastewater, i.e., nano-photocatalysis for chemical degradation of pollutants, nano-filtration, different membrane processes, ultrafiltration, nano-sensors for pathogens detection.

5.5.2.1

Membranes and Membrane Processes

The removal of undesired ingredients from water is the basic goal of water treatment and these ingredients are hindrance on the basis of their size. The membrane process gives high level of mechanization, no chemical, and requires less space, and in addition modular configuration allows flexible design. A major challenge to the membrane technology is the characteristic change between layer selectivity and penetrability. But high energy is required for the utilization of membrane based techniques in pressure ambitious membrane processes and the cleaning of the layers throughout the process. The utilization of high energy and pressure diminishes the lifetime of membrane and membrane units. In addition, the integration of functional nanomaterials into membranes gives an incredible opportunity to improve the membrane fouling resistance, thermal and mechanical stability and in addition new functions for pollutant removal with self-cleaning.

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Nano-Fiber Membranes Electro-spinning is a simple, efficient, and reasonably assessed method to make ultrafine fibers using a variety of resources. The resulting nanofibers have high specific surface area porosity and nanofiber with multilayered entire arrangement. The width, morphology, work of art, secondary structure, and spatial alignment of electro spun nanofiber scan is easily manipulated for specific applications. The potential of nanofiber membranes for water treatment is still idle, but commercially utilized for air filtration applications. Without significant fouling, nanofiber membrane can discharge micron size particles at high refusal rate from aqueous phase. However, nanofiber membrane was utilized as pretreatment earlier to reverse osmosis and ultrafiltration. Nanofibers impregnated membrane is fabricated by doping of functional nanomaterials. The unique characteristics and tunable properties make electro spun nanofibers an ideal stage for making multifunctional media/film filters by either specifically utilizing basically multifunctional materials, for example, TiO2 or by presenting practical materials on the nanofibers. For instance, by gap in ceramic nanomaterials or specific confine execute on the affinity nanofibers membranes, nanofibers gallows can be intended to discharge heavy metals and natural pollutants during the process of filtration.

Nanocomposite Membranes Nowadays, multifunction systems are produced by the addition of nanomaterials into inorganic or polymeric membrane through the process of membrane technology. Nanomaterials incorporated in such applications are catalytic nanomaterials, hydrophilic metal oxide, and antimicrobial nanoparticles. The membrane surface hydrophilicity, fouling resistance, and water permeability are enhanced by the addition of metal oxide nanoparticles such as TiO2, silica, and alumina into polymeric ultrafiltration membrane. The mechanical and thermal stability of polymeric membrane are boost by inorganic nanoparticles. The membrane biofouling could be diminished by using antimicrobial nanoparticles, i.e., CNTs and nano-Ag. Silver nanoparticles have been surface grafted or doped on polymeric membrane in order to reduce biofilm synthesis, infection of inactive viruses, as well as bacterial attachment. But, suitable substitution of silver nanoparticles is required due to its short-term efficiency against membrane biofouling. It is investigated that highest inactivation of bacteria was attained by polyvinyl-N-carbazole SWWNT nanocomposite at 3 wt% of CNTs. On the other, traditional methods are used for inactivation of single-walled nanotubes, which is insoluble in water, in this process there is no prerequisite for substitution. For inactivation, permeability of polysulfide membrane and long-term archiehydrophilicity are achieved by straight contact of CNTs. The presence of photocatalytic nanoparticles within membrane associate with their physical division function and pollutants degradation via reactivity of catalyst. Currently, researchers focused to fabricate photocatalytic inert membranes, i.e., nanoparticles of TiO2 and

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modified nanoparticles of TiO2. Bimetallic or metallic nanomaterials supported on nano zero valent iron (nZVI) were added into polymeric membrane for Reductive degradation of chlorinated compounds.

Thin Film Nanocomposite (TFN) Membranes The advancement of Thin Film Nanocomposite (TFN) membranes pay particular attention on addition of nanomaterials into active layer of TFC through the process of surface modification or doping in casting solution. The nanomaterials incorporated in such applications are CNTs, nanoparticles of titanium dioxide, and nanozeolites. The effect of nanomaterials on selectivity and membrane permeability depends upon concentration, type, and size of nanomaterials, which is integrated in membrane. The membrane permeability is frequently increased by the addition of nano-zeolite as dopants in TFN. The incorporation of nano-zeolite are considered to be more negatively charged, thicker active layer of polyamide, as well as more permeable. The TFC membrane has been utilized to obtain about 80% improvement in water permeability over membrane layers with high rejection of salt. It is investigated that, nano-zeolite (250 nm) dopants TFC membrane at 0.2 wt% attained superior salt rejection, i.e., >99.4% and higher permeability as compared to commercial reverse osmosis membrane. The hydrophilic small pores of nano-zeolites provide privileged paths for water. In any case, water permeability expanded even with pore-filled zeolites, but less as compared to pore size, which could be ascribed to defects at the zeolite-polymer interface. Nano-zeolites were additionally utilized as carriers for antimicrobial particles particularly, Ag+, imperative to fouling property to the film. The zeolite TFN improvement has achieved the beginning period of commercialization. A seawater thin film nanocomposite reverse osmosis membrane (quantum flux) is commercially available. It was accessible that the small hydrophilic pores of nano-zeolites create privileged superior paths for water. The rejection ability of membrane has been increased by the addition of nano TiO2 into active layer of TFC, whereas upholding the permeability. The organic pollutants and inactive bacteria are degraded using TiO2 under the influence of UV irradiation. These discharged organic and biological pollutants are not retained through the membrane. The incorporation of membrane and photocatalyst leads to enhance the long-term efficiency of polymeric materials. Moreover, CNTs have potential applications in TFN membranes owing to their antimicrobial activities.

Biologically Inspired Membranes Numerous biological membranes are highly permeable and selective, which makes their utilization in polymeric membranes to enhance membrane performance. Aquaporins are protein channels that control water flux cross over cell membranes. Aquaporin-Z from Escherichia coli has been incorporated into amphiphilic triblock-

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polymer, which demonstrates water permeability over original vesicles with complete rejection to urea, glucose, salt, and glycerol. One potential framework is to coat aquaporin in corporate lipid bilayers on productive nano-filtration layers. On this front, constrained achievement was accomplished. CNTs provide high water penetration (theoretical and experimental) as compared to calculated by Hagen Poiseuille infiltration equation predicts due to their atomic smoothness of nano channel and the one dimensional single-file entreating of water particles by passing through nanotubes. It was investigated that a membrane containing about 0.03% surface region of adjusted CNTs will have flux more as compared to commercially available reverse osmosis seawater membrane. But, high negative reaction for salt and small atoms is needed for aligned CNT membranes because of the absence of CNTs with consistent sub-nanometer diameter. The selectivity of aligned CNT membrane has been enhanced by function group gating at the opening of nanotubes. On the other hand, KCl has 50% rejection at 0.3 mM and diminished to zero at 10 mM. The salts were excluding physically by attachment of bulk functional groups at opening of nanotubes, whereas membrane permeability are reduce by keeping out steric. Yet, aligned CNTs are not able to discharge salt while to attain reliable salt (negative) response nanotubes diameter should be uniform and smaller as compared to 0.8 nm. A key boundary for both aligned CNT and aquaporin CNT layers is the scale-up of the nanomaterial generation and membrane manufacture. Vast scale creation and purification of aquaporins are extremely difficult. To date, synthetic vapor testimony is the most well-known approach to make aligned nanotubes. Chemical vapor deposition (CVD) model has been intended for delivering vertically aligned CNT preparing for expansive scale generation. A post-producing arrangement strategy utilizing attractive field was likewise created. Nanocomposite and TFN layers have great versatility as they can be produced by current industrial assembling technique.

5.5.2.2

Photocatalysis for Wastewater Treatment

Photocatalysis is considered as more effective technique for water purification which also affects our environmental pollution and suppresses energy crises. Metal oxide photocatalysts have emerged as promising candidate for the environmental pollution remediation and nowadays metal oxide catalysts were used for purification of water. In this treatment, photocatalyst is illuminated by light which is activated and as a result electron hole pairs are generated. These photo-generated electrons tend to jump from valence band (VB) towards the conduction band (CB) leaving behind a hole. Oxidation reduction reaction occurs which results in the formation of superoxide anions and hydroxyl radical which further react with organic dyes and degrade it as shown in Fig. 5.7a, b. TiO2 is considered as an efficient and promising semiconductor photocatalyst for degradation of organic dyes from wastewater due to its stability, non-toxicity, and high photocatalytic efficiency. On the other hand, large band gap, high charge carrier recombination rate, and low quantum efficiency in visible region of electromagnetic spectrum reduce its photocatalytic performance. Recently, other metal oxide based

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Fig. 5.7 (a) Schematic representation of wastewater treatment and dye degradation via photocatalytic process (Xiao et al. 2015). (b) Schematic representation of mechanisms of photocatalysis (Nguyen and Wu 2018)

materials have been used to overcome limitations of TiO2. Ag3PO4 is used for photocatalytic degradation in a wide range, hydrogen production from water splitting, photo reduction of CO2 into more useful fuels and O2 evaluation (Egerton 2014). Various methods have been utilized to tune the properties of TiO2 which modified its structural and optical (light absorption) properties. The methods for properties tuning are doping, coupling, and composite formation which enhanced absorption edge towards visible region of electromagnetic spectrum. This doping, coupling, and composite formation directly affects its surface and structural morphology which influence the band gap energy. These modifications result into narrowing of band

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gap which eventually results into enhancement in visible light absorption capability of photocatalyst. Among all photocatalyst, TiO2 has widely used visible light driven photocatalyst because of its absorption capability, high photocatalytic activity, and stability.

5.5.2.3

Water Splitting Process

A process used to split water into hydrogen (H2) and oxygen (O2) is known as water splitting. Water splitting has attracted great attention because this process has a potential to produce the energy. The major advantage of producing energy through water splitting is that there is no emission of carbon dioxide. The backbone of water splitting process is the photocatalytic activity. In photocatalysis, the catalyst should be visible light responsive. If catalyst is provided enough energy to split water into oxygen and hydrogen, it is the water splitting process. But, there are some restrictions to this approach. Firstly, suitable thermo dynamical energy must be provided to the catalyst. Secondly, band gap should be narrow enough so that visible photons may cross them. Thirdly, there should be stability against photo corrosion. Therefore, the utilization of the processes is based on the fulfillment of these three requirements. It is much difficult to fulfill all these requirements in normal conditions for commercial level use. Therefore, the researchers are switching to the next process where the said conditions can be modified as per approachable manners. In this second approach, two different catalysts are used instead of a single catalyst and the excitation mechanism is completed in two steps. This process is similar to the natural photosynthesis in green plants and is known as Z-scheme. In this approach the required Gibbs free energy for photocatalyst is less than one-step water splitting process and hence there is wider range of visible light. This results in the separation of hydrogen and oxygen; possibly we can also use a suitable semiconductor for one side of the system. In Z-scheme water splitting process, the energy of photon is converted into chemical energy along with a change in Gibbs free energy in a large amount and the reaction is called uphill reaction. On the other hand, there are some other photocatalytic degradation reactions which are known as downhill reactions that are used to take the advantage, and the most used catalyst in such reactions is titanium dioxide (TiO2). The water splitting using such electrode was reported by Honda-Fujishima in early 1970s in which electrode is irradiated by using ultraviolet light in which electrons and holes are produced and are shown in Fig. 5.8. A potential difference is applied between cathode and anode due to which the generated electrons reduce water to hydrogen on Pt electrode whereas a hole oxidizes water to oxygen on TiO2 electrodes. According to literature, the use of photocatalysts for water splitting is pessimistic and quiet progressive. Therefore, researchers are finding more advantageous new materials for water splitting.

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Fig. 5.8 Photosynthesis by green plants and photocatalytic water splitting as an artificial photosynthesis (Kudo and Miseki 2009)

5.5.2.4

Wastewater Treatment and Hydrogen Production by Water Splitting

Many semiconductors have been reported for an efficient oxygen evolution reaction, water splitting and HER. TiO2 is the most studied photocatalyst for water splitting due to its abundant availability, non-corrosive, cheap, environmental friendly, and stable nature. But, TiO2 cannot utilize whole visible light spectrum but only utilize the spectrum around wavelength 400 nm where it becomes active and the solar conversion efficiency is just 2% as similar to photosynthetic system in plants. However, the efficiency is increased from 16 to 32% as the wavelength is increased from 600 nm to 800 nm. Therefore, the selection of photocatalyst for whole visible light region and infrared region is critical. Many nitrides, oxides, and sulfate of do or d10 transition metal cations like CdSe, CdS, Ta3N5, TaON, C3N4, SiC, BiVO4, WO3, Cu2O, Fe2O3, etc. have been utilized as photocatalysts for water splitting. The performance of metal oxides like Fe2O3, BiVO4, and WO3 has constraints like poor charge transfer, high rate of recombination and lesser light absorption or harvesting. For example, BiVO4 can be used as a photocatalyst, but the conduction band potential is unsuitable for HER. The theoretical solar conversion efficiency of Cu2O is 18% which favors the use as catalyst, but susceptibility to self-reduction restricted its use. Similarly, CdS, a potential candidate, is also susceptible to self-oxidation. However, the photocatalytic activity of TiO2 and CdS based photocatalysts can be plentiful. The low cost and environment friendly solar hydrogen production technology can be developed by using nanosized TiO2 photocatalytic activity and the water splitting technology. But, it still has many barriers to be commercialized as low solar-to-hydrogen energy conversion efficiency, fast recombination of photo-generated electron hole pairs, backward reaction, and the poor activation by visible light. Therefore, researchers have focused on different mechanisms in order to increase the photocatalysis. The photocatalytic activity is controlled and tuned by modifying the TiO2 using metal ion doping, metal ion implantation, metal loading, dye sensitization, composite

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semiconductor, anion doping and composites with g-C3N4, C3N3, C5N3, C10N3, boron based B4.3C and B13C2, grapheme, grapheme oxide, quantum dots and carbon nanotubes. Hence, the modified TiO2 can be an efficient photocatalyst and can play an important role in efficient photocatalytic hydrogen production (Ni et al. 2007). In addition, the conjugated polymers are contenders in scaling up the photocatalytic processes due to their non-toxicity, robustness, visible light absorption, and tunable physiochemical properties. The apparent quantum yield (AQY) is reported in the ranges from 0.015 to 38.8% at 420 nm wavelength and hydrogen evaluation reaction (HER) up to 1970 μmolh1 g1 for 1,3,5-tris-(4-formyl-phenyl)triazine (TFPT) covalent organic framework. The presence of the co-catalyst and the sacrificial agents are the requirements for polymer P7 and TFPT in enhancing the catalytic performance. For instance, 500 μmolh1 g1 activities can be achieved by using co-catalysts and sacrificial agents (Tentu and Basu 2017). However, metal-free and low-cost materials have been reported as the carbon-based materials such as carbon nanotubes (CNTs), g-C3N4, and grapheme. For example, some advantages of g-C3N4 include perfect positions of band edges for water splitting, thermally and chemically stabile as well as low-cost and earth-abundant precursors. However, the grain boundary effects, low surface area, and high recombination rates are still limitations which restrict the performance of g-C3N4 (Zhang and Wang 2012).

5.5.2.5

Reverse Osmosis (RO)

Reverse osmosis is also called as hyper-filtration which is known as finest filtration and permit the separation of particles as small as ions from a solution. Reverse osmosis is applied to purify water and discharge different pollutants and salts in order to enhance properties of water such as taste and color. The most well-known use of reverse osmosis is in purifying wastewater. It utilizes a membrane that is semipenetrable, permitting the liquid that is being purified to go through it, while discharging the pollutants as shown in Fig. 5.9. When a portion of the liquid goes through the membrane, the pollutants will be filtered out which are then evaporated. Further, reverse osmosis process utilizes a method known as cross-stream to enable the membrane to consistently clean itself. The procedure of inverse osmosis requires a main thrust to push the liquid through the membrane and most widely interested force is pressure through pump that is five to ten times higher than ultrafiltration. Reverse osmosis is adequate for discharging microscopic organisms, salts, sugars, proteins, particles, fats, and different constituents that have an atomic weight of more than 0.15–0.25 kDa. The separation of ions depends upon the charge nature which implies that dissolved pollutants which have charge (i.e., salts) will probably be discharged through the membrane than those which do not have charge (i.e., organics). The greater the charge and the bigger the molecule, the more probable the molecule is discharged. The nano-filtration and reverse osmosis involves solution diffusion mechanism (Chen et al. 2017).

5 Nanotechnology: An Innovative Way for Wastewater Treatment and Purification Fig. 5.9 Schematic description of reverse osmosis

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Pure Water Applied Pressure

Semi-Permeable Membrane Salt Water

Fresh Water

Contaminants

Direction of Water Flow

5.5.2.6

Forward Osmosis

Forward osmosis (FO) involves the process of osmotic gradient in order to draw water from a low osmotic pressure towards a high osmotic pressure and can be seen in Fig. 5.10. The thermal process and reverse process are used to treat diluted draw solution in order to obtain pure water. Forward osmosis has two advantages as compared to pressure driven reverse osmosis: (i) It will not require high pressure and (ii) The layer is less prone to fouling. FO deals with a draw solute having high osmolality and easily distinguishable from water. At present, ammonia bicarbonate and NaCl are widely used chemicals to be utilized for draw solution. However, reverse osmosis and thermal treatment requires extensive energy to purify water. Recently, magnetic nanoparticles are investigated as new category of draw solute due to their reuse and easy separation. Dissolution was aid by employed hydrophilic coating in order to increase osmotic pressure. Permeate fluxes of forward osmosis were gained through the utilization of 0.065 M diacid coated Polyethylene magnetic nanoparticles and are as shown in Fig. 5.10. It is investigated that draw solutes were recovered through the utilization of magnetic nanoparticles. Recently, Fe3O4@SiO2 magnetic nanoparticles have been utilized in order to recover draw solute (Al2 (SO4)3) by flocculation (Liu et al. 2011).

5.5.2.7

Advance Oxidation Process

Advance oxidation processes (AOP) are a set of chemical treatment methods that are used to remove organic materials from wastewater by oxidation through chemical reactions with hydroxyl radicals (OH). The chemical processes can be controlled as

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PERMEATE

FEED

SALT WATER

WATER

SALT WATER

SALT

Organics, minerals and pollutants Membrane of Proprietary

Fig. 5.10 Description of forward osmosis mechanism

well as preceded by Ozone (O3), hydrogen peroxide (H2O2) and UV light. OH reacts and often made then pollutants will be quickly and efficiently converted to small inorganic molecules. The reactive species like hydroxyl radicals are strong oxidants that can oxidize any pollutant present in wastewater matrix and hence purifies water. The advance oxidation processes are partially useful for cleaning biological agents or non-degradable materials, such as aromatic pesticides, petroleum, and volatile organic compounds in wastewater. Advance oxidation process can be used to treat secondary treated wastewater which is shown in Fig. 5.11a–c. The pollutants are converted into stable inorganic compounds such as water, carbon dioxide, and salts. In AOP the reduction of chemical pollutants is performed to undergo mineralization of wastewater. AOP is considered as aqueous phase oxidation process which depends upon intermediacy of highly reactive species in the mechanism, i.e., hydroxyl radical for demolition of targeted pollutants. The efficiency of oxidation process is determined by different parameters such as concentration of catalysts, reaction contact time, intensity and wavelength of UV radiations, initial pH of solution hydrogen peroxide and structure of organic compounds (Zheng 2011). Fenton-type advance oxidation process involves Fe2+ and hydrogen as catalyst that leads to the production of hydroxyl radicals by catalytic mechanism. The chemical reactions involved are represented as (Martínez et al. 2005).

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Fig. 5.11 (a) Schematic representation of AOP classification, (b) mechanism of activated H2O2 in sludge minimization Kudo and Miseki (2009) and (c) destruction of pollutants using advance oxidation process (Ødegaard 1992)

Fe2þ þ H2 O2 ! Fe3þ þ HO þ HO Fe

3þ

þ

2þ

þ H2 O2 ! H þ ½FeðOOHÞ 2þ

½FeðOOHÞ

! Fe

2þ

þ HO2

HO2 þ Fe3þ ! Fe2þ þ Hþ þ O2

ð5:9Þ ð5:10Þ ð5:11Þ ð5:12Þ

An additional number of hydroxyl radicals are generated in photon Fenton process through interaction of UV radiation with iron species in aqueous solution as well as direct photolysis of (H2O2). The main chemical reaction is represented as (Segneanu et al. 2013): H2 O2 þ UV ! 2HO

ð5:13Þ

Fe3þ þ H2 O þ UV ! Fe2þ þ Hþ þ HO

ð5:14Þ

½FeðOHÞ2 þ UV ! Fe2þ þ HO

ð5:15Þ

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Nano-Sensors and Monitoring

The main problem for wastewater or water treatment system is monitoring of water quality in the presence of low concentration pollutants, high complexity of wastewater or water matrices and lack of fast pathogens detection.

Pathogen Detection The detection of pathogens has potential importance due their tremendous effect on public health. The conventional indicator systems are slow and unable to monitor the presence of effective pathogens and viruses such as Helicobacter and Legionella, Norwalk viruses, echoviruses, hepatitis A and E, coxsackie viruses and hepatitis E and A as well as protozoan. Many of these pathogens are etiologic operators in flareups related with drinking water. Also, pathogen recognition is the key part of analysis based water sterilization approach, in which sanitization is activated by the location of target microorganisms. Current research is going to develop nano-sensors, which has three main components: (i) signal transduction mechanism, (ii) recognition agent, and (iii) nanomaterials. The Recognition agent provides selectivity through their interaction with epitopes and antigens on the surface of pathogens. Fast response and selectivity was attained via nanomaterials, which is related with signal transduction upon the recognition event. A broad range of recognition agents have been used such as antimicrobial peptides, carbohydrates, aptamers, and anti-bodies. Nanomaterials enhance the speed of detection and sensitivity to accomplish multiplex target identification which incorporates to their novel physicochemical properties, particularly magnetic and electrochemical and optical properties. These sensors can be utilized to identify entire cells and in addition biomolecules. CNTs, noble metals and dye-doped nanoparticles, quantum dots and magnetic nanoparticles are mostly used for the detection of pathogens. Sample concentration and purification was analyzed by using CNTs and magnetic nanoparticles. Dynabead is available kits (commercial magnetic nanoparticles) for the detection of various pathogens. Quantum dots are known as fluorescent nano-crystal of semiconducting material such as CdSe, whose electrical properties depend upon the shape and size of individual crystal. The quantum dots with wider band gap and smaller size, stable and narrow fluorescent spectra and broad absorption spectra, which requires more energy to excite and emit light of shorter wavelength. Therefore, quantum dots are best candidate for the multiplex detection through one excitation light source (Theron et al. 2010; Vikesland and Wigginton 2010).

Trace Contaminant Detection Organic and inorganic pollutants can be traced and detected by the nanomaterials. Among nanomaterials, the Carbon Nanotubes (CNTs) have intrinsic adsorption and

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recuperation properties for inorganic and organic materials. They offer high rate of adsorption and recuperation with exceptional kinetics. Different other nanomaterials such as gold nanoparticles and quantum dots have also been utilized. Gold nanoparticles were used to separate the pesticides at part per billion (ppb) levels in a colorimetric analysis. The modified gold nanoparticles were applied in identification of Hg2þ and CH3Hgþ quickly with high affectability and selectivity. Fluorescence resonance energy transfer based detection used the quantum dots modified TiO2 nanotubes which showed detection limit of PAHs to the level of pico-mole per liter. A nano-sensor based on quantum dots of CoTe was immobilized on a glassy carbon cathode surface and was applied in detection of Bisphenol at different concentrations in water as low as 10 nM within 5 s.

5.5.2.9

Ultrafiltration

Water has the capacity to dissolve and contain different substances. Fresh water from surface water or groundwater is used for mechanical or residential reason, either for consumable or non-consumable utilization. Because of the anticipated purposes, a water treatment plant is ordinary to satiate the prerequisites of treated water. As a rule, regular water treatment plant comprises of physical treatment (screening, sedimentation, buoyancy, and filtration) and chemical treatment (pH alteration, coagulation-flocculation process, oxidation-decrease process, adsorption process). The complexity of treatment plant also depends upon the nature of crude water and treated water prerequisite. In industrial process, water is utilized as part of various applications, i.e., cooling water, water for flushing and compound generation, purified water, boiler food water for vaccination, etc. The distribution and cost of treatment were increased with growth of population. Ultrafiltration (UF) is considered as competitive treatment as compare to conventional methods. The formation of clear and gleaming water is sheltered extent whose infection is concerned. A complex control system was achieved through the optimal performance of overall process and controlling each step of this method. Nowadays, various pathogenic pollutants such as bacteria and viruses are discharged by membrane. Nowadays, ultrafiltration is utilized to depose in conventional water treatment system, i.e., coagulation, sedimentation, and filtration, which is known as disinfection membrane operation. All pollutants, i.e., macromolecules, viruses, and bacteria, are discharged through ultrafiltration due to permeability of membrane. The main advantages of ultrafiltration membrane as compared to conventional methods are: (i) It does not require chemicals for processes, (ii) constant and excellent quality of treated water, (iii) size-exclusion filtration as opposed to media depth filtration, (iv) good and constant quality of treated water in terms of particle and microbial removal regardless of raw food water quality, (v) process and plant compactness, and (vi) simple automation.

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Nano-Filtration

Recently, pressure driven membrane process is known as nano-filtration and developed for separation of liquid phase in aqueous solution. The characteristics of nanofiltration lie between porous ultrafiltration membranes and non-porous reverse osmosis membrane. In many applications, nano-filtration substitutes reverse osmosis because of higher flux rates and lower consumption of energy (Cadotte et al. 1988; Gozalvez et al. 2002). Commercially developed nano-filtration membranes have constant charge, which is produced through the dissociation of carboxyl acid or sulfonated (surface groups). Nano-filtration membrane permits the separation of ions through the reverse osmosis ion interaction mechanism and combination of ions as well as ultrafiltration electrical effects. The nano-filtration membrane is the comparatively new presented innovative technique in wastewater treatment system. The monovalent ions were reasonably transmitted by nano-filtration membrane due to surface electrostatic properties and the size of membrane nominally 1 nm in order to discharged uncharged solutes with conserved multivalent ions. Selective removal of solutes and fractionation from complex system was discharged by nano-filtration membrane due to its novel properties. Recently, the advancement of nano-filtration innovation has adorable process in many applications, i.e., demineralization in the dairy industry, pulp-bleaching effluents from the textile industry, virus removal and metal recovery from wastewater, and separation of pharmaceuticals from fermentation broths. Nano-filtration is one of the promising techniques for the treatment of inorganic pollutants and natural organic matter in surface water. As the surface water has low osmotic pressure, a low-pressure activity of NF is conceivable. There is a huge discharge of organic substances, i.e., (disinfection by-product precursors) by nano-filtration. Inorganic salts could be discharged through the charge effect of ions and membrane, whereas organic compounds have large molecules as compared to pore size of membrane discharged by separating mechanism (Bowen and Welfoot 2002; Tsuru et al. 2000; Van der Bruggen et al. 2002). A comparative description of the processes is presented in Fig. 5.12.

5.5.2.11

Novel Applications of Nanomaterials for Wastewater Treatment

Antimicrobial nanomaterials are potential candidate for wastewater treatment system. Nano-Ag has great potential for application in point-of-use (POU) water treatment. It can enhance water quality for top of the line utilization, or give another obstruction against aquatic pathogens for exposed population. Nano silver particles are used in many commercial available devices such as Aquapure and Marathon water system. The nano silver particles were integrated as barrier into ceramic microfilter for pathogens, which could be quite beneficial in developing countries for drinking water purification. CNT filters have been used for discharge of viruses and bacteria from water system, due to its peculiar properties, i.e., high conductivity,

5 Nanotechnology: An Innovative Way for Wastewater Treatment and Purification

Microfiltration

Ultrafiltration Suspended Particles

Macro & micromolecules

Membrane

Water

127

Membrane

Macro & micromolecules

Water High added value solutes

High added value components

Nanofiltration

High molecular weight

Multivalent Ions

Membrane

MF: 100 - 10,000 nm UF: 2 - 100 nm NF: 0.5 - 2 nm

Water Low molecular weight solutes

Fig. 5.12 Schematic description of microfiltration, ultrafiltration, and nano-filtration

Fig. 5.13 (a) Mechanism of pollutants removal from wastewater by using nanomaterials (Simeonidis et al. 2016), (b) Mechanism of pollutants removal from wastewater through the production of photo induced charge carriers in surface of TiO2 nanopartilces

fibrous shape, and antimicrobial properties. Virus and bacteria are effectively removed through thin layer of CNTs through size exclusion and depth filtration. The examples of the nanomaterial usage for wastewater treatment are shown in Figs. 5.13 and 5.14.

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Fig. 5.14 (a–b) Water purification systems using supported Ag nanoparticles into biopolymerreinforced synthetic granular nano-composites (Simeonidis et al. 2016), (c) Ag nanoparticles impregnated bacterialcidal paper, (d) woven fabric microfiltration gravity filter and (e) doped paper of Ag nanoparticles

5.6

Conclusion

Clean water is vanishing due to increase in water pollution; therefore, wastewater treatment and purification techniques are emerging and becoming popular due to increasing demand of clean water. Conventional methods are being used but are ineffective for the treatment of wastewater and purification and are not capable in complete removal of pollutants. Therefore, development of novel and cost-effective methods for the treatment of wastewater and purification are the contemporary demand of the nations. Nowadays, many advance methods based on nanotechnology are available for treatment of wastewater and purification including water splitting, reverse osmosis, advance oxidation process, advance membrane process, nano-sensors, photocatalysis, nano-filtration, etc. These multifunctional and highly efficient advance processes are providing efficient solution to wastewater treatment and purification through removal of metal ions, heavy metals, viruses, nitrates, natural organic matters and complex compounds, toxic organic and inorganic substances, which are present in water. Hence, nanotechnology is considered as advanced and emerging technology for wastewater treatment and purification through the removal of all kinds of pollutants.

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Chapter 6

Immobilized Nanocatalysts for Degradation of Industrial Wastewater Jayaseelan Arun, Marudai Joselyn Monica, Vargees Felix, and Kannappan Panchamoorthy Gopinath

Contents 6.1 6.2 6.3 6.4 6.5

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Industrial Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Nanocatalyst . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Immobilization of Nanocatalysts . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Applications . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.5.1 Degradation of Industrial Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.5.2 Detoxification of Industrial Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.5.3 Copper Adsorption of Industrial Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.6 Future Perspectives . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.7 Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

6.1

133 134 136 137 138 138 139 140 140 141 142

Introduction

Industrialization lead to economic growth and sustainability but also resulted in environmental pollution and global warming. Industrial wastewater mainly contains toxic and biodegradable heavy metals like zinc, nickel, silver, lead, chromium, arsenic, cadmium, and iron (Dil et al. 2017b; Cheng et al. 2017). Electroplating, metal finishing, and plastics release huge amount of copper into water bodies. Recently many researchers have reported various methodologies like membrane filtration, electrodialysis, Photocatalysis, and adsorption for removal of copper J. Arun · M. J. Monica · V. Felix · K. P. Gopinath (*) Department of Chemical Engineering, SSN College of Engineering, Kalavakkam, Tamil Nadu, India e-mail: [email protected] © Springer Nature Switzerland AG 2019 R. Prasad, K. Thirugnanasambandham (eds.), Advanced Research in Nanosciences for Water Technology, Nanotechnology in the Life Sciences, https://doi.org/10.1007/978-3-030-02381-2_6

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Table 6.1 Merits and demerits of various treatment regimes S. no 1

Treatment method Membrane filtration

2

Photocatalysis

3

Adsorption

4

Electrochemical

Merits Low operating conditions (pressure) Highly selective process Lesser space required Lesser by-product formation Organic contaminants were removed effectively Low cost and simple design Highly selective Eco-friendly

Demerits High cost

Lengthier process

Limited to metal concentration High cost Energy consumption was higher

ions from wastewater (Xu et al. 2016; Dong et al. 2017; Kanakaraju et al. 2017; Andrejkovicova et al. 2016). Treatment of wastewater by traditional methods was not up to the quality standards due to emerging of various contaminants (Ferroudj et al. 2013). Table 6.1 elaborates the merits and demerits of various treatment regimes present worldwide. Biological wastewater treatment was slow and limited due to presence of non-biodegradable contaminants (Zelmanov and Semiat 2008). Physical process (filtration) removes contaminants by transforming from one phase to another. Among the various techniques nanotechnology showered its vision due to incredible wastewater remediation potential and solution to environmental issues (Gupta et al. 2015). Nano-adsorbents, nanocatalysts, and nano-membranes were used in wastewater treatment process. Recovery and reuse of nanoparticles from the effluent on industrial scale is tedious and impossible. The degradation of pollutants from wastewater on industrial scale was a difficult process. In this regard, immobilization of catalyst was essential to make the process eco-friendly and degradation of pollutants from wastewater (Srikanth et al. 2017). Glass substrates, metallic supports, and photo-electrodes were used as supporting materials to immobilize catalysts. Natural fibers like chitosan and Luffa cylindrica were also used for catalyst immobilization (Nadarajan et al. 2016).

6.2

Industrial Wastewater

Without a doubt, industrialization was fundamental for the economic development of the country (Lade et al. 2015; Dai et al. 2016). Besides, the discharge of various contaminations with or without partial treatments are harmful to nature, particularly dye-based toxic wastewater effluents (Zucca et al. 2016; Bilal et al. 2016), and this is a noteworthy ecological concern. The scientific literature has demonstrated that every year tons of dyes are produced and about 10–15% of them get released into

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the water bodies in untreated condition because of the low budget of textile industry outcomes (Robinson et al. 2011). Notwithstanding the material business, dyes are likewise being utilized as a part of numerous enterprises including paper and pulp, food, cosmetic, pharmaceutical, tannery, photographic, and plastic industries. Ramírez et al. (2015) reported that various extents of blends results in cancercausing, genotoxic and mutagenic nature of it. Specialists of the fundamental water streams revealed that the balance of characteristic biological community in the streams was endangered (Punzi et al. 2015; Salleh et al. 2011), which in this way will cause health-related issues. Hence, the compelling treatment of dye-based effluents without creating any secondary contamination is vital to hinder environment disintegration. The danger of industrial effluents is fundamentally associated with the dyes what’s more, their breakdown intermediates. In this way, it is of incredible incentive to assess the toxicity status of untreated and treated dyes/effluent tests as an extra bio-treatment effectiveness instrument, since these dye pollutants, as well as their degradation metabolites, are perceived to be a potential danger to human health and aquatic biota. It has been realized that consistently expanding arrival of dye-based textile effluents may cause anomalous pigmentation of the water surface specifically influencing the amphibian widely varied vegetation, and this captures the squeezing logical concern concerning environmental safety and water quality. Various azo dyes apply different direct and indirect impacts that can prompt the allergies, cancers, and tumors arrangement (Pakshirajan et al. 2013). Unfortunately, insufficiency of satisfactory toxicity reports/figures for pollutants with respect to cell populaces makes bioremediation (e.g., degradation, mineralization, or decolorization) hesitant and untrustworthy for on location activity (Kalme et al. 2007; Saratale et al. 2009). The most widely recognized substantial metals that are regularly present in industrial wastewater include nickel, zinc, silver, lead, press, chromium, copper, arsenic, cadmium, and uranium (Dil et al. 2017a; Cheng et al. 2017). In any case, copper is generally found at high concentration in wastewater, since it is considered as the most profitable and generally utilized metal in numerous industrial applications, for example, metal completing the process of electroplating, plastics, and scratching (Azad et al. 2016; Zou et al. 2016). Besides, copper is an extremely harmful metal even at low concentration, and copper-contaminated wastewater must be treated before releasing it to nature (Nassef and El-Taweel 2015; Yang et al. 2015). The allowable farthest point of copper particles in industrial effluents was detailed by the United State Environmental Protection Agency (USEPA) to be 1.3 mg/L while it was expressed by World Health Association that copper particles content in drinking water should not be more than 2 mg/L (Aydın et al. 2008).

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J. Arun et al.

Nanocatalyst

Nanoparticles photocatalytic reaction depend on collaboration of light energy with metallic nanoparticles and are of incredible interest because of their wide and high photocatalytic exercises for different pollutants (Akhavan 2009). Typically these photocatalysts are involved semiconductors metals that can degrade assortment of persistent organic pollutants in wastewater, for example, dyes, detergents, pesticides, and volatile organic compound (Lin et al. 2014). Besides, semiconductor nanocatalyst is likewise effective for degradation of halogenated and nonhalogenated organic compounds and furthermore heavy metals in particular situation (Adeleye et al. 2016). Semiconductor nanomaterials required mild operating conditions and are extremely compelling even at a small concentration. The simple mechanism of the working of photocatalysis depends on the photoexcitation of electron in the catalyst as shown in Fig. 6.1. The light with light (UV if there should be an occurrence of TiO2) produces gaps (h+) and left electrons (e ) in the conduction band. In an aqueous media, the holes (h+) are caught by water molecules (H2O) and produce hydroxyl radicals ( OH). The radicals are unpredictable and intense oxidation agent. These hydroxyl radicals on reaction oxidize the organic pollutants into water and gaseous degradation products (Akhavan 2009). Among different nano-photocatalysts created up till now, TiO2 is a standout among the most generally connected in photocatalysis because of its high reactivity under ultraviolet light (k < 390 nm) and synthetic security. Correspondingly, ZnO has additionally been examined for its photocatalytic activity, as it contains wide band hole like TiO2 (Lin et al. 2014; Bhuyan et al. 2015). Various studies have

Fig. 6.1 Photocatalytic mechanism of action

6 Immobilized Nanocatalysts for Degradation of Industrial Wastewater

137

demonstrated the photocatalytic movement of different synthetic catalysts. Their productivity is contingent upon various factors, for example, band gap energy, particle size, dosage, toxin fixation, and pH. For example, Hayat et al. (2011) found that the photocatalytic degradation productivity of ZnO diminished with high calcination temperature that expands the molecule estimate due to agglomeration. Cds is additionally a notable semiconductor having band hole of 2.42 eV and can be worked at wavelength Br
> I
. RO has been tested in the removal of a wide range of pollutants and has proven highly effective to remove halide ions. Magara et al. (1996) eliminated around 99% of the baseline Br
ion concentration from sea water using RO with aromatic polyamide spiral modules. Percentage I
ion removal of 93.0–99.3% have been obtained, depending on the membrane type and experimental conditions used, and rates of 71–90% were reported by Wintgens et al. (2005). However, membranes are easily silted, shortening their useful life, and a pretreatment stage is required for turbidity reduction and microbial control. A further drawback of RO is its high cost, given the need for water pretreatment and the elevated energy consumption required to maintain the high operating pressures. Furthermore, management of the resulting concentrate poses an environmental challenge. Recent investigations demonstrated that mechanical alterations in the membrane structure at different pressures influence the attachment of organic matter to the membrane and its reversibility (Xie et al. 2015). The above shortcomings have limited the generalized application of RO in water treatment plants.

7.2.1.2

Nanofiltration

NF includes both RO (non-porous diffusion) and ultrafiltration (pore sifting) and is an intermediate process between them (Fane et al. 2011; García-Vaquero et al. 2014). NF generally operates at a lower work pressure in comparison to RO, reducing its cost. Different types of membranes are effective to separate dissolved organic matter, small organic molecules (pesticides, pharmaceutical compounds, endocrine disruptors, etc.), cations, and/or salts (Chon and Cho 2016; Darwish et al. 2007; Laurell et al. 2015; Maher et al. 2014; Vergili 2013) according to their cutoff molecular weight. In a water treatment plant, the utilization of NF membranes

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reduced THMs by 90% (Ribera et al. 2013). A study on the effect of NF on monovalent ions reported a maximum percentage I
ion removal of 80% using an NF-70 membrane at a pressure of 1200 kPa (Lhassani et al. 2001). The authors observed that the membrane was more selective for Cl
ions than for I
ions, although the opposite was observed at pressures >800 kPa, demonstrating the important influence of operating conditions on the selectivity of membranes for monovalent ions. Diawara et al. (2003) achieved the selective separation of monovalent halides using an NF-45 membrane, whose effectiveness for halide ions followed the order (Hofmeister series) F
> Cl
> Br
> I
, finding that the membrane convection coefficient and solute permeability were correlated with the hydration energy of halide ions. Llenas et al. (2011) also observed the retention of halides according to the Hofmeister series at low pressures, although the order of Cl
and Br
was reversed at high pressures. These authors also used NF as pretreatment system for RO and found that it reduces the formation of incrustations on the membrane. The cost of NF is slightly lower in comparison to RO, given the lower work pressures used, enabling the recovery of a higher water flow, while its effectiveness for ion removal is comparable (Pontié et al. 2008; Zhou et al. 2015). Accordingly, its use has increased, mainly in industrial applications and pretreatment systems for water intended for human consumption. Bórquez and Ferrer (2016) recently observed that 99.9% of dissolved solids and Na+, Mg2+, Ca2+, and Cl
ions could be removed by combining the NF system using membrane NF90-2540 with a RO and/or ion-exchange stage. Finally, López-Roldán et al. (2016) reported that the implantation of membrane systems in a water treatment plant improved the chemical quality of water, reducing its carcinogenic index.

7.2.1.3

Electrodialysis and Reverse Electrodialysis

In ED, a continuous electrical current forces the passage of ion species through membranes with opposite charge. The removal rate is directly proportional to the current applied and inversely proportional to the flow rate through each cell pair. The membranes are formed by an ion-permeable porous polymer matrix. According to published results, ED can concentrate around 70–245 g L
1 NaCl, achieving additional purification of the main multivalent ions (Ca2+, Mg2+, and SO42
) through ion complexation reactions in concentrated brines (Reig et al. 2014). RED is a modification of ED in which the electrode polarity is periodically reversed (Khan et al. 2013). The membranes permit transporting anions (in the case of anion-exchange membranes) or cations (in the case of cation-exchange membranes), producing an electric potential difference. Ion transport is subsequently compensated by electron transport due to redox reactions in the electrodes. The electric current generated can be used in an external electric device. RED has been successfully used in water desalination, providing a reliable and inexpensive alternative to salty water RO. Thus, Valero and Arbós (2010)

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implemented the RED system in a pilot plant and a water desalination plant, reporting that the pilot plant achieved a reduction of 75% in Br
ions, 60% in Cl
ions, 30% in total organic carbon, 75% in NO3
ion concentration, 70% in SO42
anion concentration, 80% in Ca2+ ions, 70% in K+ ions, and 8% in Mg2+ ions, with a water recovery rate of 90%. The formation of THMs was considerably reduced by the decrease in halide ions and NOM, with final concentrations of 64–60 μg L
1, markedly lower than maximum of 100 μg L
1 established by Spanish legislation (RD 140/2003). Van der Hoek et al. (1998) tested RED as an alternative to filtration with RO in three integrated membrane systems (IMS) that used ozone as a disinfection process, obtaining a reduction of 72% in Br
ion concentration before the disinfection stage, with a resulting decrease of 5 μg L
1 in formation of the bromate ion. The utilization of monovalent ion selective membranes permits the separation of mixtures of both monovalent and multivalent ions. This relative selective permeability is obtained by depositing a very thin layer on the surface of conventional membranes, permitting only the passage of monovalent ions (e.g., Cl
) and restricting the passage of divalent ions (e.g., sulfate). This layer can also have an opposite charge to that of the membrane, producing selectivity towards monovalent ions (Güler et al. 2014).

7.2.1.4

Advantages and Disadvantages of Membrane Techniques

Polymer membranes undergo swelling, biological incrustation, and desquamation, and their thermal and chemical resistance is poor, limiting their operating life and hampering their cleaning (Duscher 2014). For this reason, porous inorganic membranes are currently used, due to their superior thermal stability, resistance to solvents and chemical products, and greater mechanical resistance, allowing their utilization under greater pressures and with a wider pH range (König et al. 2014; Zhu et al. 2015). ED-treated waters require minimum pretreatment and offer greater water recovery in comparison to RO, but new types of membranes are needed to improve their effectiveness and prolong their useful life. In addition, neither ED nor RED can eliminate neutral species such as dissolved organic matter, and both require a previous stage to remove these precursors.

7.2.2

Electrochemical Techniques

Electrochemical techniques have been successfully used to remove pollutants from industrial wastewaters and natural waters (Dong et al. 2017; Gabarrón et al. 2016; Strathmann 2010). The most widely applied electrochemical techniques are electrolysis and capacitive deionization (CDI).

7 New Technologies to Remove Halides from Water: An Overview

7.2.2.1

153

Electrolysis

In electrolysis, the passage of an electric current through an electrolytic solution gives rise to oxidation-reduction reactions that produce the removal of chemical compounds. Kimbrough and Suffet (2002) studied the removal of Br
ion from natural waters with electrolysis, which allowed oxidization of the Br
ion to a mixture of hypobromite, hypobromous acid, and bromine gas that could be eliminated by degasification with carbon dioxide, reducing the post-disinfection concentration of brominated THMs. Bo (2008) proposed the combination of disinfection with Br
ion removal by electrolysis, utilizing silver cathodes and anodes in parallel, finding that Cl
ions were oxidized into chlorine gas, which acts as a disinfectant, and Br
ions were oxidized into bromine gas, which volatilizes and therefore does not require extraction. The effectiveness of this process depends on the [Cl
]/[Br
] ratio, electrical current applied between electrodes, ionic strength of the water, distance between electrodes, and residence time. It removes 48–62% of baseline Br
ion concentrations when [Br
]0 is 200 μg L
1. The final result is a 73% reduction in the concentration of brominated THMs.

7.2.2.2

Capacitive Deionization

CDI is a promising technique for water desalination, based on electro-adsorption at low cell voltage (0.6–1.2 V) using a pair of juxtaposed highly porous electrodes (Suss et al. 2015). Ions are removed from water by the electrical current and temporarily stored in the double electrical layers formed within the porous electrodes (Edmunds and Smedley 2013). It was recently described as a low-cost and readily implemented method to remove heavy metals (Huang et al. 2016b). The utilization of flow electrodes allows electrode regeneration without halting the system, increasing the efficiency of the process (Suss et al. 2015). Liu et al. (2016) used CDI to remove DBP precursors, Br
ions, and dissolved organic matter, and achieved a reduction of 1160–1510 μg L
1 in brominated DBPs formed, supporting the consideration of CDI as an alternative option for controlling DBPs in water treatment systems.

7.2.2.3

Membrane Capacitive Deionization

In membrane CDI (MCDI), ion-exchange membranes are placed between the carbon electron and the water to be treated, hindering the passage of ions from the electrode area to the water and increasing the efficiency of the process (Bian et al. 2016; Dykstra et al. 2016).

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7.2.2.4

J. Rivera-Utrilla et al.

Advantages and Disadvantages of Electrochemical Techniques

CDI and MCDI are not currently implemented to treat water intended for human consumption; however, this technology may become a feasible alternative after the development of new electrodes and membranes needed to operate at an industrial scale.

7.2.3

Adsorption Techniques

The implementation of adsorption methods for halide ion removal in water treatment plants is relatively easy and their operating costs are low. The different adsorption techniques used are based on the use of hydrated oxides, coals, activated carbons, silver-doped activated carbons, carbon aerogels, ion-exchange resins, and aluminumbased adsorbents, as described below.

7.2.3.1

Hydrated Oxides

Hydrated oxides can possess cationic or anionic adsorbent properties according to the basicity of the central metal atom, and their characteristics depend on the method used for their synthesis. Synthesis techniques are simple to apply and inexpensive, and the materials obtained have excellent selectivity for certain ions or ion groups (Chubar 2011).

Layered Double Hydroxides Increasing interest is being paid to layered double hydroxides (LDHs), also known as hydrotalcites, because of their capacity for selective pollutant removal from waters (Wang and O’Hare 2012; Wu et al. 2013). They are inorganic materials formed by double layers of divalent and trivalent cations, whose charges are neutralized by anions in the matrix interlayer, which act as exchangeable anions. They have an extensive surface area and numerous ion-exchange sites, allowing them to act as adsorbents and ion exchangers. Furthermore, they can be obtained from low-cost, easily regenerated precursors. Table 7.4 lists some examples of LDHs used to remove halide ions from waters. As can be seen in Table 7.4, LDHs have proven as effective ion-exchange resins in the removal of Br
and I
ions from natural waters, with a high ion-exchange rate. The main drawback of LDHs is the preferential adsorption of bicarbonate and sulfate ions over Br
ions.

S

S S S S

I
(mg L
1) 342

330

– 100 100 100

1 1 1 – –

1269

Br
(mg L
1) –

–

100 – – 100

– – – 0.2 0.2

–

Mg-Al-(NO3)

Zn-Al

Mg-Al Mg-Alb Mg-Alc Mg-Ald

Cu6Al2(OH)18 Cu6Cr2(OH)18 Cu6Ga2(OH)18 Mg-Fe Mg-Al-Fe

Zn-Al

YES

YES YES YES YES YES

NO NO NO NO

NO

Competition with other anions YES

–

– – – 60 60

27.5 – – 94

–

Br
(%) –

b

a

14

9600 mL g-1 e 4200 mL g-1 e 1700 mL g-1 e – –

– 35 88.4 96.1

60

I
(%) 59

Halide removal

Water type: N ¼ natural water, S ¼ synthetic water, WW ¼ wastewater Molar proportion 2 c Molar proportion 3 d Molar proportion 4, calcined at 400  C e Adsorption coefficient, % removal not available

S

S S S N N

Water typea N, WW

Baseline anion concentration

LDH type

Table 7.4 Application of different LDHs for halide removal from waters (Watson et al. 2012)

10

10 10 10 10 10

1 1 1 1

–

48

24 24 24 0.16 0.16

24 – – 24

–

[Adsorbent] Time (g L
1) (h) 20 4

6

– – – 7–8 6.5–7.5

– – – –

7

pH 9.2

–

25 25 25 – –

30 30 30 30

–

T ( C) 25

Echigo et al. (2007) Kaufhold et al. (2007)

Kentjono et al. (2010) Thomas and Rajamathi (2009) Liang and Li (2007) and Lv et al. (2008) Pless et al. (2007)

References

7 New Technologies to Remove Halides from Water: An Overview 155

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J. Rivera-Utrilla et al.

Double Hydrated Hydroxide Sol-Gel Chubar et al. (2005) used the sol-gel method to synthesize a double hydrated oxide Fe2O3Al2O3H2O able to remove F
, Cl
, and bromate ions from the medium. In a later study, Chubar (2011) synthesized other double hydrated hydroxide with a Br
ion adsorption capacity of 20.25 mg L
1 for a baseline Br
concentration of 198.2 mg L
1.

7.2.3.2

Carbon Materials

Carbon materials are widely used to treat natural waters and wastewaters, and their capacity to remove halide ions has been investigated.

Coal and Activated Carbon Balsley et al. (1998) studied the I
ion adsorption capacity of lignite and sub-bituminous coals, finding the adsorption rates to be pH-dependent. Comparable I
ion adsorption results were observed on a commercial activated carbon by Kaufhold et al. (2007), who postulated that the adsorption mechanism consisted in I
ion oxidation to iodine, which is then adsorbed on the activated carbon.

Silver-Doped Activated Carbon Various authors have impregnated carbon materials with silver in an attempt to improve their halide ion removal capacity. Thus, Gong et al. (2013a) found that the capacity of silver-doped porous carbon spheres to adsorb Br
ions was both temperature-dependent, increasing from 80% at 20  C to 94% at 25  C, and pH-dependent, with a study pH range of 4.0–7.0 and optimal pH of 5.0, at which 90% of the baseline Br
ion concentration was removed. Watson et al. (2016) reported that the I
ion adsorption capacities of a silver-doped activated carbon and the same activated carbon with no modification were similar (25–26 μg L
1) but that the Br
ion adsorption capacity of the silver-doped material was much higher, removing 95% of the baseline concentration. Chen et al. (2017) also found that silver-doping of activated carbons markedly improved their capacity for Br
anion removal, independent of the synthesis method used. They concluded that the effectiveness for Br
ion removal of the doped material depended on: (i) the nature of the water matrix, with a reduction in percentage removal at higher concentrations of Cl
ions or dissolved NOM; (ii) the degree of silver doping, with an increase in percentage Br
ion removal with higher doping; and (iii) the physical and chemical characteristics of the original activated carbon surface, observing a higher percentage Br
ion removal with larger surface area.

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Silver Chloride-Doped Activated Carbon Ho and Kraus (1981) reported that activated carbon pretreated with silver chloride (20% in weight) removed 98.0% of I
ions, independent of the presence of Cl
ions. Karanfil et al. (2005) found a similar percentage I
ion removal between activated carbons doped with silver chloride or with silver alone but observed a much higher percentage silver lixiviation with the latter (up to 50%) than with the silver chloridedoped carbons (6%). According to these findings, activated carbon doped with silver chloride is a promising material for halide removal from waters.

Silver-Doped Carbon Aerogels Silver-doped carbon aerogels have also proven effective to remove halide ions from water (Sánchez-Polo et al. 2006, 2007). They have significantly higher halide removal capacity in comparison to activated carbon, achieving a maximum adsorption capacity of 7.32 μmol g
1 for Cl
, 1.98 μmol g
1 for Br
, and 3.01 μmol g
1 for I
. The adsorption capacity increased by 92% for Br
ions and 154% for I
ions when the aerogels were carbonized and activated.

7.2.3.3

Ion-Exchange Resins

Ion-exchange resins are solid synthetic water-insoluble materials in porous or gel form (zeolites, clays, etc.), which selectively exchange ions as a function of their chemical structure. Magnetic ion-exchange (MIEX®) resins have mainly been used to remove Br
ions and organic matter from natural waters, contributing to the removal of bromo-derivative precursors (Bazri et al. 2016; Boyer and Singer 2006; Ding et al. 2012; Hsu and Singer 2010; Karpinska et al. 2013; Phetrak et al. 2014), and their Br
ion removal capacity of these resins depends on the medium pH. These resins were found to reduce baseline Br
ion concentrations by 93% under optimal conditions and by 60% under less favorable conditions (Singer and Bilyk 2002). Gibert et al. (2017) recently found that a hybrid system combining MIEX® resin with ultrafiltration could remove dissolved organic carbon (DOC) and Br
ions from water, with an effectiveness that depended on the resin concentration, achieving 55% DOC removal and 37% Br
ion removal when 3 mL L
1 MIEX® resin was used. However, the percentage Br
ion removal decreased to zero at 8 h of treatment, attributed to the presence of chemical species that compete with Br
ions for resin binding sites. Watson et al. (2015) reported that Br
ion adsorption on MIEX® resin was low under high alkalinity conditions, likely due to competition between bicarbonate and Br
ions for binding sites on the anion-exchange resin; thus, the percentage Br
ion removal decreased from 49  4% at low pH values to 20  4% at higher pH values. In the case of I
anions, however, the relationship between their adsorption and alkalinity or DOC was not statistically significant, and higher percentage I
removal

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percentages were observed when the baseline I
concentration was elevated, suggesting that I
ion competes effectively with bicarbonate or DOC for MIEX® resin binding sites.

7.2.3.4

Halide Removal by Coagulation with Aluminum Salts

The use of aluminum salts has also been proposed to remove halide ions from waters. Ge et al. (2007) found that 93.3–99.2% of Br
ions were removed by coagulation with aluminum chloride and that the percentage removal decreased in the presence of organic matter but was independent of the pH in the range between 4.0 and 8.0. The removal capacity of aluminum chloride was lower in natural water due to interference from other anions and dissolved organic matter. Ge and Zhu (2008) studied the effect of the presence of other ions on percentage Br
ion removal by coagulation with aluminum salts, finding a decrease of 11.5% with the addition of HCO3
, 21.2% with SO42
, 14.6% with Cl
, 8.0% with NO3
, and 40.8% with H2PO4
.

7.2.3.5

Nanotechnology

Water treatment companies continuously improve coagulation processes or add nonselective adsorbents to reduce the presence of DBP precursors and achieve disinfection with minimum DBP generation. The search for new approaches has been stimulated by more restrictive legislation on maximum DBP concentrations in waters intended for human consumption (Qu et al. 2013). The development of nanotechnology has been responsible for novel approaches to water treatment and disinfection and other environmental problems based on the intrinsic characteristics of nanoparticles, including their large surface area, high reactivity, and surface plasmon resonance, among others. These properties have led to proposals for the utilization in water disinfection of silver nanoparticles as antimicrobials (Ahmed et al. 2016; Durán et al. 2016; Aziz et al. 2015) and of photocatalytic TiO2 (Gupta and Tripathi 2011) or iron oxide (Xu et al. 2012) nanoparticles due to their photocatalytic properties. It is worth to mention that silver and iron oxide nanoparticles contribute to the majority of publications as regard to their application in water treatment. The four major areas of application of nanomaterials in water treatment are (i) adsorptive removal, (ii) catalytic degradation, (iii) disinfection, and (iv) membrane filtration (Li et al. 2008; Prathna et al. 2018). Depending on the adsorption mechanism, nanoparticles can be designed to have high selectivity for specific pollutants, including heavy metals, microorganisms, and organic pollutants. For instance, iron nanoparticles have been prepared to remove heavy metals (Geng et al. 2009; Hua et al. 2012; Kaushal and Singh 2017; Zhou et al. 2014) and inorganic nanoparticles have been developed as biocides and/or antimicrobial agents (Frimmel and Niessner 2014; von Moos and Slaveykova 2014). In

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159

addition, silver nanoparticles with excellent antibacterial and antifungal properties have been designed for preventive disinfection in drinking water treatment infrastructures or in places with limited sanitation procedures and high risk of pathogen development (Simeonidis et al. 2016). Polo et al. (2017) studied the removal of halides from water by silver nanoparticles (AgNPs) and hydrogen peroxide (H2O2). The AgNP/H2O2 process proved efficacious for bromide and chloride removal from water through the selective formation of AgCl and AgBr on the AgNP surface. The removal of Cl
and Br
anions was more effective at basic pH, reaching values of up to 100% for both ions. The formation of OH• and O2•
radicals was detected during the oxidation of Ag (0) into Ag(I), determining the reaction mechanism as a function of solution pH according to the following reactions: Agn 0
NPs þ 2H2 O2 ! Agn 0
1
NPs
Agþ þ O2 •
þ 2H2 O Agn
1 0
NPs
Agþ ⇆Agn 0
1
NPs þ Agþ Agn
1 0
NPs
Agþ þ O2 •
! Agn

1
NPs
Agþ þ O2

ð7:6Þ ð7:7Þ

þ

ð7:8Þ

Agn

1
NPs
Agþ ! Agn
1 0
NPs
Ag0

ð7:9Þ

Agn





1

þ

ð7:5Þ


NPs
Ag ⇄Agn





1


NPs þ Ag

Agn 0
NPs þ H2 O2 þ Hþ ! Agn 0
1
NPs
Agþ þ OH • þ H2 O OH • þ H2 O2 ! HO2 • þ H2 O •

HO2 ⇄O2

•


ð7:11Þ

þ

þ H ðpK a ¼ 4:8Þ

ð7:12Þ

O2 •
þ O2 •
þ 2 Hþ ! H2 O2 þ O2 •

ð7:13Þ

•

HO2 þ HO2 ! H2 O2 þ O2 •

HO2 þ O2

•
•

þ H2 O ! H2 O2 þ O2 þ HO

HO2 þ O2

•


ð7:14Þ





! HO2 þ O2

H2 O2 þ HO
⇆HO2
þ H2 O ðpK 11 ¼ 11:8Þ O2

•


•




þ HO2 ⇄HO2 þ O2

HO2
þ Hþ ⇄H2 O2

ð7:10Þ

ð7:15Þ ð7:16Þ ð7:17Þ ð7:18Þ ð7:19Þ

Moreover, the results obtained showed that: (i) the efficacy of the oxidation of Ag (0) into Ag(I) is higher at pH 11 and (ii) the presence of NaCl and dissolved organic matter (tannic acid [TAN]) on the solution matrix reduces the efficacy of bromide removal from the medium due to: (1) precipitation of AgCl on the AgNPs surface and (2) the radical scavenger capacity of TAN. AgNPs exhausted were regenerated by using UV or solar light, and toxicity test results showed that AgNPs inhibit luminescence of Vibrio Fischery bacteria.

160

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Fig. 7.1 XPS spectra of an Ag-cloth: (a) Survey scan and (b) high-resolution spectrum of Ag0 3d5/2 and 3d3/2 and Ag+ 3d5/2 and 3d3/2

The same authors (Polo et al. 2016) carried out a research on the behavior of a novel material, silver-doped polymeric cloth (Ag-cloth), in the removal of bromides and iodides from water adding H2O2. Ag-cloth was prepared using an electrospinning methodology to transform methacrylic-based polymer in the cloth material. After cloth generation, Ag nanoparticles were included by specific interactions between the polymer chemical groups and Ag atoms. UV radiation of Ag (I) cloth was used for reducting Ag(I) to Ag(0) on material surface. The surface chemistry of the Ag-cloth was determined by XPS, obtaining a general spectrum (Fig. 7.1a) that confirmed the presence of Ag on the surface of the Ag-cloth. Figure 7.1b depicts the high-resolution spectrum corresponding to the energy of the Ag bond, formed by two individual peaks, with a bond energy of 368.0 and 374.0 eV. Each peak can be deconvoluted into two peaks, thereby examining the contribution of Ag0 and Ag+. The peaks assigned to Ag0 constitute 95.05% of the total duplet area, with only 4.95% corresponding to Ag+ (Fig. 7.1b). Results indicated that Ag0 reacts with H2O2 in the first phases of the process, yielding Ag+, superoxide radical, and water; however, as the process advanced, this radical favored Ag+ reduction (Reactions 7.7, 7.8, and 7.9). Figure 7.2 shows the results obtained for halide chemisorption as a function of initial H2O2 concentration. The removal capacity of the Ag-cloth was higher for iodide than for bromide ions, which is due to the higher affinity of I
forwards Ag+ (lower Ksp), compared to Br
(higher Ksp). An increase of the H2O2 concentration increased the capacity of the Ag-cloth to remove halides up to a maximum concentration of 55 μM. For higher H2O2 concentration a slight decrease in adsorption capacity was observed. This is due to the reduction of Ag+ into Ag0 by action of the superoxide radical (Reactions 7.7, 7.8, and 7.9). The halide removal capacity of the Ag-cloth was analyzed as a function of solution pH (Fig. 7.3). The halide removal capacity of the Ag-cloth increased from pH 0.5 to 5.0 because of the increasing reactivity between H2O2 and Ag0 (Wang et al. 2013). The halide removal capacity of the material decreased at pH > 7.0. This behavior can be explained by considering Reaction (7.18). The superoxide radical disproportionation ( Reaction 7.18) rate depends on the medium pH and decreases at

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161

Fig. 7.2 Influence of the initial H2O2 concentration on halogen adsorption capacity (Xm). Halide: (■), Br
; (●), I
. [I
]0 ¼ 2.5  10
5 M; [Br
]0 ¼ 2.5  10
5 M; [H2O2]0 ¼ 25–75 μM; pH ¼ 6.5  0.5; WAg-cloth ¼ 0.022  0.002 g

Fig. 7.3 Influence of the medium pH on the halide adsorption capacity (Xm). Halide: (■), Br
; (●), I
. [I
]0 ¼ 2.5  10
5 M, [Br
]0 ¼ 2.5  10
5 M, [H2O2]0 ¼ 5.5  10
5 M, pH ¼ 0.5–12.0, WAg-cloth ¼ 0.0215  0.0015 g

higher pH values, according to the relationship k10 ¼ 5.0  1012 [H+] M
1 s
1 established previously (He et al. 2012). Consequently, the rise in solution pH increased the concentration of the superoxide radical, favoring surface Ag+ reduction by Reactions (7.7), (7.8), and (7.9) which reduced the halide removal capacity of the Ag-cloth. Recently, Polo et al. (2018) proposed the use of magnetic microparticles with Ag (0) on their surface (Ag-MPs) and their subsequent oxidation with H2O2 to remove

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Fig. 7.4 Influence of TAN on Br
and Cl
ion removal by the Ag-MPs/H2O2 system. [Ag-MPs]0 ¼ 100 mg L
1, [H2O2]0 ¼ 1.0  10
3 mol L
1, [Br]0 ¼ 1.0  10
3 mol L
1, [Cl
]0 ¼ 1.0  10
3 mol L
1, pH ¼ 7.0

Br
and Cl
ions from water through the formation of the corresponding silver halides adsorbed on their surface and their subsequent elimination by applying an external magnetic field. The main objectives of this research were: (i) to study halide (chloride and bromide) removal from water using Ag-MPs; (ii) to analyze the influence of operational variables (pH, initial H2O2 concentration, etc.) on halide ion removal from water; (iii) to evaluate the regeneration of Ag-MPs by solar or ultraviolet radiation; and (iv) to study the cytotoxicity of Ag-MPs. The results obtained pointed out that: (i) Halide ion removal from water can be achieved by using Ag-MPs after their oxidization with an oxidizing agent such as hydrogen peroxide. (ii) The effectiveness of the Ag-MPs/H2O2 system depended on the medium pH, being favored at acid and neutral pH values, and on the presence of dissolved organic matter, which interfered with Ag-MPs oxidization and, therefore, decreased the amounts of halides removed (Fig. 7.4). (iii) Ag-MPs can be regenerated with UV or solar radiation, although the efficacy of the Ag-MP/H2O2 system was reduced with each regeneration cycle. (iv) The cytotoxicity study indicated that Ag-MPs were not toxic at the concentrations studied.

7.2.3.6

Advantages and Disadvantages of Adsorption Techniques

In general, adsorption methods of halide removal have lower operating costs and are more readily implemented in comparison to membrane or electrochemical techniques. However, their effectiveness is decreased when halide ions must compete

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with organic matter or other ions for adsorption sites on the adsorbent material surface.

7.3

Technologies Used for Fluoride Removal from Water

In general, the presence of F
ions in waters intended for human consumption is a local problem; therefore, treatments for their removal are usually independent of those used to remove Cl
, Br
, and I
ions removal. Their presence is related to the geology of the area in which aquifers are located or to drought conditions that require the use of groundwater with high F
ion concentrations, often above the maximum limit of 4 mg L
1 set by the World Health Organization for drinking water (World Health Organization 2008). Fluoride ion concentrations can be elevated in water for human consumption if the groundwater is in contact with minerals such as fluorite [CaF2], apatite [Ca5(PO4)3(F, Cl, OH)], or cryolite [Na3AlF6], which can dissolve under certain conditions (Edmunds and Smedley 2013; Farooqi et al. 2007; Vithanage and Bhattacharya 2015). F
ion concentrations can also be high due to the reutilization of water from some industrial activities, which was found to be responsible for F
ion concentrations, >1.0 mg L
1 in the resulting drinking water (Ku et al. 2002; Mohapatra et al. 2009). Consumption of water with high F
ion concentrations can have severe effects on teeth and bones. Thus, prolonged exposure can cause tooth staining and fragility (dental fluorosis) and may also be responsible for more severe diseases in children and adults, including cancer, renal failure, digestive and nervous disorders, respiratory problems, Alzheimer’s disease, thyroid disease, and calcium and phosphorus metabolism disorders, as well as for the stunting of children’s growth and development (Barbier et al. 2010; Choi et al. 2015; Kut et al. 2016; Sajil Kumar et al. 2015; Zhang et al. 2016). Table 7.5 lists these effects as a function of the F
ion concentration with which they are associated (Dissanayake 1991). Drinking F
ion-contaminated water and the possible health consequences continue to be an important public health problem, especially in developing countries, which are less able to implement technologies to reduce F
ion concentrations to acceptable levels.

Table 7.5 Health effects of the consumption of water with F
ions

[F
] (mg L
1) 10.0

Disease developed Dental caries Optimal dental health Dental fluorosis Dental and bone fluorosis Paralyzing fluorosis

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Techniques used to remove F
ions from water include: reverse osmosis, nanofiltration, adsorption, coagulation/precipitation, ion exchange, and electrochemical and membrane processes.

7.3.1

Reverse Osmosis and Nanofiltration

The fertilizer industry generates effluents with high concentrations of F
ions. Dolar et al. (2011) studied the capacity of RO for their removal from synthetic waters and from fertilizer plant effluent, using six different membranes, and achieved >80% F
removal in synthetic waters and >96% removal in the effluent. Percentage removal rates of 98% were obtained by other researchers, regenerating the membrane after each cycle (Mohapatra et al. 2009; Richards et al. 2010; Sehn 2008). A hybrid system of precipitation with hydrated lime plus RO was used to remove F
ions from real industrial waters with a high F
ion content (3200 mg L
1), observing removal rates >98% after neutralization with hydrated lime at pH 7–8, which increased to approximately 99.9% after the RO treatment of optimally neutralized industrial effluent (Ezzeddine et al. 2015).

7.3.2

Adsorption Techniques

Membrane techniques diminish the presence of F
ions in waters but require high pressures and therefore elevated operating costs. Other methods to remove F
ions have been considered, including adsorption on different materials (e.g., metal oxides, metal hydroxides, carbon materials, membranes), which is one of the most effective, economic, and environmentally friendly techniques for F
removal (Biswas et al. 2007; Jamode et al. 2013).

7.3.2.1

Metal Oxides and Hydroxides

One of the simplest methods to remove F
ions is by adsorption on activated alumina. Craig et al. (2017) recently analyzed variations in its adsorption capacity as a function of different operating conditions (e.g., surface charge, hydration period, particle size). They reported that the adsorption capacity of activated alumina remained constant and high under their study conditions, and they recommended utilizing particle sizes of 0.125–0.250 mm, with which the F
ion adsorption rate was markedly increased. Low-cost materials based on aluminum hydroxide were proposed by other researchers (Chen et al. 2016) for F
removal, such as carbonized compost coated with aluminum hydroxide. Study of the adsorption kinetics revealed that this material had a good maximum F
ion adsorption capacity of 36.5 mg g
1 adsorbent,

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Table 7.6 Alumina-based adsorbents used for F
ion removal

pH 6.15

Contact time (h) 24

Adsorption capacity (g L
1) 15.43

Adsorbent Nano alumina

[F
]0 (mg L
1) 100

Adsorbent dose (g L
1) 1.0

Alumina hydroxide

–

–

0.5

4

63.50

Fe3O4@Al (OH)3 Mesoporous calciumdoped alumina Al2O3

160

1.0

6.5

1

88.48

2–1000

0.1

6.5

12

450.00

10–150

1.0

6.0

48

83.30

Activated alumina prepared with Boehmite Activated alumina Lanthanumdoped activated alumina Ordered mesoporous alumina Hollow Al (OH)3 spheres γ-Al2O3

10.8

–

4

–

–

0.5

1.0

7.0

–

3.02

0.5

1.0

7.0

–

16.90

18

0.6

6.0

12

135.91

Yang et al. (2014)

200

1.0

7.0

–

16.77

60

5.0

–

5

8.25

10

1.6

6.7

2

25.41

Zhang and Jia (2016) Xu et al. (2016) Xu et al. (2017)

Porous starch loaded with metallic ions (Al, Zr, Fe, and La)

References Bhatnagar et al. (2011) Sujana et al. (2009) Zhao et al. (2010) Li et al. (2011)

Li et al. (2011) LeyvaRamos et al. (2008) Shi et al. (2013)

reducing the baseline F
ion concentration from 10 mg L
1 to 1 mg L
1. The removal capacity was pH-dependent but remained constant in the pH range 6.0–8.0. According to these findings, F
ion can be effectively removed from drinking water by using a modified compost of crop residue biomass. Table 7.6 lists other aluminabased materials that have been used for F
ion removal from water.

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7.3.2.2

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Double Hydroxides

LDH or hydrotalcite is a type of clay that has a positive surface charge and can therefore adsorb anions or oxanions such as F
, arsenite, or arsenate, among others (Gong et al. 2013b; Theiss et al. 2014). The point of zero charge of LDHs is within the pH range from 9 to 12, and they therefore have a positive surface charge at the usual pH values of waters and can eliminate F
ions from the medium (Goh et al. 2008). LDHs can have high F
ion adsorption capacity, with reported values of 17–213 mg g
1 (Lv et al. 2006). F
ion removal has also been studied using LDHs with Ca-Al, but their small particle size hinders their industrial application and their regeneration is hampered by their poor stability at alkaline pH values (Lv et al. 2006; Theiss et al. 2014). Sun et al. (2017) synthesized Ca-Al LDHs with different Ca/Al molar ratios for use in F
ion removal from waters, observing that their morphology, layer structure, and particle size distribution were influenced by the synthesis conditions, which had a major effect on the F
ion removal rate. The highest F
ion removal capacity (146.6 mg g
1) was shown by LDHs with a Ca/Al molar ratio of 2 and synthesized at pH 12, which reached equilibrium at the moment of contact. It is therefore possible to change LDH particle morphology, although this also modifies their F
ion removal capacity. A hybrid adsorbent formed by a Li/Al LDH impregnating and D201 commercial polystyrene anion exchanger was proposed by Cai et al. (2016). It demonstrated good chemical and mechanical stability with an improved working pH range of 3.5–12.0. Fluoride ion adsorption was higher on the hybrid adsorbent than on D201 or activated alumina, achieving a maximum adsorption capacity of 62.5 mg g
1 for a baseline concentration of 4.1 mg L
1. In addition, the treatable water volume was 11-fold higher for the hybrid adsorbent than for the D021 exchanger. This hybrid adsorbent is therefore a promising candidate for water defluorination at an industrial scale.

7.3.2.3

Ion-Exchange Resins and Fibers

Ion-exchange resins and fibers are widely used to remove ionic pollutants from water (Bhatnagar et al. 2011), but their effectiveness for F
ion removal is low. Thus, Ku et al. (2002) only removed 5% of F
ions from water with a baseline concentration of 15 mg L
1, although they markedly improved the capacity of the resin by adding aluminum, obtaining an adsorption rate of 4.6 mg g
1 at a pH of 4.0. Hence, a major increase in the F
ion adsorption capacity of ion-exchange resins can be achieved by modifying adsorption sites, which is generally performed by doping the material with chelating agents that can form hydrogen bonds with the F
ion. Meenakshi and Viswanathan (2007) compared F
ion removal capacity between a chelating resin with sulfonic acid as functional group and an anion-exchange resin, under different equilibrium conditions. They observed that 95% of F
ions were removed from a solution with baseline concentration 3 mg L
1 by 1 g of chelating resin versus 65% removed by the anion-exchange resin.

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Demirkalp et al. (2016) proposed using a Lewatit TP 260 chelating resin doped with Al(III) or Al(OH)3 for F
ion removal from water. They observed higher percentage F
ion removal with the doped versus non-doped resins, obtaining a removal rate of 0.40 mg g
1 for the Al(III)-doped resin and 0.55 mg g
1 for the Al (OH)3-doped resin. The percentage F
ion removal was not affected by the presence of Cl
or SO42
ions but was lower in the presence of NO3
ions. Prabhu et al. (2016) used a composite formed by a synthetic teabag-like chitosan resin modified with lanthanum to remove F
ions from water. Study of the adsorption isotherms revealed a maximum adsorption capacity (17.50 mg g
1) at 12 min and neutral pH for a baseline F
ion concentration of 20 mg L
1. Results also suggested that the adsorption process was mainly controlled by electrostatic interactions and the ion-exchange mechanism.

7.3.2.4

Zeolites

Zeolites are also used to remove F
ions from water because of their large inner surface, although it is necessary to modify their active centers by the addition of metal cations or cationic tensioactives (Teutli-Sequeira et al. 2013). Onyango et al. (2004) modified the zeolite surface for this purpose by exchanging Na+ with Al3+ or La3+ ions and concluded from the adsorption isotherm data that F
ion removal was governed by an ion-exchange mechanism when the zeolite was doped with Al3+ and by a physical adsorption process when it was doped with La3+. In general, higher percentage F
ion removal was achieved with the Al3+-doped versus La3+-doped zeolite under the different experimental conditions tested.

7.3.2.5

Carbon Materials

Activated carbon is used as an adsorbent to remove a wide range of pollutants from water, due to its large surface area, microporosity, and surface chemistry (Bhatnagar et al. 2013; Navarro et al. 2006), including F
ions, as detailed below. Daifullah et al. (2007) studied the F
ion adsorption capacity of activated carbon derived from pyrolyzed rice straw and oxidized with HNO3, H2O2, or KMnO4, recording the highest capacity (19.5 mg g
1 F
at water pH) for the material oxidized with KMnO4. Studies of F
adsorption rates on carbonized bone and hydroxyapatite-doped carbonized bone attributed the adsorption mechanism to electrostatic interactions between carbon surface charge and F
ions, and reported that the effective pore volume diffusivity did not depend on the F
mass adsorbed in equilibrium (Leyva-Ramos et al. 2008, 2010; Medellin-Castillo et al. 2014, 2016). Velazquez-Jimenez et al. (2014) found that the F
ion adsorption of a commercial activated carbon was increased by doping with Zr(IV), achieving the highest removal percentage with a Zr(IV)-doped sample that was also immersed in an oxalic acid solution. In another study, the F
ion adsorption capacity of granular activated carbon was increased fivefold by doping with lanthanum oxyhydroxides, achieving

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86% removal at 10 min and 92.6% at 50 min with the doped material, whose adsorption capacity was not altered by the presence of other anions. The authors explored the types of interaction between La(III) and activated carbon functional groups as well as the F
adsorption mechanism, finding that La(III) has a preference for binding to surface carboxyl and phenol groups by H+ displacement, while F
is adsorbed by displacement of the OH
of La(OH)3 (Vences-Alvarez et al. 2015). Activated carbon derived from lemon rinds, at a dose of 10 g L
1, attained a maximum percentage F
ion removal of 83% at pH ¼ 4 (Kumar et al. 2016).

7.3.2.6

New Nanostructured Materials

Interest has grown over recent years in metal-organic frameworks (MOFs), which are crystalline materials formed by binding metal atoms or aggregates via organic ligands. MOFs have cavities with molecular dimensions, generally being classified as microporous (pore size 7, although its performance was not affected by the presence of Cl
or Br
ions. Karmakar et al. (2016) synthesized a MOF with aluminum fumarate that demonstrated excellent properties for F
ion removal from waters, including high thermal stability, large surface area (1156 m2 g
1), and a point of zero charge of 8.1, meaning that its surface charge is positive at natural water pH, facilitating F
adsorption. They reported a maximum adsorption capacity of 600 mg g
1 at 293 K, 550 mg g
1 at 303 K, 504 mg g
1 at 313 K, and 431 mg g
1 at 333 K, finding no interference with adsorption from the presence of other ions, such as Cl
, bicarbonate, nitrate, or phosphate. The above data support this material as a candidate for the treatment of natural waters with high F
ion concentrations. Silica xerogels (SXGs) are currently the most widely studied group of aerogels (Teutli-Sequeira et al. 2013; Andrade-Espinosa et al. 2010) and are of major interest in the field of wastewater treatment (Mohammadi and Moghaddas 2015; Han et al. 2016). SXGs have a large surface area due mainly to the presence of mesopores. They exhibit a greater adsorption capacity in comparison to other frequently used adsorbents, with high mechanical resistance. They proved to be good adsorbent materials when their surface is modified with nanoparticles of iron or other metals that form oxides or hydroxides. Thus, recently, Hernández-Campos et al. (2018) studied the behavior of lanthanum (La)-doped silica xerogels in the removal of fluorides from water. Four xerogels were synthesized, one acting as blank (X-B), two doped with LaCl3 and dried at different temperatures (X-LaCl and X-LaCl-M), and a fourth doped with La2O3 (X-LaO). The results showed that fluorides are only removed when La-doped xerogels are utilized. In addition, X-LaCl yielded the highest adsorption capacity, removing 28.44% of the initial fluoride concentration at a solution pH of 7. Chemical characterization of these materials confirmed that fluoride removal took

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169

Table 7.7 Fluoride mass adsorbed (mmol g
1) in the presence of dissolved organic matter

Material XB XLaCl XLaCl-M XLaO

Adsorption capacity (mmol/g) [DOM] Distilled water 5 mg L
1 2.71 5.18 43.54 46.40 41.84 48.45 3.96 6.13

[DOM] 50 mg L
1 4.77 48.89 4.12 6.88

[DOM] 250 mg L
1 – 56.44 61.21 53.80

place by precipitation of LaF3 on the surface of La-doped xerogels. The presence of dissolved organic matter (DOM) in solution increased the removal capacity of La xerogels (Table 7.7). Analysis of the influence of solution pH revealed that the capacity of all xerogels to adsorb F
was highest at a solution pH of 7. Oladoja et al. (2016) studied the behavior of nano magnesium oxide (NMgO), as a defluoridation sorbent, in both synthetic feed water and groundwater (GW) that were contaminated with chromium (Cr(VI)), one of the naturally occurring oxyanions in GW. Competitive sorption studies was carried out, using each of the ionic species, either as the sorbate of interest or the interfering ionic specie, in the two aqueous matrixes. The efficiencies (%) of interfering Cr(VI) in F
sorption by NMgO and the competitive coefficient values were far lower than the values obtained for the same parameters, when the role of F
was studied as an interfering ionic specie in Cr(VI) sorption by NMgO. Therefore, the competitive effects of the presence of Cr(VI) in the sorption of F
by NMgO were insignificant; thus, it was concluded that Cr(VI) had non-interactive effect in the defluoridation efficiency of NMgO. The magnitudes of the efficiency of competition of the interfering ionic specie were higher in the GW system than in the synthetic feed water system. In contrast, the values of the competition coefficient were lower in the GW system than in the synthetic feed water system. The determination of the effects of the competitive sorption on the quality characteristics of the GW showed that carbonate also exhibited antagonistic effects on the sorption process. Chen et al. (2009) developed a technology for the granulation of Fe–Al–Ce nanoadsorbent (Fe–Al–Ce) in a fluidized bed. The coating reagent, a mixture of Fe–Al–Ce and a polymer latex, was sprayed onto sand in a fluidized bed. The granule morphology, coating layer thickness, granule stability in water and adsorption capacity for fluoride was investigated by analyzing samples for different coating time. The coating amount was from 3% to 36%. With increasing coating amount, granule stability decreased and fluoride adsorption capacity increased. Coated granules with a coating amount of 27.5% had a fluoride adsorption capacity of 2.22 mg g
1 (coated granules) at pH 7 and initial fluoride concentration of 0.001 M. A column test showed that 300 bed volumes can be treated with the effluent under 1.0 mg L
1 at an initial fluoride concentration of 5.5 mg L
1, space velocity of 5 h
1 and pH of 5.8. These authors reported that the coating granulation of the Fe–Al–Ce adsorbent can produce granules that can be used in a packed bed for the removal of fluoride from drinking water.

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A novel adsorbent of sulfate-doped Fe3O4/Al2O3 nanoparticles with magnetic separability was developed for fluoride removal from drinking water by Chai et al. (2013). The nanosized adsorbent was characterized and its performance in fluoride removal was evaluated. Kinetic data revealed that the fluoride adsorption was rapid in the beginning followed by a slower adsorption process, nearly 90% adsorption can be achieved within 20 min and only 10–15% additional removal occurred in the following 8 h. The calculated adsorption capacity of this nanoadsorbent for fluoride by two-site Langmuir model was 70.4 mg g
1 at pH 7.0. Moreover, this nanoadsorbent performed well over a considerable wide pH range of 4–10, and the fluoride removal efficiencies reached up to 90% and 70% throughout the pH range of 4–10 with initial fluoride concentrations of 10 mg L
1 and 50 mg L
1, respectively. The observed sulfate–fluoride displacement and decreased sulfur content on the adsorbent surface revealed that anion-exchange process was an important mechanism for fluoride adsorption by the sulfate-doped Fe3O4/Al2O3 nanoparticles. Moreover, a shift of the pH of zero point charge (pHPZC) of the nanoparticles and surface analysis based on X-ray photoelectron spectroscopy (XPS) suggested the formation of inner-sphere fluoride complex at the aluminum center as another adsorption mechanism. With the exception of PO43
, other co-existing anions (NO3
, Cl
and SO42
) did not evidently inhibit fluoride removal by the nanoparticles. These findings demonstrated the potential utility of this nanoadsorbent that could be developed into a viable technology for fluoride removal from drinking water.

Table 7.8 Comparison of different halide ion removal technologies

Technology Membrane

Electrochemical Adsorption

Treatment method RO NF ED/RED Electrolysis CDI LDH Double sol-gel

Br
removal (%) 90.0–99.8 93.0–97.0 72.0–80.0 79.0–99.0 50.0–86.1 27.0–94.0 9.0–80.0

I
removal (%) 80.0–92.0 55.0–91.0 92.0–97.0 – 69.7–77.0 14.0–96.0 –

Hydrated oxides Activated carbon Activated carbon-Ag Silver-doped carbon aerogels

80 mg g
1 adsorbent –

71% with thermally activated alumina at 500  C 13.0–46.0

F
removal (%) >90.0 >50.0 43.3–63 High High >70 Medium/ low 90 mg g
1 adsorbent Low

72.0–73.0

–

Medium

3.01– 5.78 μmol g
1 adsorbent 2.0–83.0

1.98–5.03 μmol g
1 adsorbent

Medium/ high

–

–

MIEX

7 New Technologies to Remove Halides from Water: An Overview

7.4

171

Comparison of the Different Technologies to Remove Halides from Water

Table 7.8 displays percentage Cl
, Br
, and I
ion removal values as a function of the technology and treatment method applied. Membrane techniques yield higher percentage Br
ion, and I
ion, and NOM removal from natural waters, contributing to a reduction in halogenated DBPs. Electrochemical techniques perform well in halide removal from waters, but their effectiveness is reduced by the presence of other anions and NOM and their operating costs are high. CDI requires further development to enhance its applicability but is worthy of consideration for halide removal by the drinking water treatment industry.

7.5

Conclusions

The increasing demand for hydric resources requires a greater exploitation of alternative sources of drinking water while water quality requirements are becoming increasingly strict, as evidenced in the WHO guidelines. Consequently, there is a need to implement effective technologies for halide removal from waters to avoid DPB formation. There are innumerable individual DBP species that cannot plausibly be controlled and regulated, and the removal of DBP precursors offers the key advantage of minimizing the formation of all brominated and/or iodized DBPs, including known or unknown and regulated or nonregulated products in a simple and effective manner, reducing the formation of all brominated and/or iodized DBPs. Br
and I
removal methods are classified as membrane, electrochemical, and adsorption techniques. Membrane techniques have demonstrated excellent effectiveness to remove halides but are costly and energetically inefficient. Electrochemical techniques (electrolysis, CDI, and MCDI) also have good halide removal capacities; however, unlike membrane techniques, they do not effectively remove NOM, which is an essential step in the prevention of DBP formation. After further technological development, CDI and/or MCDI may prove suitable for application in drinking water treatments. Variable results have been obtained for bromide and/or iodine removal using adsorption methods (LDH, activated carbons doped with silver, carbon aerogels, ion-exchange resins, aluminum coagulation, and flocculation), which are limited by interference with halide adsorption from competitor anions and NOM. Nevertheless, the utilization of adsorption techniques is a promising research area, given their relatively low cost and easy application. Several research papers have shown that nanotechnology based systems can be efficient and effective in removal of halide ions from water. Most of the studies have been conducted in batch experiments, and under controlled conditions, but experiments with actual contaminated waters are lacking. Large-scale commercial production of nanoparticles would considerably bring down the production cost. These

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materials can be a useful solution for the removal of pollutants, in general, and halide ions, in particular; however, more studies are needed in order to fully understand the following issues: (i) the desorption process and nanoadsorbent regeneration, (ii) the evaluation of the adsorption process in large-scale, (iii) the release of metal ions from the nanoparticles into the environment, (iv) the nanoparticles disposal after the water treatment, and (v) the evaluation and analysis of nanoparticles life cycle, their toxicity, impacts, and distribution in the different ecosystems. Acknowledgments The authors are grateful for the financial support of the Ministry of Science and Innovation (CTQ2016-80978-C2-1-R).

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aluminum (III) mixed oxide. J Ind Eng Chem 46(16):5346–5356 Bo L (2008) Electrolytic cell and process for removal of bromide ions and disinfection of source waters using silver cathode and/or dimensionally stable anode (DSA): a process for the reduction of disinfectant/disinfection byproducts in drinking water. US Patent US7384564B2 Bórquez R, Ferrer J (2016) Seawater desalination by combined nanofiltration and ionic exchange. Desalin Water Treat 57(58):28122–28132 Bougeard CM, Goslan EH, Jefferson B, Parsons SA (2010) Comparison of the disinfection by-product formation potential of treated waters exposed to chlorine and monochloramine. Water Res 44(3):729–740 Boyer TH, Singer PC (2006) A pilot-scale evaluation of magnetic ion exchange treatment for removal of natural organic material and inorganic anions. Water Res 40(15):2865–2876 Cai J, Zhang Y, Pan B, Zhang W, Lv L, Zhang Q (2016) Efficient defluoridation of water using reusable nanocrystalline layered double hydroxides impregnated polystyrene anion exchanger. Water Res 102:109–116

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Chapter 8

Nanotechnology Explored for Water Purification A. Laha, D. Biswas, and S. Basak

Contents 8.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.2 Water Filtration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.2.1 Nanofiltration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.2.2 Nanocomposite . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.2.3 Nano-Fibre Membranes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.2.4 Nano-Wire Coating . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.3 Nanotechnology Used in Desalination of Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.3.1 Nanomaterials in Reverse Osmosis (RO) for Desalination . . . . . . . . . . . . . . . . . . . . . . 8.3.2 Engineered CNT-Based Membranes for Effective Desalination . . . . . . . . . . . . . . . . . 8.3.3 Mesoporous Cobalt Oxide Silica Membranes for Desalination Application . . . . 8.3.4 Nanoporous Single Layer Graphene as Desalination Media . . . . . . . . . . . . . . . . . . . . . 8.4 Nanotechnology Used for Portable Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.5 Nanotechnology Used for Effluent and the Wastewater Treatment . . . . . . . . . . . . . . . . . . . . . . 8.5.1 Nano-Absorbent . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.5.2 Bioactive Nanoparticles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.5.3 Bio-membrane for Water Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.5.4 Nanocatalyst . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.5.5 Nanocatalytic Membrane . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.6 Challenges and Constraints of Nanotechnology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.7 Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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A. Laha Reliance Industry Ltd., Navi Mumbai, Maharashtra, India D. Biswas Indian Jute Industries Research Association, Kolkata, West Bengal, India S. Basak (*) Indian Council of Agricultural Research, Central Institute for Research on Cotton Technology (CIRCOT), Mumbai, Maharashtra, India © Springer Nature Switzerland AG 2019 R. Prasad, K. Thirugnanasambandham (eds.), Advanced Research in Nanosciences for Water Technology, Nanotechnology in the Life Sciences, https://doi.org/10.1007/978-3-030-02381-2_8

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Introduction

About two-third of Earth’s surface is covered by water. But only 3% of the total available water on the Earth’s surface can be used as drinking water or in industrial applications. There is no laboratory or industry that makes water for the above use. On the other hand, World’s population is increasing rapidly, population to reach around two to three billion in next 40 years. Also huge industrial revolution is happening simultaneously. Thus, the demand of fresh water is increasing exponentially. Hence there is a real need to develop novel and innovative water purification technology for drinking water as well as fulfil the industry requirement (Patra and Gouda 2013; Qu et al. 2013). Water scientists are continuously trying to develop a novel water purification technology that would be sustainable, robust, energyefficient and cost-effective. They have opted different approaches to resolve the present issue. Most traditional techniques for the water filtration processes have been achieved by using geolite, chlorine-based chemicals, potassium permanganate, alum, etc. Membrane separation technology (nonwoven textile made membranes) and reverse osmosis are also popular in the field of the conventional filtration. However, efficacies of these traditional chemicals are less and larger quantities of the chemicals have been consumed for it (Wegmann et al. 2008; Gehrke et al. 2015; Sharma and Sharma 2012). Halogen-based products (chlorine, bromine, iodine) also have harmful side effects towards the human being and now a days commercially banned in different parts of the world market. Different kind of lightweight nanotechnology-based products like nanocomposite, nanofibre membrane and nano wire coating has great potential to provide filtered water (Kim and Deng 2011; Fathizadeh et al. 2011, 2017; Joshi and Bhattacharyya 2011; Lee et al. 2011; Mejía et al. 2017). Concerning the effluent treatment, traditional techniques like extraction, absorption and chemical oxidation are generally effective but often very expensive. Nanotechnologies can play an important role due to their high surface area, reactivity and absorption power. In addition, this advance technology can also solve the technical challenges by absorbing effluents, bacteria, virus, harmful pathogens, dyes, heavy metals (lead, cadmium, zinc), pesticides, insecticides, etc. from the water (Qian and Hinestroza 2004; Karim et al. 2009; Schoen et al. 2010; Hassan et al. 2009; Seshama et al. 2017). Desalination is another part of the water purification and conventionally performed by membranes (chemical, mechanical and electrical mediated) or thermal processes, i.e. evaporation of water and vapour condensation employing heat and pressure, etc. Conventionally used large-scale water treatment plant for performing the desalination is enormously expensive in terms of the energy and chemical consummated and also it requires skilled manpower. Therefore, nanotechnology is an effective and emerging approach in this regard. Connected to this area, carbon nano tube (CNT)-based membrane, mesoporous cobalt oxide and silica membrane are emerging in the domain of the research field (Nednoor et al. 2007; Lin et al. 2012). This promising technology also has emerged in the agricultural sector and has great potential to be used for plant protection, water purification and for the improvement of soil. However, in most of

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the cases newly developed technologies are not coming into the commercial market (especially in the developing and the underdeveloped countries) due to the size variation, lack of uniformity and for large initial investment.

8.2

Water Filtration

Filtration or purification is a process of separating the unwanted materials dispersed in water, by using porous structures. The cheap sources of porous structures for water filtration are different membranes. These membranes are mostly made by using textile structures. Among the various textile structures, application of nonwoven structure is found to be most suitable and cheapest material for water filtration. There are varieties of nonwoven available with wide range of thickness, pore size and pore distribution. The pore size of nonwoven structures can be as low as nanoscale. Pore size is the main parameter which determines the effectiveness of the filter. Textile structures are most widely used as filters because they have uniform pore size and pore distribution, stability to thermal and chemical conditions, resistance against clogging and very good filtration efficiency. Based on the filtration requirements actual textile structures are selected. Membrane separation technology is rapidly advancing and largely accepted as automated process for water filtration technology. The physical barrier provided by the membranes made of fibrous material depends on their pore size, pore distribution and particle size to be separated out. From last few decades researchers are concentrating on nanotechnology for improving selectivity and filtration efficiency (Patra and Gouda 2013).

8.2.1

Nanofiltration

Use of Nanofiltration technique is recently becoming the most common trend for water filtration. It is a pressure-driven process, wherein molecules and particles less than 0.5–1 nm can pass through Nanofilter. These filters also work by a unique charge-based repulsion mechanism for separation of various ions (Qu et al. 2013). They are predominantly used to reduce the hardness, colour, odour, and heavy metal ions from groundwater. They also have very good potential to convert the seawater into distilled water in a very cost-effective manner (Sharma and Sharma 2012; Gehrke et al. 2015).

8.2.2

Nanocomposite

Recently scientists have introduced Nanocomposite as a new group of filtration materials. It comprises two parts, namely matrix and surface-functionalized

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nanofibres. Different nanofillers like metal oxides (Al2O3, TiO2), antimicrobial nanoparticles (nanosilver, CNTs), zeolites, etc. are added to the matrix and surfacefunctionalized nanofibres. These nanofillers have high surface area and they assist to improve hydrophilicity, photocatalytic effect and antimicrobial activity of the nanocomposite (Wegmann et al. 2008). Researchers are also using ordered mesoporous carbons as nanofillers to develop new type of thin film nanocomposites (Kim and Deng 2011). Selective plasma treatment is carried out to the ordered mesoporous carbons to improve the hydrophilicity up to a certain percentage. This phenomenon helps to increase the permeability of the pure water. These types of thin film semipermeable nanocomposite membranes are widely used in the reverse osmosis process. Fathizadeh et al. (2011) developed thin film nanocomposite membrane of polyamide and nano-NaX zeolite (40–150 nm) coated with interfacial polymerization of trimesoyl chloride and m-phenylenediamine monomers on the surface of porous polyethersulfone. These membranes are also very good at filtration of ground water as well as seawater. Fathizadeh et al. (2017) has recently found very fine 2-dimensional membranes of fine graphene oxide suitable for water purification due to their unique physico-chemico-mechanical properties.

8.2.3

Nano-Fibre Membranes

A fibrous material is defined as nanofibres when the diameter is in nano level (1–100 nm) (Mejía et al. 2017). Nanofibres are characterized by high surface area to volume ratio. Using electrospinning technique a wide range of polymers, namely Polyamide-66, polycaprolactone (PCL), polystyrene (PS) fibres, poly methyl methacrylate (PMMA), polylactic acid (PLA) fibres, Chitosan fibres, etc., are spun into nanofibre webs. This nanofibre webs can be used as nanofiltration very efficiently. These nanofilters are very good to reduce the water hardness and remove natural and synthetic organic material dissolved in water (Qian and Hinestroza 2004; Joshi and Bhattacharyya 2011). Argonide Corporation, Sanford, FL, USA, developed one patented technology, NanoCeram®. This is an electropositive filter medium that is implemented in a filter cartridge. NanoCeram® is characterized by fibre diameter in nano range and very high surface area (300–600 m2/g). They are used for ultrapure water systems for laboratory or commercial applications, as a microbiological sampler, etc. (Karim et al. 2009).

8.2.4

Nano-Wire Coating

Schoen et al. (2010) developed new textile-based multi-scale technology for water filtration. This filtration system may find application in developing countries also as the raw materials used for this technology are very cheap, abundantly available and

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requires very low voltage for the filtration. Unlike slow-operating conventional filters, where trapping of bacteria is through small pores, this new filter technology is a high-speed electrical sterilization of water combo of cotton, carbon nano tube and silver nanowire. Very small electric current is required to kill bacteria and other microorganisms. This is a gravity fed device with accuracy level of more than 98% bacteria killing in few seconds. This excellent performance of the new textile-based multi-scale technology is coupled with available high surface area of the nanowire, and use of low voltage instead of using size inclusion process allows for effective bacterial inactivation. These types of filters can be used to filter of drinking water as well as industrial applications.

8.3

Nanotechnology Used in Desalination of Water

The earth is 70% covered with water even though its habitats are facing stringent shortage of drinking water in recent days. The reason is salinity of seawater which human beings cannot consume to fill the thirst. Therefore, to ensure the steady source of drinking water, desalination of the saline water may be considered as the best option. In theoretical terms desalination refers to the removal of salts and few identified minerals from saline water for obtaining drinkable water by means of both physical and chemical methods (Schoen et al. 2010; Hassan et al. 2009). Evaporation of water from sea surface and subsequent condensation to form rain water droplets is the natural desalination process performed by the nature which is difficult to replicate in practicality to serve the needs of the growing population. Conventionally, desalination is done by using either membranes (chemical, mechanical and electrical mediated) or thermal processes, i.e. evaporation of water and vapour condensation employing heat and pressure. A large-scale water treatment plant for desalination by conventional method is enormously expensive as it is energy and chemical intensive, as well as skilled manpower consuming. Thus, it does not support the economic viability of the process. Desalination of water is now an emerging field of research where alternative advanced technologies are being tried of among which nanotreatment-based desalination process is promising one. Different nanostructured materials like nanoporous single layer graphene, mesoporous cobalt oxide with nanosilica membrane, etc. have been explored by the researchers for the conversion of the seawater to the fresh water.

8.3.1

Nanomaterials in Reverse Osmosis (RO) for Desalination

Desalination is performed through reverse osmosis (RO), i.e. the osmotic pressure difference between the pure water and the saltwater to remove salts from water.

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Membranes that are commonly used in RO are cellulose acetate, aromatic polyamide, polypiperazine-amide, polybenzimidazoline, polyoxadiazole, polyfurane, polyetherpolyfurane, sulfonated polysulfone, polyamide via polyethylenimine, polyvinylamine, polypyrrolidine, polypiperazine-amide, cross-linked fully aromatic polyamide and cross-linked aralkyl polyamide (Lee et al. 2011). Nanofiltration in combination with RO has been used for desalination by the researchers. A nanofilter membrane due to larger membrane pore size (0.05–0.005 μm) requires less pressure (70 and 140 psi) in comparison to RO. Nanofilters are also used to remove cations, organic pollutants, biological contaminants, natural organic matter, minute quantities of U (VI), arsenic and nitrates from seawater.

8.3.2

Engineered CNT-Based Membranes for Effective Desalination

Engineered membranes having ion selectivity, i.e. ‘ion-channel’, have been created by the researchers to efficiently separate salt ions through the membranes with high water flux. CNTs with hollow graphite cores and metal oxide frameworks have been modified to provide a water channel-like function, i.e. water transport occurs in a single-file fashion (Nednoor et al. 2007). A higher salt rejection has been reported and a nearly double water flux obtained by using CNT/polymeric membranes.

8.3.3

Mesoporous Cobalt Oxide Silica Membranes for Desalination Application

The performance of cobalt oxide silica membrane (CoOxsi) prepared by sol-gel technique has been studied by Lin et al. (2012). Desalination of brackish, seawater and brine has been examined by the researchers which revealed that 99% of salt rejection is possible. The (CoOxsi) membrane works comparatively better than other inorganic membranes based on zeolite and silica (Lin et al. 2012; Elma et al. 2015).

8.3.4

Nanoporous Single Layer Graphene as Desalination Media

Single layer Graphene having sub-nanometre pores in its morphological structure has the potential to selective separation of salt ion from saline water (Cohen and Grossman 2012). Sint et al. (2008) studied the possibility of transport of gases and ions through pores in graphene membranes and identified graphene as prospective

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Fig. 8.1 Desalination mechanism with Graphene membrane (Sint et al. 2008; Cohen and Grossman 2012)

material for water filtration (Sint et al. 2008) (Fig. 8.1). Graphene pores chemically functionalized with hydrogen are better at rejecting salt ions, but have lower flow rates than pores functionalized with hydroxyl groups. Based on their simulations, Cohen and Grossman (2012) identified the ideal sizes for hydrogenated pores (23.1 Å2) and hydroxylated pores (16.3 Å2) to maximize water throughput. A graphene membrane with sub-nanometre pores is a promising reverse osmosis membrane. High pressure applied to the salt water (left) drives water molecules (red and white) across the graphene membrane (right), while salt ions (spheres) are blocked. However, the big challenge is scalable manufacturing of large graphene membranes with sub-nanometre pores with a narrow size distribution, while maintaining the structural integrity of the graphene and keeping the costs in low level.

8.4

Nanotechnology Used for Portable Water

Nano technology in the water purification process can provide very much efficient filtration process. As per the report of the researchers, nano-based filtration membranes are more efficient rather than the chlorine, iodine, bromine or other chemicalbased membranes. In this direction, nano scientists in India have developed a new kind of portable water purification system based on the nanoparticles filtration. They have reported that the nanoparticle used, not only remove the bacterial contamination from the water but also the heavy metals are getting absorbed in the nanoparticle. Connected with the concerned technology, researchers have developed a two-stage filtration process that can filter 10 L of water within a short time span of approximately 1 h. Indeed, they have delivered very few amount of nanosilver ions (for meet with international safety standard) in the water, without using any electrical system. As per their report, silver nanoparticles have been trapped in tiny cage-like

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structures, made by clay-based material. First stage of the filtration process is to kill the harmful Gram-positive and Gram-negative bacteria and dangerous microbacteria. Thereafter, the heavy metals (lead, arsenic, etc.) present in the water also have been absorbed by the nano particles present in the filter. Thus, the purified filtrate produced is very much suitable for drinking and cooking purpose. However, constant control release of the nanosilver ions in the water is still a big challenge for the researchers. The production of the nanofibre-based membranes is also one another greatest achievements by the researchers in the field of portable water filtration. High surface area based nanofibres can easily absorb harmful bacteria, microbes and also heavy metals from the water. High temperature resistance, eco-friendliness and the longer life span of the membranes are major advantages behind the usage. As per the report, nanofibres are produced by the electrospinning process and different nanoparticles have been successfully incorporated inside the nanofibre for bacteria killing and for the metal absorption. Very recently, NASA has developed a very much efficient portable nanomesh that can create clear drinking water for the astronauts. Concerning the reported technology, it is the constant reusage of the wastewater by purification process using nano technology by a closed loop system. Most of the modern filtration system needs high level of electricity for fast and efficient filtration process. However, they have used highly electrically conductive carbon nano tubes. Carbon nano tubes are few nanometre in diameter and can be used where small, lightweight and efficient filter material is required. Very recently in 2015, Fast Company has introduced with the revolutionary concept of Nanotech bottle, which can instantly purify the water. As per the report of the researchers, water taken from the shallow river, drain, etc. by the nanotech bottle is instantly suitable for drinking. Report also elucidated the fact that the nanoscale holes present in the filter (mainly carbon based) of the bottle help to clean the water by absorbing dangerous microbes and bacteria.

8.5

Nanotechnology Used for Effluent and the Wastewater Treatment

Water contamination is a major problem in the whole world today. Most of the chemical and textile industries are liberating poisonous, harmful effluent which is very much dangerous for the mankind. Different treatment processes have been followed by the industries and the researchers to remove the harmful contaminants from the water. However, in most of the cases larger quantity of the chlorine, carbon, iodine and bromine-based chemicals is required to get clear odour-free clean water. From the last one decade, researchers are trying to develop nano-based cleaning system of the effluent so that the cleaning will be more effective. In addition, less quantity of the nano-based chemicals is required for the treatment. Indeed,

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nanoparticles have high surface area, have high absorbing, reacting and interacting capabilities and can be used effectively for removing toxic metal ions, harmful microbes and chemical solutes from effluent water (Khalil et al. 2011; Prachi et al. 2015). Various classes of nanomaterials are like metal-containing nanoparticles, carbonaceous nanomaterials, zeolites and dendrimer-based branched chain polymers. Nanotechnology-based different nanomaterials like nanostructured catalytic membranes, biomimetic membrane, bioactive nanoparticles, nanosorbents, nanocatalysts, etc. have been used effectively for the wastewater treatment (Khalil et al. 2009).

8.5.1

Nano-Absorbent

Most of the nano-absorbent has high sorption capacity and can be used effectively for the water treatment. Researchers have developed carbon-based nano-absorbent which can easily absorb heavy nickel ions from the water. It also has high mechanical strength and chemical resistance. Nano clays can absorb hydrocarbon-based dyes and phosphorous-based contaminants from the effluent water. Especially it can be used effectively in the effluent generated from the textile industries. Researchers have also developed re-generable polymeric nano-absorbent and the carbon-based absorbent for removing the organic and the inorganic contaminants from the water (Marcells et al. 2009).

8.5.2

Bioactive Nanoparticles

Silver nanoparticles are very common for removing the bacterial and microbe-based contaminants from the water. Researchers have developed that the nanoparticles can also be developed from the biological sources. As for examples, Silver nanoparticles (AgNPs) synthesized from the extracellular of bacteria Bacillus cereus, having very high antibacterial efficacy against the both Gram-positive and the Gram-negative bacteria. It also has been reported that the magnesium oxide nanoparticles and cellulose acetate (CA) fibres with embedded silver nanoparticles are very effective biocides against Gram-positive bacteria and Gram-negative bacteria (Hyeok et al. 2009).

8.5.3

Bio-membrane for Water Treatment

Bio-membrane-based water treatment is one of the new and the advanced way of water purification, based on its specific design and fabrication. This kind of bio mimetic membrane has been effectively developed by Albuquerque-based Sandia

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National Laboratories and also by the University of New Mexico. As per their report, the invention composed of self-assembly and atomic layer deposition tuned nanopores which generally provide high flux desalination. The membranes remove impurities like salt and others from the water with applied pressure and adequate electrical energy. At low pressure level (5.5 bar), the nonporous bio mimetic design enables high salt rejection and faster water flow through the membrane. The concerned filtration process basically followed the reverse osmosis principle. This particular technique has high filtration efficiency by investing low cost due to the combine usage of nano fabrication with the protein channels of bio-membranes (Hyeok et al. 2009; Jian et al. 2009).

8.5.4

Nanocatalyst

Nanocatalysts, having high surface area and shape-dependent properties, have been used in the water treatment because of its catalytic activity at the surface. Popularly explored catalytic nanoparticles are semiconductor materials, zero-valence metal and bimetallic nanoparticles. These materials have been used for the degradation of the environmental contaminants such as PCBs (polychlorinated biphenyls), azo dyes, halogenated aliphatic, organochlorine pesticides, halogenated herbicides, nitro aromatics, etc. In large scale, hydrogen has been used for making active catalyst which has been performed by redox reactions. However, there is need to reduce the hydrogen economy by making metallic catalysts. Silver (Ag) nanocatalyst, N-doped TiO2 and ZrO2 nanoparticles catalysts have been highly efficient for degradation of the microbial contaminants in water. TiO2-AGS composite is very efficient for Cr (VI) absorption from the wastewater. Wastewater with specific contaminants like traces of halogenated organic compounds can be selectively biodegraded by using advanced nanocatalytic activities (Hyeok et al. 2009; Jian et al. 2009; Hongwei et al. 2012).

8.5.5

Nanocatalytic Membrane

Nanocatalytic membranes also have been used for the treatment of the contaminated water. These membranes have highly uniform catalytic sites, less contact time of catalyst and as a result it is easier for the industrial scale up. Filtration process normally has been performed by nanostructured TiO2 films and membranes under UV and visible-light irradiation. Researchers have reported that the N-doped ZnO nanostructured material could be used for making multifunctional membrane and it is very much efficient to remove water contaminants by enhancing photo degradation activity under visible light irradiation (Khalil et al. 2009; Marcells et al. 2009; Hyeok et al. 2009). The concerned membrane also showed antibacterial activity due to the presence of zinc oxide and helped in delivering the clean water.

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Challenges and Constraints of Nanotechnology

As nanotechnology sector is gradually improving day by day, new and novel nanomaterials are also developed. Despite lot of the potential advantages, nanotechnology applications in the commercial sector are still marginal due to its high cost and lower uniformity and reproducibility. Nanotech products need high initial investment which can only be counterbalanced if the technology is successful in large scale up. In the field of filtration and the agriculture, it has been observed that most of the technologies developed are only remaining as patent or publication (mainly claimed by academic scientists and small-scale industries). However, no new revolutionary nano-based product has come into the field level. The nanomaterial varies in shape and size. In countries like India, lack of information and methods of the characterization of nanomaterials make the existing technology extremely difficult. In addition, there is high amount of energy demand during the synthesis of the nanoparticle with lower recovery and recycling rate. Another major constraint of using nanomaterial is its uniform mode of action. Indeed, due to severe environmental condition it can be agglomerates and its efficacy of filtration may be reduced. Lack of trained workers and the engineers connected with this field are also causing further concern in the way of commercialization. There are two major research needs for the commercial applications of nanotechnology for the water purification and the effluent treatment. First, the performance of various nanotechnologies in treating real natural and wastewater needs to be tested and the concerned studies need to be done under more realistic conditions. Secondly, the long-term efficacy of these nanotechnologies is largely unknown as most lab studies were conducted for relatively short period of time. Therefore the research addressing the long-term performance of water and wastewater by using nanotechnologies is in great demand.

8.7

Conclusion

Nanotechnology is one of the upcoming emerging techniques for the purification of the water in terms of the filtration, cleaning of the effluent water, seawater, purification of the portable water, etc. Different kinds of nano-coated membranes, nanocomposite, nanofibre-based products, nano filler, nano crystals, carbon nano tube-based membranes, etc. show their promising experimental behaviour in the field of the water purification. High surface area, reactivity and absorption power of the nanomaterials make the technology more promising for the future prospect. Major advantages of the nano technology are cost saving, less waste generation, cleaner and green process condition, less green-house gas emission and absorption of toxic gases from atmosphere. However, applicability of the nanotechnology for the commercialization purposes varies widely. Very few of the developed nanotechnologies are available on the commercial market, while others require significant

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research before they can be considered for large-scale applications. Main challenges behind the commercialisations are technical hurdles, cost-effectiveness, and potential environmental and human risk.

References Cohen TD, Grossman JC (2012) Water desalination across nanoporous graphenes. Nano Lett 12:3602–3608 Elma M, Wang DK, Yacou C, Motuzas J, Diniz JC (2015) Interlayer free: nickel doped silica membranes for desalination. Desalination 365:308–315 Fathizadeh M, Aroujalian A, Raisi A (2011) Effect of added NaX nano-zeolite into polyamide as a top thin layer of membrane on water flux and salt rejection in a reverse osmosis process. J Membr Sci 375:88–95 Fathizadeh M, Xu WL, Zhou F, Yoon Y, Yu M (2017) Graphene oxide: a novel 2-dimensional material in membrane separation for water purification. Adv Mater Interfaces 4:1600918. https://doi.org/10.1002/admi.201600918 Gehrke I, Geiser A, Schulz AS (2015) Innovations in nanotechnology for water treatment. Nanotechnol Sci Appl 8:1–17 Hassan SS, Awwad NS, Aboterika AH (2009) Removal of synthetic reactive dyes from the textile waste water by Sorels cement. J Hazard Mater 162:994–999 Hongwei B, Zhaoyang L, Darren DS (2012) Hierarchical ZnO nanostructured membrane for multifunctional environmental applications. Colloid Surf A Physicochem Eng Aspect 410:11–17 Hyeok C, Souhail R, Al-Abed S, Dionysios D, Dionysiou S (2009) Nanostructured titanium oxide film and membrane-based photocatalysis for water treatment. Nanotech Appl Clean Water 34:39–46 Jian X, Leonidas B, Dibakar B (2009) Synthesis of nanostructured bimetallic particles in poly ligand functionalized membranes for remediation applications. Nanotech Appl Clean Water 67:311–335 Joshi M, Bhattacharyya A (2011) Nanotechnology—a new route to high performance functional textiles. Text Prog 43:155–233 Karim MR, Rhodes ER, Brinkman N, Wymer L, Fout GS (2009) New electropositive filter for concentrating enteroviruses and noroviruses from large volumes of water. Appl Environ Microbiol 75:2393–2399 Khalil A, Gondal MA, Dastageer MA (2009) Synthesis of nano-WO3 and its catalytic activity for enhanced antimicrobial process for water purification using laser induced photo-catalysis. Catal Commun 11:214–219 Khalil A, Gondal MA, Dastageer MA (2011) Augmented photocatalytic activity of palladium incorporated ZnO nanoparticles in the disinfection of Escherichia coli microorganism from water. Appl Catal A Gen 402:162–167 Kim ES, Deng B (2011) Fabrication of polyamide thin-film nano-composite (PA-TFN) membrane with hydrophilized ordered mesoporous carbon (H-OMC) for water purifications. J Memb Sci 375:46–54 Lee KP, Arnot TC, Mattia D (2011) A review of the reverse osmosis membrane materials for desalination. J Memb Sci 1:370–379 Lin XC, Ding LP, Smart S, Diniz JC (2012) Cobalt oxide silica membranes for desalination. J Colloid Interface Sci 368:70–76 Marcells A, Omole F, Owino IK, Omowunmi A, Sadik N (2009) Nanostructured materials for improving water quality: potentials and risks. Nanotech Appl Clean Water 45:233–247

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Mejía ML, Zapata J, Cuesta DP, Ortiz IC, Botero LE, Galeano BJ, Escobar NJ, Hoyos LM (2017) Properties of antibacterial nano textile for use in hospital environments. Rev Ing Biomed 11:13–19. https://doi.org/10.24050/19099762.n22.2017.1178 Nednoor P, Gavalas VG, Chopra N, Hinds BJ, Bachas LG (2007) Carbon nanotube made biomimetic membranes: mimicking protein channels regulated by phosphorylation. J Mater Chem 17:1755–1765 Patra JK, Gouda S (2013) Application of nanotechnology in textile engineering: an overview. J Eng Technol Res 5:104–111 Prachi GP, Madathil D, Nair ANB (2015) Nanotechnology in waste water treatment: a review. Int J ChemTech Res 5:2303–2308 Qian L, Hinestroza JP (2004) Application of nanotechnologyy for high performance textiles. J Text Apparel Technol Manage 4:1–7 Qu X, Alvarez PJ, Li Q (2013) Applications of nanotechnology in water and wastewater treatment. Water Res 47:3931–3946 Schoen AP, Hu L, Kim HS, Heilshorn SC, Cui Y (2010) High speed water sterilization using one-dimensional nanostructures. Nano Lett 10:3628–3632 Seshama M, Khatri H, Suther M, Basak S, Ali SW (2017) Bulk Vs nano ZnO: influence of fire retardant behaviour on sisal fibre yarn. Carbohydr Polym 175:256–262 Sharma V, Sharma A (2012) Nanotechnology: an emerging future trend in wastewater treatment with its innovative products and processes. Int J Enhanced Res Sci Technol Eng 1:121–128 Sint K, Wang B, Kral PJ (2008) Selective ion passage through functionalised graphene nanopores. J Am Chem Soc 130:16448–16449 Wegmann M, Michen B, Graule T (2008) Nanostructured surface modification of microporous ceramics for efficient virus filtration. J Eur Ceram Soc 28:1603–1612

Chapter 9

Nanomaterials in the Development of Biosensor and Application in the Determination of Pollutants in Water Germán A. Messina, Matías Regiart, Sirley V. Pereira, Franco A. Bertolino, Pedro R. Aranda, Julio Raba, and Martín A. Fernández-Baldo

Contents 9.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.2 Relevant Characteristics of Nanomaterials . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.3 Nanomaterials Used in Biosensors Methodologies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.3.1 Carbon Nanomaterials . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.3.2 Metal Nanoparticles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.3.3 Metal Oxide Nanoparticles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.3.4 Magnetic Nanoparticles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.3.5 Up-Conversion Nanoparticles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.3.6 Quantum Dots . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.3.7 Silica Nanomaterials . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.3.8 Polymer and Biomaterials . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.4 Recent Developments Applied to Pollutants Determination in Water Samples . . . . . . . . . 9.5 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

9.1

195 198 199 199 202 203 204 204 205 205 206 207 210 210

Introduction

Recently, nanomaterials have aroused much interest due to the increased need for control and monitoring of different analytes present in samples of environmental relevance (Farré et al. 2011; Pumera 2011; Arain et al. 2018; Liu et al. 2018). A nanomaterial comprises nanoparticles (NPs) that are less than 100 nm at least in one G. A. Messina · M. Regiart · S. V. Pereira · F. A. Bertolino · P. R. Aranda · J. Raba M. A. Fernández-Baldo (*) Instituto de Química de San Luis (INQUISAL) – Consejo Nacional de Investigaciones Científicas y Técnicas (CONICET), Universidad Nacional de San Luis (UNSL), San Luis, Argentina e-mail: [email protected] © Springer Nature Switzerland AG 2019 R. Prasad, K. Thirugnanasambandham (eds.), Advanced Research in Nanosciences for Water Technology, Nanotechnology in the Life Sciences, https://doi.org/10.1007/978-3-030-02381-2_9

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Fig. 9.1 Scheme of different methods that can be used for the nanomaterials synthesis

dimension (Gajanan and Tijare 2018). The controlled synthesis and tuning properties of nanomaterials require knowledge of different disciplines such as physics, chemistry, electronics, computer science, biology, medicine, engineering, agriculture, and others that may lead to the emergence of novel and multifunctional nanotechnologies (Gajanan and Tijare 2018). Various methods can be used to synthesize nanomaterials in different forms such as colloidal NPs, metallic NPs, nanoclusters, nanopowders, nanotubes, nanorods, nanowires, and thin films, among others (Mackenzie and Bescher 2007). The conventional methods with some modification can be utilized to obtain nanomaterials (Hasan et al. 2018). Figure 9.1 shows a flow chart of different methods that can be utilized for the synthesis of nanomaterials. The physical, chemical, biological, and hybrid methods have been developed for the nanomaterials synthesis (Prasad et al. 2016; Choi et al. 2007; Hasan et al. 2018). The selection of a synthesis method depends on the material of interest or the type of nanomaterial, their sizes, and the desired quantity (Zhang and Wei 2016; Hasan et al. 2018). Besides, bottom-up and top-down are the main approaches for synthesis of nanomaterials (Hasan et al. 2018; Liu et al. 2018). Bottom-up is an approach in which the miniaturization of material elements (atomic level) followed by self-assembly results in the creation of nanostructures (Hasan et al. 2018). Moreover, during the self-assembly process, the basic unit of a larger structure is composed of nanomaterials (Lan et al. 2017; Hasan et al. 2018). This approach yields lesser defects and a more homogeneous chemical composition (Arlett et al. 2011; Hasan et al. 2018). On the other hand, in the top-down approach, the large (macroscopic) structure can be externally controlled during processing of the desired nanomaterials (Luo et al. 2009; Hasan et al.

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Fig. 9.2 Schematic representation of the biosensor. This device compromise three main components: a sensitive biological recognition element, a transducer, and a signal processor

2018). A major drawback in this approach is the presence of imperfections in the surface structure. Surface defects in this approach can have an impact on physical and surface properties of NPs due to the high aspect ratio (Hasan et al. 2018; Maduraiveeran et al. 2018). In this context, the exciting properties of nanomaterials have attracted the world scientific community toward their application in various sectors such as health, food, security, transport, and information technology (Luo et al. 2009; Hasan et al. 2018; Prasad et al. 2014, 2017; Aziz et al. 2015, 2016). The intelligent use of nanomaterials is predicted to enhance the performance of miniaturized biosensor-based devices with high sensitivities and detection limits (Choi et al. 2007; Luo et al. 2009; Li et al. 2015a, b, c, d; Kurbanoglu et al. 2017; Maduraiveeran et al. 2018; Faraz et al. 2018). A biosensor is a miniaturized analytical device that integrates a biological element on a solid-state surface which enables a reversible biospecific interaction with the analyte, and a signal transducer (Turner et al. 1987) (Fig. 9.2). The biological element is a layer composed of molecules qualified for biorecognition, such as enzymes, receptors, peptides, antigens, antibodies, single-stranded DNA, even living cells are applicable (Tansil and Gao 2006; Arlett et al. 2011). If antibodies or antibody fragments are applied as the biological element, the device is called immunosensor (Bravo et al. 2017; Piguillem et al. 2018). Many devices are connected with a flowthrough cell, enabling a flow injection analysis (FIA) mode of operation (Fernández Baldo et al. 2009). Biosensors combine high analytical specificity with the processing power of modern electronics components to achieve highly sensitive detection systems (Fernández Baldo et al. 2009). There are two different types of biosensors: biocatalytic and bioaffinity-based biosensors (Regiart et al. 2017). The biocatalytic biosensor uses mainly enzymes as the biological compound, catalyzing a biochemical signaling reaction (Regiart et al. 2017). The bioaffinity-based biosensor, designed to monitor the binding event itself, uses specific binding proteins, lectins, receptors, nucleic acids, membranes, whole cells, or antibodies for biomolecular recognition

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(Bravo et al. 2017). Biosensors technology seeks to improve analytical performance by reducing the consumption of reagents, decreasing the analysis time, increasing reliability and sensitivity through automation, and integrating multiple processes in a single device (Regiart et al. 2017). Recently, some biosensors for the pollutants in water samples determination have been fabricated using incorporated nanotechnology (Xu et al. 2013; Wang et al. 2013; Devasenathipathy et al. 2014; Lai et al. 2014; Wei et al. 2014; Chamjangali et al. 2015; Li et al. 2015a, b, c, d; Sun et al. 2015; Bapat et al. 2016; Hayat et al. 2016; Ramnani et al. 2016; Zeng et al. 2016; Mishra et al. 2017; Scala-Benuzzi et al. 2018a). These kinds of novel devices have a high speed of response, accuracy, lower cost, and less operator intervention (Bidmanova et al. 2016; Jarque et al. 2016). Others advantages of these analytical methods are the reduction of the amount of solvents and reagents required in sample pretreatment as well as in the measurement steps (Regiart et al. 2017). These benefits are the consequence of the automation and miniaturization which reduced the adverse environmental impact of analytical methodologies (Fernández Baldo et al. 2009). Besides, the use of some nanomaterials (quantum dots, carbon nanotubes, magnetic and metallic nanoparticles) as a bioaffinity platform for the immobilization of biomolecules or electrode modification had permitted the development of biosensors with enhanced sensitivities and improved response times (Zhang and Wei 2016; Maduraiveeran et al. 2018). This chapter focuses on the application of biosensors with incorporated nanotechnology in the determination of pollutants in water samples.

9.2

Relevant Characteristics of Nanomaterials

Nanomaterials are currently undergoing rapid development due to their potential applications in the field of nanoelectronics, catalysis, magnetic data storage, structural components, biomaterials, and biosensors (Maduraiveeran et al. 2018). In the last years, the use of NPs, nanotubes, and nanowires in biosensor diagnostic devices are being explored (Lan et al. 2017; Liu et al. 2018). With the advancement in properties of nanomaterials, their dimensions at the nanoscale level, new biodevices (smart biosensors) that can detect minute concentration of a desired analyte are emerging (Liu et al. 2018). Nanomaterials are generally used as transducer materials that are an important part for biosensor development (Lan et al. 2017). Also, these nanomaterials are used as such as bioaffinity platform for the immobilization of biomolecules (DNA, enzymes, antigens, or antibodies) or for electrode modification in the development of biosensors (Zhang and Wei 2016; Bravo et al. 2017). The engineered nanomaterials provide higher electrical conductivity, have nanoscale size, can be used to amplify desired signals, and are compatible with biological molecules (Zhang and Wei 2016). For example, carbon materials can be utilized for conjugation of biomolecules (enzyme, antibody, DNA, or cell) (Fernández Baldo et al. 2009). It has been found that the use of nanomaterials may lead to increased biosensor performance including increased sensitivities and low limit-of-detection of

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several orders of magnitudes (Bravo et al. 2017; Regiart et al. 2017). Therefore, nanostructured materials show increased surface-to-volume ratio, chemical activity, mechanical strength, electrocatalytic properties, and enhanced diffusivity (Piguillem et al. 2018). Nanomaterials have been predicted to play an important role toward the high performance of a biosensor (Bravo et al. 2017). To probe biomolecules such as bacteria, virus, or DNA, biocompatibility of nanomaterials is an important factor for designing a biosensor (Maduraiveeran et al. 2018). Furthermore, a variety of samples such as body fluids, environmental samples, food samples, and cells culture can be explored to analyze using biosensors with incorporated nanomaterials (Kurbanoglu et al. 2017).

9.3

Nanomaterials Used in Biosensors Methodologies

Several nanomaterials have been used for the biosensors development in view of their excellent optical, electronic, thermal, and mechanical properties. They are recognized as one of the most interesting materials for the design of next-generation biosensors. With their high surface area to volume ratio, excellent magnetic, great electronic conductivity, and physicochemical properties (Maduraiveeran and Jin 2017), different kinds of nanomaterials (including metal nanomaterials, carbon nanomaterials, magnetic nanoparticles, silica nanoparticles, up-conversion nanoparticles, and quantum dots, between others) have been successfully applied to develop various biosensors for target analytes detection, like several contaminants in water (Zeng et al. 2016). Table 9.1 summarizes and compares the most relevant articles related about nanomaterials used in biosensors fabrication for the pollutants determination in water. In this section, the basic properties of these types of nanomaterials used in biosensors are discussed.

9.3.1

Carbon Nanomaterials

Carbon nanomaterials, including carbon nanotubes, carbon nanofibers, fullerene, graphene quantum dots, graphene, and micro/meso/macro porous carbon, are obtaining a high consideration for their extraordinary properties (Wang and Dai 2015; Yang et al. 2015). Among them, carbon nanotubes, graphene, and porous carbon are the most frequently used carbon nanomaterials in biosensors for contaminants detection in water (Ramnani et al. 2016).

9.3.1.1

Carbon Nanotubes

Carbon nanotubes (CNTs), one-dimensional carbon nanomaterials, are sp2 hybridized carbon atom rolled graphene sheets that were discovered by Iijima in 1991

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Table 9.1 Nanomaterials used in biosensors fabrication for the pollutants determination in water Analytes Herbicide MCPA

Nanomaterials β-CD/MWNTs

Detection methods CV

Dichlorvos Olaquindox Cd2+

MWNTs/ALB MWNTs MWNTs

CV DPV DPV

Sample Natural water Water Water Water

Cd2+

MWNTs/PPCV/Bi

DPV

Water

Cu2+

MWNTs/GCE

NPASV

Thiocyanate

AuNPs/MWCPE

SWV

River water Water

Hydroquinone

rGO–MWNTs

DPV

Phoxim

P3MT/NGR

CV

Isoproturon carbendazim Imidacloprid

GR

SWSV

GO

CV

Napropamide Pb2+ Cd2+

ECL ECL SWASV

Nitrite

Sulfonated GR GQDs Sn/poly( p-ABSA)/ GR Au-rGO/PDDA

Nitrite

GC–GNs–Th

SO32 Hydrazine

GR-CS/AuNPs GCE GR/CCLP-AuNPs

Hydroquinone Hydroquinone

CFG PEDOT/NGE

River water River water Water

Refs. Rahemi et al. (2012) Yan et al. (2013) Xu et al. (2013) Afkhami et al. (2014a, b) Chamjangali et al. (2015) Liang et al. (2014) Afkhami et al. (2014a, b) Hu et al. (2012) Wu et al. (2015) Noyrod et al. (2014)

Lake water Water Water Water

Lei et al. (2014)

Jiao et al. (2015)

AM

Lake water Water

AM

Water

AM

Water

DPV DPV

Sea water Lake water

DPV

Wang et al. (2014a, b) Dong et al. (2014) Wang et al. (2014a, b)

Shervedani et al. (2016) Wang et al. (2013) Devasenathipathy et al. (2014) Li et al. (2015a, b, c, d) Si et al. (2014) (continued)

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Table 9.1 (continued) Analytes Anthracene Thiols Nitrite

Nanomaterials GR-PANI NGR/CoPc Fe3O4–MWNTs PhCOOH

Detection methods SWV AM DPV

Sample Water Water Water

Refs. Tovide et al. (2014) Xu et al. (2015) Pistone et al. (2013)

CV cyclic voltammetry, DPV differential pulse voltammetry, SWV square-wave voltammetry, SWSV square wave stripping voltammetry, SWASV square-wave anodic stripping voltammetry, AM amperometric, NPASV normal pulse anodic stripping voltammetry, β-CD β-cyclodextrin, ALB AChE liposomes bioreactor, NPs nanoparticles, GNs graphene nanosheets, GR graphene, GO graphene oxide, MWCPE multi-walled carbon nanotube/carbon paste electrode, NGR nitrogendoped graphene, PPCV poly(pyrocatechol violet), PDDA poly(diallyldimethylammo-nium chloride), rGO reduced graphene oxide, CFG carboxyl functionalized graphene, GQDs graphene quantum dots, AuNPs Au nanoparticles, PANI polyanilino, CCLP calcium ions cross linked pectin film, PEDOT poly(3,4-ethylenedioxythiophene), CoPc cobalt phthalocyanine, P3MT poly (3-methylthiophene), MWNTs PhCOOH MWNTs chemically functionalized with the benzoic acid group

(Iijima 1991). These can be classified into single wall carbon nanotubes (SWCNTs) and multiwall carbon nanotubes (MWCNTs), according to the number of rolled layers. Since their discovery, NTs have an increased attention due to their unique thermal, electronic, and mechanical properties. In this way, their great mechanical flexibility, unique thermal conductivity, excellent electrochemical stability, and fast electron transfer make them a unique material for the application in biosensors (Tîlmaciu and Morris 2015; Lawal 2016). To improve their solubility and biocompatibility, CNTs can be functionalized with carboxyl or amino groups or form combinations with other materials, such as polymers, ionic liquids, or metal nanoparticles. Chemical groups are able to connect with biomolecules (e.g., antibodies and/or enzymes) or organic molecules (Besteman et al. 2003; Balasubramanian and Burghard 2006). In the last decades, CNTs-based biosensors have been extensively used for the detection of contaminants in water (Xu et al. 2013; Wei et al. 2014).

9.3.1.2

Graphene

Graphene (G), two-dimensional carbon nanomaterials, is a sheet of sp2 bonded carbon atoms that are arranged into a rigid honey comb lattice, exhibiting the highest mechanical strength between know materials, excellent electrical conductivity, large specific surface area, high electron transfer capabilities, exceptional pliability and impermeability, and good biocompatibility (Wang et al. 2016a, b). Since its discovery in 2004, graphene has been successfully applied into various fields, such as energy storage, catalysis, sensor, and electronic devices. In addition, graphene can be easily oxidized into graphene oxide (GO), which contains many hydrophilic groups such as carbonyl, epoxy, hydroxyl, and carboxyl groups (Lei et al. 2014; Vilian et al. 2014). These hydrophilic groups make GO aqueous dispersibility and

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easy to be functionalized with biomolecules (e.g., antibodies, enzymes), which are highly important features in biosensor applications (Si et al. 2014; Wu et al. 2015).

9.3.1.3

Porous Carbon

Porous carbon with a high surface area, accessible surface chemistry, and short pathway for mass and electron transfer has attracted considerable attention due to the promising applications in biosensors, especially in electrochemical biosensors. Jun et al. (2000) synthesized the first nanostructured carbon type CMK-3. They used an ordered mesoporous silica material (SBA-15) and sucrose as a carbon source. The resulting material was a negative replica of the porous structure of SBA-15 formed by interconnected carbon nanorods. This material became attractive due to its interesting textural, structural, and morphological properties. Compared to the corresponding template, the nanostructured carbons exhibit a hydrophobic nature, excellent mechanical strength and thermal stability, thus becoming a material of great interest for several applications (Niu et al. 2016; Regiart et al. 2016). According to the International Union of Pure and Applied Chemistry (IUPAC) recommendation, porous carbon materials can be grouped into three classifications based upon their pore sizes: microporous NC (54.52 mg g
1) > MMT (12.70 mg g
1) (Wang and Wang 2007). Few complex systems have also been employed by the researchers which involved multiple modifiers in an essence to enhance adsorption potential of the clays. One such modification was impacted to synthesize hydrogels of CNC—the latest outcomes of nanotechnology that resulted in enhancement of properties of clay leading to high adsorption. Increase in number of functional group and swelling ability of these gels, prepared by cross-linking MMT with polyethylene glycol diacrylate formed by using acrylamide and itaconic acid sodium salt, led to 457 mg g
1 adsorption in just 30 min. Three types of interactions supported the adsorption phenomenon: (1) Basic dye bonding with the carboxylate group of polymer structure, (2) hydrophobic interaction of dye aromatic ring and polymer, and (3) clay-polymer interactions. Kinetic and equilibrium studies supported that these hydrogel NC can be

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Fig. 10.6 Different mechanisms operate for the dye adsorption on the clay materials

Fig. 10.7 Proposed overview of the iron oxide nanoparticle doped clay interaction with the methylene blue dye molecules resulting in its degradation in the presence of UV light irradiation by generation of radicle ions (Chen et al. 2016a, b)

alternatively used as unique, quick with high capacity adsorbent materials for the excellent removal of cationic dyes (Kaplan and Kasgoz 2011). Doping metallic impurities into the clay structure also affect the removal capacity of clays. Several metal NPs like iron oxides (Chen et al. 2016a; b), TiO2 (Mishra et al. 2017), Pt/TiO2 (Ding et al. 2008), etc. have been tried to get enhanced efficacy for dye degradation. Figure 10.7 details the proposed overview of iron oxide NP doped clay interaction with methylene blue that resulted in its degradation in the presence of UV light irradiation by generation of radicle ions (Chen et al. 2016a, b). Recently bio-nanoclay composites have received a lot of attention due to their unique properties, expandable adsorption capacity, and biologically degradable nature—making them the center of study in nanotechnology. Chitosan-based nanoclay has been used for removal of various dyes by employing different modifications to affect the adsorption process and very high adsorption capacities were also achieved for some of these bio-clays, i.e., chitosan-g-poly(acrylic acid)/MMT (1859 mg g
1) (Wang et al. 2008a; b). These experiments led the way towards focusing on these bio-nanocomposites of clay for water reclamation.

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10.7.2 Removal of Other Organic Contaminants Besides organic dyes there are hundreds of organic molecules/contents found in the drinking or wastewater and many of them are sever enough to cause fatal damage for living organisms. Several types of phenols, aldehydes/ketones, carbon tetrachloride, herbicides, pesticides, naphthalene, etc. must be treated completely before to get clean water. Table 10.2 gives overview of these organic contaminants treated with modified clay or its NCs by various processes of removal.

10.7.2.1

Phenolic Compounds

These compounds rated as carcinogenic are successfully treated with the help of CNC. Phenol adsorption was impacted on nanoclays modified by using different intercalating agents like tetrabutylammonium chloride (TBAC), N-acetyl-N,N,N-trimethyl ammonium bromide (CTAB), and hexadecyl trimethyl ammonium chloride (HDTMA). The adsorption capacity follows the order TBAC (0.53 mg mg
1) > HDTMA (0.52 mg mg
1) > CTAB (0.49 mg mg
1) (Sonawane et al. 2008). Efficient removal of 2,4,6-trinitrophenol (anionic) and trichlorophenol (nonionic) from water was achieved by employing batch experiments and NC of MMT nanoclay to get montmorillonite-poly-4-vinylpyridine-co-styrene (Mt-PVPcoS) and montmorillonitepolydially-diamethylammonium (Mt–PDADMAC) nanocomposites. Phenols interact with NC based on van der Waals and electrostatic interactions resulting in difference in adsorption capacity. Mt-PVPcoS showed almost twice high removal capacity for trinitrophenol (45.8 mg g
1) as compared to trichlorophenol (20.8 mg g
1). Highly positive charged CNC, i.e., Mt-PDADMAC, did not show satisfactory results for these two contaminants (Ganigar et al. 2010). Photocatalytic degradation of phenol into adipic acid and 2,4,6-triphhenoxy phenol was carried out under irradiation in batch as well as continuous stirred tank reactor using ZnO-bentonite NC. This 20–30 nm NC resulted in 67% phenol removal in batch mode as compared to 70% in case of continues mode (Meshram et al. 2011). Better degradation potential (79%) against 4-nitrophenol was achieved when sepiolite loaded with TiO2 NPs was used as photocatalyst in the presence of UV light. Degradation was attributed to multiple factors, i.e., nano size, larger BET surface area, increased absorption, mesopores, and generation of OH and O2
which help in oxidation of 4-nitrophenol (Kaplan and Kasgoz 2011).

10.7.2.2

Pharmaceutical Ingredients

Ibuprofen removal was affected from water and wastewater by use of surfacemodified Cloisite 15A NC. Surface modification was affected by three-step approach involving use of 3-aminopropyltriethoxysilane, mono-tosyl β-cyclodextrin, and polyvinylpyrrolidone. This complex CNC resulted in only 1.1 mg g
1 adsorption

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Table 10.2 Organic contaminants other than dyes removal by clay nanocomposites CNC MMT organically modified with polyamidoamine (PAA-MMT) and polyethylene oxide (PEO-MMT)

Organic compound Toluene

Organophilic bentonite (O-Bt) nanoclays with TBAC, CTAB and HDTMA as intercalating agents

Phenol

Montmorillonite-poly-4vinylpyridine-co-styrene (Mt-PVPcoS); montmorillonitepolydiallydiamethylammonium (Mt–PDADMAC) MMT with HDTMA by Na ions

Trichlorophenol; trinitrophenol

Bentonite with inorganicorganic-intercalated by Co2+, Ni2+, and Cu2+ along with HDTMAB MMT-chitosan

Salicylic acid, naproxen, clofibric acid, and carbamazepine Herbicide (clopyralid)

Cloisite 20A integrated with polyurethane foams organoclay nanocomposites

Oil

CTAB modified MMT

Herbicide (bentazone)

MMT modified with cationic polymer hexadimethrine

Pesticide

Petroleum hydrocarbons

Remarks Removal efficiencies were enhanced for organically modified nanocomposites vs. MMT, i.e., for PEO-MMT removal efficiency increased from 38% to 98% and PAA-MMT from 22% to 97% Langmuir adsorption isotherm showed adsorption capacity of O-Bt-TBAC ¼ 0.53 mg mg
1, O-Bt-CTAB ¼ 0.49 mg mg
1 and O-Bt-HDTMA ¼ 0.52 mg mg
1 Results indicated that Mt-PVPcoS adsorption efficiency was better than Mt– PDADMAC. Mt-PVPcoS remove trinitrophenol 99.5% while for trichlorophenol the removal was 40–60% qe ¼ 4–10 g g
1

Maximum adsorption capacity was found for the salicylic acid with removal of 5.5 μmol/g with Cu2+ modified with HDTMAB Adsorption depends upon concentration and arrangement of chitosan in MMT. At acidic pH adsorption increases up to 5–10% Sorption capacity was enhanced up to 16% and oil removal efficiency was attained 56% in oil-water system and adsorption capacity was 21.5 g g
1 qe ¼ 500 mg g
1 at pH 3

Modified nanocomposites applied on different pesticides and the adsorptions were effective removal of pesticides and removal was occurred as 70–95% of pesticide removal from water

References Ratanarat et al. (2003)

Sonawane et al. (2008)

Ganigar et al. (2010)

Sharafi et al. (2010) RiveraJimenez et al. (2011) Celis et al. (2012)

Nikkhah et al. (2015)

ShirzadSiboni et al. (2015) Gámiz et al. (2015)

(continued)

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Table 10.2 (continued) CNC MMT modified with TMOAB

Organic compound Insecticide (Diazinon)

Hectorite-TiO2 loaded on MMT-TiO2

Herbicide (Dimethachlor)

MMT modified with polymeric Al-Fe (MMT-p-AlFe)

Organic arsenic as dimethylearsinate (DMA)

TiO2 supported sepiolite

Lignin and phenol

ZnO-bentonite

Phenol

Remarks Maximum removal efficiency ¼ 99.95% with TMOAB/ MMT 0.5 g/L in 60 min and maximum adsorption capacity ¼ 1428.5 mg g
1 Photocatalytic degradation of results showed that size of NC matters most hence simple TiO2 has highest degradation ability than TiO2-MMT nanocomposite. Cation-exchange capacity comparison was analyzed by simple MMT and MMT-p-AlFe resulting in 67 meq 100 g
1 for simple MMT and 83 meq 100 g
1 of MMT-p-AlFe. 18.19 mg g
1 adsorption capacity indicated from -bath experiments for NC 80–100% photocatalytic degradation at 11 pH after 24 h in the presence of UV and H2O2 70% Photocatalytic degradation was carried via UV irradiation in special tank at flow of 10 mL min
1 of phenol at pH 12

References Hassani et al. (2015a, b)

Belessi et al. (2007)

Ramesh et al. (2007)

Uğurlu and Karaoğlu (2011) Meshram et al. (2011)

capacity in 120 min (Rafati et al. 2018). Another pharmaceutical excipient– diclofenac, which is frequently available in the treated effluents as well as surface waters, was effectively removed by using poly-4-vinylpiridine-co-styrene modified Na-MMT. High removal potential of the adsorbent was attributed to multiple types of interactions, i.e., electrostatic, hydrophobic, pi-pi, and van der Waals (Kohay et al. 2015).

10.7.2.3

Organic Solvents

Adsorption characteristics of clays for developing affinity towards organic solvents like methyl tertiary-butyl ether, a potential contaminant occurring in surface as well as ground water, was done by multistep sequential synthesis of the MMT NC. The synthetic process adopted consisted of wet ball milling followed by ultrasonication, sedimentation, and surfactant ion exchange. Two surfactants, HDTMA and tetramethylammonium bromide (TMA) used in the study impacted adsorption process which obeys the order clay-HDTMA > clay-TMA. The reason was

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attributed to larger chain length which enhances hydrophobicity of adsorbent as well as basal spacing which enable accommodation of higher quantity of MTBE (Abbas et al. 2017).

10.7.2.4

Pesticides and Related Compounds

Organic compounds like herbicides, pesticides, and insecticides are necessary for environmental hazardous contaminants degradation but their excess amount found its way to drinking water. The organically modified CNC, especially bio-nanoclay composites, are most eco-friendly nanomaterials in removal of these types of contaminants. These bio-nanocomposites’ adsorption potential is affected by change in solution pH. Removal percentage of clopyralid—a herbicide by MMT-Chitosan NC enhanced by 5–10% by lowering the pH to acidic side (Celis et al. 2012). Very high adsorption capacity, i.e., 1428.5 mg g
1, was reported for diazinon based on adsorption isotherms when MMT surfactantly modified with trimethyl-octylammonium bromide (TMOAB) was utilized. The experimental work indicated that maximum removal efficiency was achieved in just 60 min by use of 0.5 g L
1 adsorbent (Hassani et al. 2015a; b). TiO2 NP doped MMT and hectorite prepared by sol-gel approach catalytically degraded dimethachlor under UV radiation. Degradation depends on TiO2 concentration doped that in turn effects the particle size of NP in CNC, pore volume, and surface area. Adsorption follows the order hectorite—15% TiO2 > hectorite—30%TiO2 > hectorite—55%TiO2 > MMT—15% TiO2 > MMT—30%TiO2 > MMT—55%TiO2 > TiO2 depicting that TiO2-NC of hectorite exhibited higher activity as compared to TiO2-MMT; cause of this behavior was attributed to two reasons: (1) Poor accessibility of pesticides into MMT pillared structure and (2) Effective charge transfer in hectorite NC. Overall the entire NCs prepared showed very low percentage adsorption, i.e., p, p0 -DDT > pendimethalin > p, p0 -DDE > metribuzin > α-endosulfan > β-endosulfan > endosulfan sulfate. Factors like smaller particle size, lager surface area, and organophilic nature facilitates such higher removal rate (Shabeer et al. 2014).

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Humic Acid

Presence of humic acid—a natural organic contaminant—affects the quality of water directly by altering its taste, odor, and color and indirectly by adding complications to water treatment processes and resulting in formation of by-products by generation of dis-infection by-products during chlorination process (Jiang and Cooper 2003). Affinity of clays towards humic acid can be increased by undergoing modifications in their structure (Chang and Juang 2004; Jiang and Cooper 2003). Nano-organoclay MMT can perform much better than the traditional alum-based adsorbent in two ways. Firstly, this new adsorbent is operative at neutral pH, i.e., 7.1 as against pH 8 observed for alum. Secondly, by removing 90% of the humic acid this is two times greater than the alum (40% removal).

10.8

Treatment of Inorganic Contaminants

Inorganic contents are the most hazardous contaminants available in their dissolved or undissolved state when found in water. Heavy metals are extremely dangerous for nearly all biological organisms and much of the research work have been published on their fatal and hazardous consequences in the living systems. This chapter hence greatly emphasizes on the implication of nanotechnology with respect to CNC for the treatment of wastewater inorganic contaminants.

10.8.1 Heavy Metals Several heavy metals like Cu (II), Cd (II), Pb (II), Ni (II), Co (II), Cr (VI), Hg (II), etc. as detailed in Table 10.3 are removed from water using competitive and non-competitive modes to entail combined or selective adsorption potential of these nano-organoclays and their composites. Cr(VI) have been subjected for removal from drinking and wastewater due to its highly carcinogenic nature. Various combinations of CNC with metal oxide NP, polymer, chitosan, etc. (Ballav et al. 2014; Chen et al. 2014; Kumar et al. 2011; Moussout et al. 2018; Pandey and Mishra 2011; Setshedi et al. 2013; Yao et al. 2012; Yuan and Wu 2007; Yuan et al. 2009) have been used for Cr(VI) removal from water (Table 10.3). In-situ prepared organically modified MMT-polypyrrole composite was found Cr (VI) selective as demonstrated in binary adsorption systems in the presence of coexisting cations (Ni (II), Co (II), Cu (II)) and anions (NO3
1 and Cl
1). The single metal batch experiments conducted by varying pH, initial Cr (VI) concentration, CNC dosage, and temperature revealed that Cr(VI) adsorption process is temperature dependent demonstrating an increase in Langmuir maximum adsorption capacity from 112.3 mg g
1 at 292 K to 209.6 mg g
1 at 318 K. The

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Table 10.3 Recent studies of the removal inorganic contaminants using clay nanocomposites

CNC MMT, beidellite, saponite, and fluor-tetrasilicic expandable mica pillared with cationic, Keggin ion-like, Al clusters (Alclst) Allophane nanoclay

Heavy metal/ inorganic waste Phosphate

Phosphorus

Kaolinite modified with polyvinyl alcohol

Pb(II); Cd (II)

MMT-supported magnetite NPs

Cr(VI)

Magnetite NPs with diatomitesupported (MagDt-H) clay composites Bentonite with acid activated (AA-B) and manganese-oxide coated (MC-B)

Cr(VI)

Organo-bentonite intercalation with cetyl-trimethylammonium (CTA) and polyacrylic acid (PAA) nanocomposites

Pb(II)

MMT with Ti-pillared

As(III) and As(V) Phosphate

Bentonite pillared with hydroxy-aluminum (Al-bent), hydroxy-iron (Fe-bent), and mixed hydroxy-iron–aluminum (Fe–Al-bent)

MMT-TiO2

Pb(II)

Hg(II)

Remarks Maximum sorption occurs between pH range of 3.0–4.2 and was maximum. For AlclstMont > Alclst-Beid > AlclstSap > Alclst-TSM Removal is 70% reducing concentration of P from 14.2 to 4.2 mg/L The adsorption capacity for Pb (II) was 56.18 mg/g and Cd (II) was 41.67 mg/g Based upon Langmuir and Freundlich adsorption isotherms qe ¼ 15.3 mg g
1 than unsupported 10.6 mg g
1 By comparative study of Langmuir adsorption isotherms qe ¼ 69.2 mg g
1 Langmuir adsorption theorem showed qe for AA-B is 8.92 mg g
1 and MC-B is 58.88 mg g
1 qe calculated on basis of Langmuir isotherm observed to be 52.3 mg g
1 for simple bentonite while 93.0 mg g
1 for PAA-bentonite NC with 99.6% efficient removal of lead from water qe As(III) ¼ 1.5–28 mg L
1; qe As(V) ¼ 1.0–25 mg L
1 Based on Langmuir isotherm the adsorption of phosphate onto Al-bent was 12.7 mg g
1, Fe-bent was 11.2 mg g
1 and Fe–Al-bent was 10.5 mg g
1. The phosphate adsorption was pH independent while by increasing temperature the adsorption capacity was also increased qe of TiO2-montmorillonite at 25  C, 35  C, and 45  C came out as 123.8, 118.3, and 116.5 mg g
1 mg g
1, respectively

References Kasama et al. (2004)

Yuan and Wu (2007) Unuabonah et al. (2008) Yuan et al. (2009)

Yuan et al. (2010) Eren et al. (2009)

Rafiei et al. (2016)

Na et al. (2010) Yan et al. (2010)

Dou et al. (2011)

(continued)

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Table 10.3 (continued)

CNC Cloisite (10A) with chitosan bio-nanocomposite

Heavy metal/ inorganic waste Cr(VI)

Polypyrrole-coated palygorskite

Cr(VI)

Montmorillonite with poly[N(4-vinylbenzyl)-N-methyl-Dglucamine nanocomposite Exfoliated polypyrroleorganically modified MMT

As(V)

Cr(VI)

Polypyrrole-coated Halloysite nanotube

Cr(VI)

Polypyrrole-Sepiolite nanofibers Pyrrole-MMT

Cr(VI)

Bentonite with polyvinyl alcohol

Hg(II)

Zeolite modified with CTAB

Sulfate

Magnetic/raw diatomite and illite clay nanocomposites loaded with Fe3O4 nanoparticles to form MDC, MIC, RDC, and RIC nanocomposites

Phosphate

Illite-Smectite nanoclay

Pb(II)

Cr(VI)

Remarks Based on Langmuir isotherm adsorption capacity was calculated as 357.14 mg g
1 Adsorption increased from 87.9 to 99.3% by increase in adsorption dosage from 0.2 to 0.6 g qe ¼ 55 mg g
1

References Pandey and Mishra (2011) Yao et al. (2012)

The Langmuir maximum adsorption capacity of Cr(VI) was found to be 112.3, 119.34, 176.2, and 209.6 mg g
1 at 292, 298, 308, and 318 K, respectively Maximum Langmuir adsorption capacity was 149.25 mg g
1 at 2.0 pH qe ¼ 302 mg g
1

Setshedi et al. (2013)

Adsorption is highly H dependent and Langmuir isotherm indicated qe was 166.7 mg g
1 qe Hg(II) for 0, 10, 30, 50% bentonite was 460.18, 455.12, 392.19, and 360.73 mg g
1, respectively Maximum adsorption capacity by Freundlich isotherm was 38.02 mg g
1 at 40  C MDC, MIC, RDC and RIC nanocomposites were studied at varying adsorption dosage and phosphate concentration. The adsorption capacity of MDC and MIC were greater than that of RDC and RIC. The removal efficiency of phosphate onto MDC was 98.67% with adsorption capacity of 11.89 mg g
1 and for MIC was 98.89% with adsorption capacity of 5.48 mg g
1, respectively qe ¼ 256.41 μg g
1

Urbano et al. (2012)

Ballav et al. (2014) Chen et al. (2014) Chen et al. (2015) Wang et al. (2014)

Chen and Liu (2014) Chen and Liu (2014)

Yin et al. (2018)

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Fig. 10.8 Removal mechanism of (a) Pb by PANI/clay (Piri et al. 2016) and (b) As (V) by organoclay having intercalated and exfoliated structure (Shokri et al. 2016)

selectivity was attributed to replacement of chloride ion from CNC by HCrO4
1 ion resulting in reduction of Cr (VI) to Cr (III) (Setshedi et al. 2013). A different ion-exchange mechanism, as shown in Fig. 10.8a, was reported for selective adsorption of Pb (II) onto polyaniline/CNC sheet having thickness ~ 40–50 nm. The CNC was prepared by in-situ synthesis using aniline monomer, ammonium persulfate, and clay. The maximum adsorption was achieved at pH 5–6 as at this pH release of imine groups protons is facilitated that in turn results in availability of more binding sites for Pb(II) ions. The system doesn’t work well for Cd (II) sorption as it resulted in only 0.12 mg g
1 adsorption capacity as against Pb (II), i.e., 70.42 mg g
1 inferring to the selectivity of prepared system for later (Piri et al. 2016). Type of mechanism responsible for adsorption depends on the CNC structure. Study carried out on As (V) removal by using two different types of organoclay embedded polysulfone membrane structures, i.e., intercalated and exfoliated (Fig. 10.8b) revealed that adsorption is governed by two major phenomena which are accessibility of metal ion into the clay structure and hence available active sites of the organoclay. Exfoliated structure owing to its greater accessibility revealed higher metal removal as compared to other structure (Shokri et al. 2016). In contrast to this study better As (V) removal was achieved by using intercalated CNC resin of MMT, i.e., Poly[N-(4-vinyl benzyl)-N-methyl-D-glucamine]-MMT, the protonated amine of which electrostatically interacts with metallic ion. Interfering ions like sulfate and phosphate being similar to arsenate with respect to charge and structure didn’t have considerable impact on As (V) adsorption on to this resin (Urbano et al. 2012). Metal adsorption in these nanoclays composites was also affected by the type and amount of clay and functional groups used for modification. A detailed study to show this impact was done on eight hydrogel/clay NCs prepared by using two combinations of monomers (acrylamide and 2-methylpropane sulfonic acid), i.e., 25/75 and 75/25 (weight percentage) with four ratios of clays 0, 10, 30, and 50 (weight percentage). Studies pointed out that metal uptake varies with change in monomers ratio, i.e., for ratio 25/75 the metal order follows Cd (II) > Cu (II) > Pb (II) while for 75/25 the order changes to Cu (II) > Cd (II) > Pb (II) in non-competitive mode while for competitive the selectivity order is Cu (II) > Pb

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(II) > Cd (II) for both the ratios. Incorporation of clay into CNC initially facilitates uptake of metal ions but as the clay percentage is increased metal adsorption decreased due to low metal affinity of the clay. Further, reduction of total basic group and swelling ability with clay >10 wt. % was also attributed to decrease in metal removal capacity (Kaşgöz et al. 2008). Like organic contaminants removal the bio-nanoclay composites are extremely helpful in removal of the metallic impurities as well. Chitosan combination with MMT and Cloisite was used effectively to remove Se (VI) and Cr (VI) from water, respectively (Bleiman and Mishael 2010; Pandey and Mishra 2011). Later bio-nanoclay when subjected to remove Cr (VI) resulted in very high maximum adsorption of 357.14 mg g
1. Amine groups available at CNC surface at pH < 8 gets protonated, thereby electrostatically adsorbing chromium anions (Pandey and Mishra 2011). A similar CNC of chitosan-polyvinyl alcohol/bentonite having exfoliated structure prepared by varying clay ratio was found to be suitable for selective removal of Hg (II) ions, but addition of clay resulted in reduction in removal capacity (Wang et al. 2014).

10.8.2 Anionic Contaminants Several inorganic fertilizers of phosphates, nitrates, and sulfates are dynamically used to enhance the fertility of land but their unused amount in land causes waste in water. Their relegation from clean water is as necessary as heavy metal removal from water; however, not much work is noticed on the utilization of these modified clays and their NC for treatment of water contaminated from these impurities. Phosphate removal was subjected by use of CNC and pillared clays by undergoing different kind of structural modifications. Various clays like MMT, beidellite, and saponite pillared were modified with cationic Al clusters to use in phosphate removal in the presence of NaNO3. Lower pH facilitates adsorption process and as shown by transmission electron microscopy (TEM) analysis majority of phosphate was found in between Al cluster layers which probably deals with transformation of Al-OH bonds to Al–H2PO4
(Kasama et al. 2004). Allophane nanoclay was also found helpful in removal of 70% phosphate from P-rich effluent (Yuan and Wu 2007). pH-independent adsorption was supported by use of bentonite clay pillaring with three different materials, i.e., hydroxy-aluminum (Al-Bent), hydroxy-iron (Fe-Bent), and the mixer of these two, i.e., hydroxy-iron–aluminum (Fe–Al-Bent). The hydroxy-aluminum (Al-Bent) pillared bentonite showed highest adsorption capacity from all three clays. The prime factor observed for enhancement in adsorption capacity is increase in temperature (Yan et al. 2010). Doping of Fe2O3 NPs on to raw diatomite and illite clays was tried but results obtained were not promising; adsorption capacity 11.89 and 5.48 mg g
1, respectively (Chen et al. 2016a; b). Phosphate ion coexists as several species depending upon pH, i.e., H2PO41
and H3PO4 occur at acidic pH while H2PO41
, HPO42
, and NaHPO41
exist at neutral pH (Kasama et al. 2004) which give rise to different adsorption phenomenon (Li et al. 2016).

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Nitrate adsorption was significantly accelerated by employing the clays modification with different surfactants. The outcomes of study conducted on organoclays composites based on halloysite, kaolinite, and bentonite modified with two different ratios of HDTMA (i.e., 2 and 4) showed increase in nitrate adsorption as compared to the unmodified ones. Incorporation of surfactants changes the surface properties of clay hybrids, thereby facilitating electrostatic interaction between surfactant cationic and anionic contaminants. Hence, suggesting that nitrate adsorption is not solely governed by the surface area enhancement of organoclays. Among these prepared nanoclays bentonite clay modified with 2 and 4 ratios of surfactant gave better results, i.e., 12.83 and 14.76 mg g
1 of organoclays, respectively (Xi et al. 2010).

10.9

Treatment of Biological Contaminants in Water

Microbes (unicellular/multicellular) like bacteria, fungi, algae, etc. are also found in the drinking water which are responsible for some serious disease in living organisms. Effective role is played by CNC in water decontamination—the cytotoxicity of which towards microorganisms depend on size, shape, nature, or type of nanoclay (i.e., type of modifier used, hydrophilic/hydrophobic character) as well as on the bacterial strain (Kryuchkova et al. 2016). These nanomaterials does not solely adsorb, like in case of organic and inorganic contaminants, on bacteria cell to remove it but there are several other phenomenon responsible for the antimicrobial activity as will be discussed later in the chapter. Antimicrobial studies carried out on both Gram-positive and -negative bacteria showed that unmodified Cloisite Na+ showed no activity while Cloisite 30B (modified with methyl tallow bis-2-hydroxyethyl quaternary ammonium) have considerable bacteriostatic activity against Gram-positive bacteria and both bacteriostatic and bactericidal activities against Gram-negative bacteria. On the other hand, Cloisite 20A (modified with dimethyl dehydrogenated tallow quaternary ammonium) depicted bactericidal effect only against Gram-positive bacteria. Higher activity of 30A was attributed to its less hydrophobic character as compared to 20A which promotes interaction between the NH4+ cation present in quaternary ammonium ions and bacterial cell surface having anionic charge. This interaction leads to the change in cell membrane permeability which consequently results to cell death. Considering the fact that Gram
ve bacteria have outer membrane that is not present in Gram +ve strains makes the former more resistant to attack (Hong and Rhim 2008). Incorporation of the 30B into Whey protein isolate or agar NC films exhibited antimicrobial activity only against Gram +ve strain (L. monocytogenes) while no activity is recorded towards Gram
ve strain (E. coli). On the other hand, 20A nano-films prepared using same base materials reduced its activity to almost negligible level against both type of strains (Rhim et al. 2011; Sothornvit et al. 2009). In an effort to enhance the antimicrobial activity of clay composites an interesting study was carried out involving Gelidium corneum/clay nanocomposite films. These films were prepared by using different clay materials (Cloisite 30B and Cloisite Na+)

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in varying wt. % followed by incorporation of antimicrobial agents, i.e., grape fruit extract and thymol into the clay films at 04 concentrations. The prepared materials showed very interesting findings with respect of microbial activity, i.e., (1) antimicrobial activity increases with increase in concentration of antimicrobial agents and (2) the CNC films were active against both Gram +ve (L. monocytogenes) and Gram
ve (E. coli) strains (Lim et al. 2010). Incorporation of metal or metal oxide NPs into clay structure have resulted in significant enhancement in antimicrobial activity of these NCs which is ascribed to their dual action, i.e., direct attachment of nanocomposite with bacterial cell wall and indirect role played by metallic nanoparticles. Figure 10.9a, b describes the absorption of these metallic nanohybrids into the bacterial periplasmic compartment as a result of release of metal ions from composite materials on interaction with aqueous media (Fig. 10.9c, d) and bind with functional groups of proteins and enzymes, present on the bacterial cell wall resulting in cell inactivation and lysis (Bagchi et al. 2013; Girase et al. 2011; Shu et al. 2017; Sohrabnezhad et al. 2014). The surface of spent CNC when scanned under scanning electron microscope (SEM) gives considerable idea of the active and non-active materials. As can be seen in Fig. 10.9g the surface of pure MMT which is not active against microbes is clean and clear while the modified MMT nanocomposite (Fig. 10.9h) after treatment with E. coli showed deposition of dead bacteria cells on the surface indicating its high antibacterial activity (Wang et al. 2008a; b). The type of NP doped and matrix used also impact the activity to considerable level. The Cu NP/MMT NC observed to show high mortality rates with order E. faecalis > E. coli > S. aureus > P. aeruginosa, however incorporation of oxide of copper into MMT NC induce bactericidal effect against E. coli (Bagchi et al. 2013;Sohrabnezhad et al. 2014). Significant enhancement is also observed when these Cu NPs modified MMT were added in different amounts in polymeric layer of Low density polyethylene (Bruna et al. 2012). The use of both metal and metal oxide NPs doping into the clay structure, i.e., Ag/ZnO-bentonite, has given significant enhancement in the antimicrobial activity as compared to the Ag-bentonite and ZnO-bentonite. The change in ratios of Ag/ZnO also impacted the response of the prepared CNC towards the selected strains, i.e., E. coli and E. faecalis (Motshekga et al. 2013). The antibacterial activity of these doped CNC is considerably affected by the synthetic approach also. The silver NP prepared ex-situ followed by doping into clay showed remarkable variation in their response towards E. coli as compared to the ones prepared using in-situ synthesis. Three techniques used for Ag NP synthesis, i.e., calcination, UV, and NaBH4 reduction when used for in-situ Ag-clay nanohybrid synthesis resulted in formation of similar sized particles which hence conforms to same CFU value/mL (Fig. 10.9e) in contrast to the ex-situ techniques which produce different particle sizes and hence varied activity as shown by Fig. 10.9f. The studies also pointed to the fact that nanohybrid clays in both the cases (in-situ and ex-situ) gave better activity than the pure Ag NP, unmodified clay (PGN clay), and organically modified clay (I 44P) as presented in Fig. 10.9e, f (Girase et al. 2011).

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Fig. 10.9 (a, b) TEM images of absorption of Ag-ZnO/Halloysite nanotubes into E. coli and its accumulation in bacterial periplasmic compartment, (c) release of Cu+2 ions from the Cu NP/MMT as a function of time, (d) SEM image of water suspension showing released Cu NP from Cu NP/MMT, (e) antibacterial activity against E. coli of I 44P clay, pure Ag, Ag-clay nanohybrid in-situ synthesized by calcination, UV and NaBH4 reduction methods, (f) Antibacterial activity against E. coli of PGN clay, pure Ag, Ag-clay nanohybrid ex-situ synthesized by calcination, UV and NaBH4 reduction methods, (g) SEM image of pure MMT after treatment with E. coli, and (h) SEM image of quarternized chitosan/organic MMT after treatment with E. coli showing dead bacteria cells on surface (Bagchi et al. 2013; Girase et al. 2011; Shu et al. 2017; Wang et al. 2008a, b)

By using pillaring phenomenon of inorganic metals on MMT, a harmful alga from water was removed. The MMT doped with Cu (II)/Fe (III) oxides was able to remove 94.8% cyanobacterial Microcystis aeruginosa in just 20 min contact time. These Cu (II)/Fe (III) oxides supporting on pillared montmorillonite were completely magnetic ensuring their easy separation. The cytotoxicity of this CNC increases with lowering in pH, increase in ionic strength and presence of Ca+2 ions indicating that charge-related mechanism controls the decaying of algal cells. The decrease in pH facilitates the reduction in negative charge prevailing on the surfaces of algal cells and CNC leading to interaction between these two. The surface potential of the two is also varied by interaction with the Ca+2 cations which also decrease the electrostatic repulsion and hence promotes enhanced interaction leading to removal of algae (Gao et al. 2009). Such high removal rates reveal the effectiveness of pillared clay nanocomposite in water decontamination. Ag NPs as discussed earlier can considerably enhance the antimicrobial activity of the clays. These NPs among other heavy metal NPs are most effective in removing and degrading of the microbes, hence work best for antimicrobial activity from ancient times (Petrik et al. 2012). In this regard many Ag NP based CNC like Ag NP/alumina/mixed clay (quartz, calcite, feldspar, and mica) (Yahyaei et al. 2014), Ag/poly (4-vinylpyridine)/halloysite nanotubes composite (Zhang et al. 2012), and Ag/chitosan/clay (Rhim et al. 2006) have been prepared and checked for their decontamination potential towards both Gram-positive and Gram-negative bacteria with promising results. Table 10.4 details few other combinations of CNC that have been used successfully for treating microbial impurities.

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Table 10.4 Antimicrobial activity of clay nanocomposites CNC Organo-MMT and polymerchlorhexidine acetate nanocomposites Pillared MMT-CuO + Fe2O3 magnetic material

Method/ mechanisms Growth inhibition

Microbe E. coli and S. aureus

Magnetic removal by coagulation

M. aeruginosa (cyanobacteria)

MMT-nanosilicate platelets (NSP) modified with three surfactants. Sodium dodecyl sulfate (cationic surfactant NSQc), Tallow alkyl amine (anionic (NSQa) and Triton X-100 (nonionic (NSQb) along polyurethane (PU) MMT-Cu and bacterial cellulose (BC) nanocomposites

Growth inhibition or microbiostatic effect

E. coli and S. aureus

Growth inhibition and degradation

E. coli and S. aureus

MMT with Cu2+/ LDPE clay-polymer nanocomposites (CPNs)

Degradation

E. coli

Vermiculite with Ag and Cu in polyethylene (PE) nanofillers

Growth inhibition

E. faecalis

Chitosan/MMT

Flocculation

Microcystis aeruginosa

Remarks Strong inhibition activity against both microbes ~ 18.3 mm inhibition zone CuO and Fe2O3 (1:1) ratio with pillared MMT showed 48% removal of cyanobacteria while CuO and Fe2O3 (2:1) ratio with MMT showed 92% removal NSQc depicted most eye catching results; addition of only 1% of NSQc along with PU showed highest antimicrobial activity than NSQa and NSQb NSPs

BC-Cu-MMT gave maximum reduction for E. coli which was 77.9% and for S. aureus which was 74.1% Results indicated that antibacterial effect of CPNs increases with the proportion of MtCu2+ added, at 4% MtCu2+ 94% of reduction occur Good antibacterial behavior after 96 h in comparison to pure PE. The number of colonies decreased from 1.5  108 CFU per mL to zero The NC showed efficient removal of the MA within agitation time 16–50 min resulting in removal efficiency of 94.7%

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Conclusion

The organo-nanoclays and clay nanocomposites underwent several structural and surface functional modifications that have enabled these clays to behave as effective adsorbents not only for inorganic contaminants but for organic pollutants as well. Many approaches have been used to modify these naturally available clays that include surface modifications by use of quaternary ammonium ions, incorporation of metals and their oxides, merging them with polymers, bio-materials or amalgam of all these. The effectiveness of these modifications impacts the uptake capacities by altering the electrostatic interactions or other forces that ensures bonding. These forces are responsible for not only high adsorption capacities of these modified clays but have also acted as indirect source to induce the cytotoxicity against microorganisms. The domain of nanoclays and their composites still carries a whole new era to be discovered and exploited.

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Chapter 11

Application of Nano-Photocatalysts for Degradation and Disinfection of Wastewater Jayaseelan Arun, Vargees Felix, Marudai Joselyn Monica, and Kannappan Panchamoorthy Gopinath

Contents 11.1 11.2 11.3

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Nano-Photocatalyst . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.3.1 Nanomaterials as Photocatalysts . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.3.2 Modification of Photocatalyst . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.4 Mechanism of Disinfection . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.4.1 Photocatalytic Disinfection Mechanisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.4.2 Extracellular Target Sites . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.5 Future Perspectives . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.6 Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

11.1

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Introduction

Water pollution has the major impacts on health, aquatic ecosystems, economic, sustainability of environment, and social welfare of the society. Demand on water quality has increased throughout the world in recent days. Inefficiency of current methods to provide purified water leads the way to new and effective techniques to satisfy the needs. Water disinfection was an important asset in recent days to fulfill the water needs globally. The most preferred chlorination method was ineffective due to development of resistant waterborne pathogens. Application of nanoparticles for water purification has attained vision in recent days worldwide. J. Arun · V. Felix · M. J. Monica · K. P. Gopinath (*) Department of Chemical Engineering, SSN College of Engineering, Kalavakkam, Tamil Nadu, India e-mail: [email protected] © Springer Nature Switzerland AG 2019 R. Prasad, K. Thirugnanasambandham (eds.), Advanced Research in Nanosciences for Water Technology, Nanotechnology in the Life Sciences, https://doi.org/10.1007/978-3-030-02381-2_11

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Nanotechnology provides various solutions for wastewater treatment and related environmental problems (Gupta et al. 2015). Nanowires, nanotubes, nanofilms, and quantum dots are the different forms of nanomaterials used in wastewater treatment process. Catalyst modification methods were developed to shift the light adsorption capacity at visible spectrum range (Caballero et al. 2009). Posttreatment methods are preferred to regenerate the catalysts in case of costlier catalyst. Immobilization of catalyst onto the surface of glass, polymers, metals, and fibers was carried out to recover the catalyst (Sunnotel et al. 2010; Kubacka et al. 2009; Coutinho and Gupta 2009). Photocatalysis is based on the photoexcitation of semiconductor oxide after absorption of light radiation (Reddy et al. 2007). Photocatalysis was an effective sterilization process against L. acidophilus, S. cerevisiae, and E. coli (Matsunaga et al. 1985). Photocatalysis was used for water disinfection by many researchers worldwide. Economical feasibility of photocatalytic disinfection has been explored widely to overcome environmental challenges generated by the modern society (Booshehri et al. 2017).

11.2

Wastewater

Population growth, industrialization, and environmental challenges have made water quality and sufficiency as a huge concern. Poor water quality influences numerous regimes on human welfare, economic and social implications. Water disinfection was preferred to kill pathogens and is proficient through different strategies like chlorination, ultraviolet treatment, and ozonation. Chlorination resulted in evolution of resistant waterborne pathogens, and furthermore tends to form disinfection by-products (DBP) when chlorine is added to water. Since some of the waterborne microscopic organisms have mutated there, this prompts the development of higher measures of DBP (Nieuwenhuijsen et al. 2000). Ozonation was most expensive than chemical disinfection methods and would be able to shape the destructive bromate when ozone responds with bromide ions in water. Like ozone treatment, UV treatment does not leave any residues in treated water and thus offers no recontamination in the water. Subsequently, new methodologies were needed to be considered to improve purification standards. Nanomaterials can demonstrate an assortment of enhanced properties from their mass scale due to the increase in surface area. Without a doubt, the industrialization was vital for the economic development. Be that as it may, it is dependably associated with a cost paid in regard to environmental contamination (Lade et al. 2015; Dai et al. 2016). Moreover, the discharge of various toxins with or without fractional medications, which are dangerous in nature, particularly industrial-based unique dyes or potentially dyes-based poisonous wastewater effluents (Zucca et al. 2016; Bilal et al. 2016), is a noteworthy natural concern. The logical writing has demonstrated that every year tons of dyes were produced, and almost 10–15% of the aggregate dyes get released in encompassing condition because of the low yields of textile procedures (Robinson et al. 2001). Notwithstanding the textile industries,

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dyes are likewise being utilized as a part of numerous ventures including paper and pulp, food, cosmetic, pharmaceutical, tannery, photographic, and plastic ventures. Ramírez-Montoya et al. (2015), reported the decolorization of dyes with various molecular properties using free and immobilized laccases from Trametes versicolor. In addition, inferable from their carcinogenic, genotoxic, and additionally mutagenic nature, the release of these substance operators into the primary water streams genuinely imperiled the harmony of natural biosystem (Sathishkumar et al. 2014) along these lines posturing health-related issues. In this manner, the successful treatment of dyes-based effluents without delivering any optional contamination is a key to hinder environment crumbling. During the previous quite a long while, the natural specialists are consistently attempting to build up another or enhance the current imaginative innovations. Clearly, different treatment approaches have been endeavored for the treatment of dyes or effluents (Dos Santos et al. 2007). Consequently, extra exertion and more extensive approval for industrial applications are compulsory to ease this hazardous issue. The modern usage of chemicals, as a green catalyst for remediation purposes has gotten bunches of intrigue as of late, since they work in a more extensive pH, higher temperature, higher saline concentration (Rao et al. 2014).

11.3

Nano-Photocatalyst

11.3.1 Nanomaterials as Photocatalysts Nanoparticles photocatalytic reaction depend on collaboration of light energy with metallic nanoparticles and are of incredible interest because of their wide and high photocatalytic exercises for different pollutants (Akhavan 2009). Typically these photocatalysts are used as semiconductors that can degrade assortment of persistent organic pollutants in wastewater, for example, dyes, detergents, pesticides, and volatile organic compound (Lin et al. 2014). Besides, semiconductor nano-catalyst is likewise effective for degradation of halogenated and nonhalogenated organic compounds and furthermore heavy metals in particular situation (Adeleye et al. 2016). Semiconductor nanomaterials are required a mild operations conditions and extremely compelling even at a small concentration. The simple mechanism of photocatalysis depends on the photoexcitation of electron in the catalyst. The UV light with occurrence of TiO2 produces gaps (h+) and left electrons (e-) in the conduction band. In an aqueous media, the holes (h+) are caught by water molecules (H2O) and produce hydroxyl radicals (˙OH). The radicals are unpredictable and intense oxidation agent. These hydroxyl radicals on reaction oxidize the organic pollutants into water and gaseous degradation products. Among different nano-photocatalysts created up till now, TiO2 is a standout among the most generally connected in photocatalysis because of its high reactivity under ultraviolet light (k < 390 nm) and synthetic security. Correspondingly, ZnO has additionally been widely examined for its photocatalytic activity, as it contains wide band hole simply like TiO2 (Lin et al. 2014).

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Productivity was contingent upon various factors, for example, bandgap energy, particle size, dosage, toxin fixation, and pH. For example, Hayat et al. (2011) found that the photocatalytic degradation productivity of ZnO diminished with high calcination temperature that expands the molecule estimate due to agglomeration. Compact discs nanoparticles have been pulled in serious interest as photocatalyst for treatment of industrial dyes in wastewater (Zhu et al. 2009). In spite of the fact that the previously mentioned catalyst are notable for their photocatalytic action be that as it may, they are dynamic just under ultra violet radiations (l < 387 nm). This is because of the wide bandgap energy, i.e., 3.2 eV as in the event of TiO2. In this manner, advance adjustments for the catalyst have been concentrated to build their exercises under noticeable light source (daylight) for degradation of organic pollutants (Dutta et al. 2014).

11.3.2 Modification of Photocatalyst The utilization of noticeable amount of light in photocatalytic treatment of wastewater is contemporary. To accomplish this objective, the nano-material/ semiconductor requires a few changes to diminish the bandgap energy frame UV to noticeable district. There are number of accessible examinations assessing photocatalytic action of changed nano-catalyst under visible light. The general strategies utilized for change of the catalyst incorporate dye sharpening, doping metal impurities, crossover nanoparticles or composites utilizing thin bandgap semiconductors, or anions (Ni et al. 2007). The new metals and anions in the composite make a thin bandgap additionally called as impurity energy levels, which upon introduction to visible light directs electron into semiconductor for starting the catalytic reaction (Qu et al. 2013). ZnO and TiO2 nanomaterials have wide bandgap of 3.2 eV and have been broadly examined for their photocatalytic action. Be that as it may, in solar spectrum both catalysts can as it were assimilate a little bit of the UV district, which diminish their efficiency (Chen and Zhou 2004). In any case, adjustment in catalyst by stacking metals on its surface can solve this issue. The altered composite material reduces the bandgap energy in this manner exchange the excited electron to semiconductor under illumination of solar radiation. Numerous advantages were raised once the Nanocatalyst was modified as shown in Fig. 11.1. Moreover, not all conductive metals are viable for doping to enhance photocatalytic activity, e.g., Pt and Ru are incapable for doping, while different metals, for example, Au, Ag, and Pd demonstrated excellent photocatalytic activities (Barakat et al. 2013) during late years, different doped nano-catalyst have been developed, for example, ZnO:Co, Ni, ZnS: Mn, ZnS:Cu, CdS:Eu, CdS:Mn, ZnSe:Mn, ZnS:Pb,Cu (Chandrakar et al. 2015). There are numerous dopants, for example, Cr, Si, Co, Mg, Mn, Fe, Fe, Al, In, and Ga are utilized having capacity to upgrade the surface area of metal oxide nanostructure (Jamal et al. 2012). Among different dopant, anions, for example, nitrogen, are

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Fig. 11.1 Various advantages of nanocatalysts

likewise considered as most practical and cost effective for modern application (Qu et al. 2013). Doping of nanomaterials mitigate the surface area of the catalyst and secures the nanocomposite from size reduction, change fit as a fiddle. Eskizeybek et al. (2012) incorporated altered ZnO nanocomposite utilizing organic homopolymer polyaniline (PANI). The changed PANI/ZnO nano-catalyst demonstrated 99% of removal of organic pollutants in wastewater, for example, methylene blue and malachite green dyes during photocatalysis, even with a little dosage of catalyst, i.e., 0.4 g/L of wastewater. Dutta et al. (2014) integrated γ-Fe2O3 nanoparticles by warm disintegration technique and observed a high photocatalytic movement toward degradation of rose Bengal and methylene blue dyes under illumination of visible light. The utilization of grapheme for adjustment of catalyst is likewise getting consideration of analyst due its extraordinary attributes. The composites of graphene with other semiconductor materials could altogether upgrade the electron mobility through interfacial electron exchange process amid photocatalysis. Also, graphene has capacity to advance the partition effectiveness of photograph instigated electrons and gaps in photocatalyst (Li et al. 2016).

11.4

Mechanism of Disinfection

11.4.1 Photocatalytic Disinfection Mechanisms Characterizing the mechanism which brings about microbial inactivation has been the distraction of numerous scientists. There is still much debate over which process or set of procedures lead to death of an organism passing to photocatalytic activity.

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Fig. 11.2 Nanoparticle disinfection mechanism of action

Be that as it may, a large portion of the examiner now demonstrates that the devastation of the cell layer is a vital procedure for inactivation. It has been known for some time that, if given adequate time, photocatalysis would in the long run oxidize the greater part of the organic material which constitutes microorganisms (Greist et al. 2002). As much as 96% of the cell’s dry weight is made out of organic macromolecules, which incorporate proteins, polysaccharides, lipids, lipopolysaccharide and, what’s more, nucleic acids (DNA and RNA). Monomers are 3% of the dry weight and incorporate amino acids, sugars, nucleotides, and monomer precursor and the 1% was inorganic particles (Magdigan and Martinko 2006). It has been effectively demonstrated that TiO2 can incur extreme damage on molecules, for example, amino acids (Hidaka et al. 1997a, b; Tran et al. 2006) and DNA (Yang and Wang 2008), when treated in detachment. Figure 11.2 elaborates the photocatalytic disinfection mechanism of nanocatalysts. Be that as it may, huge numbers of these particles are complex and collectively have distinctive methods of chemical and physical versatility to guarantee the survival of the cell. There are, for instance, a few classes of phospholipids, every one of which exists in numerous sorts due to variability in fatty acids composition among species based on genetic contrasts and development conditions (Oleary 1962; Magnuson et al. 1993). Not exclusively are the individual properties of these constituents essential, yet additionally the way in which they are basically organized (for instance, in lipid bilayers or on the other hand cross-linked in peptide chains) gives extra opposition to inactivation.

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11.4.2 Extracellular Target Sites The cell membrane and cell wall are evident focuses in the extracellular condition. Be that as it may, these external layers are themselves complex structures with various sites for attack. In the instance of microscopic organisms, no less than three arrangements can be set up: (1) Peptidoglycan layer, introduce in both Gram-negative and Gram-positive bacteria; (2) the lipopolysaccharide layer (LPS), the peripheral layer found just in Gram-negative microscopic organisms; and (3) the phospholipid bilayer—two in Gram-negative and one in Gram-positive organisms.

11.4.2.1

Peptidoglycan Layer

Not much has been distributed on the impacts of photocatalytic attack of the peptidoglycan layer. Peptidoglycan is a peptide-cross-linked polysaccharide grid that encompasses the cell. The essential structure is a sheet shaped from singular strands of peptidoglycan lying contiguous each other. In Gram-positive microscopic organisms, this layer can represent as much as 90% of the cell wall with a few (up to 25) sheets stacked upon each other. In Gram-negative microbes, it makes up just around 10% of the cell wall. Peptidoglycan presents unbending nature to keep up shape and internal pressure. It is likely that the peptidoglycan layer might be defenseless to radical attack (Lu et al. 2003). In any case, in both Gram-negative and Gram-positive microorganisms, peptidoglycan is extremely permeable and permits particles of around 2 nm to go through (Demchick and Koch 1996). While these pores are adequate to permit the section of oxidative species, for example, the hydroxyl radical and superoxide, it might at present be troublesome for such particles to saturate to the inward film due to their reactivity in the outside condition. On the off chance that inactivation were a capacity of peptidoglycan thickness, Gram-positive microscopic organisms would be anticipated that would pick up the preferred standpoint in surviving photocatalytic attack. This has been recommended in the writing by analysts who have uncovered types of the two gatherings to the same photocatalytic conditions (Pal et al. 2007; Dalrymple et al. 2010). Notwithstanding, the examination between inactivation rates of Gram positive furthermore, Gram-negative species construct just in light of peptidoglycan layer thickness is of little value because of the distinction in area of the layers in each gathering. Actually, the examination depends on the many-sided quality and thickness of the cell wall all in all. In any case, it has not been built up whether the peptidoglycan layer is a real basic focus of attack by radicals or on the other hand capacities to just retard the dispersion of oxidants to the basic indispensable locales, especially in Gram-positive microbes. While a few scientists observed a reduction in inactivation rate with expanded layer density and intricacy (Kühn et al. 2003), it is as yet hard to make complete decisions about the part of peptidoglycan in the instrument of inactivation (or in the obstruction thereof), since cell wall many-sided quality is a shapeless

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parameter. It isn’t known whether thought was given in these examinations to other unobtrusive contrasts among the life forms, for example, membrane lipid substance and composition.

11.4.2.2

Intracellular Target Sites

The complex mixture of substances and structures in the cell is all things considered called the cytoplasm. In the cytoplasm, there exist DNA, RNA, and ribosomes alongside other broke down or suspended materials. Every one of these substances are essential for the best possible working of the living being. Obviously, the cytoplasm is secured by the cell layer, which is impermeable to most essential molecules and particles. All different substances enter through active transport or diffuse through trans-film proteins, whose channels open and close agreeing to the necessities of the cell. The three major central point’s restrict the availability of intracellular target sites are: (1) under typical water treatment conditions (pH 5–8), photocatalyst particles aggregate to shape composites more prominent than 300 nm (French et al. 2009; Yurdakal et al. 2007); (2) it is typically accepted that oxidation happens through surfacebound radicals which are most certainly not allowed to diffuse into the cell (Mills and Le Hunte 1997); (3) regardless of whether a few radicals move towards becoming idle in arrangement, they would be exceedingly receptive and liable to experience oxidizable substrate on the cytoplasmic membrane. This implies that attack of intracellular constituents must happen through the age of different oxidants, for example, lipid radicals, hydrogen peroxide and superoxide, or surface-bound radicals on particles which break the film because of large perforation in the layer (Huang et al. 1997). Aside from coordinate attack, superoxide and hydrogen peroxide can deliver hydroxyl radicals in the intracellular condition through the Fenton reaction including “free” iron (Stefan and Irwin 1999). Numerous cells tend to direct their iron take-up as a component of barrier against the formation of the more dynamic hydroxyl radical framed from the Fenton reaction (Dubrac and Touati 2002) In any case, once created inside the cell, the hydroxyl radical is allowed to attack biomolecules.

11.4.2.3

Nucleic Acids

DNA is especially powerless to oxidative stress. ROS may attack DNA either at the sugar or at the base, offering ascend to an extensive number of products (Hidaka et al. 1997a, b). Attack at the sugar at last prompts sugar fragmentation, base loss, and a strand break with a terminal garmented sugar residue. Harm to nucleic acid in photocatalytic frameworks shows up as an indirect result from the age of superoxide By utilizing strains of E. coli, inadequate in qualities known to give protection from ROS and control iron take-up, Gogniat and Dukan (2007) proposed that DNA harm came about because of hydroxyl radical attack produced by the Fenton reaction. They observed an increase in defenselessness to photocatalysis by the mutant strains,

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especially during recovery. Likewise, this may support the observation that DNA can be attacked by no less than two unique modes, which incorporate direct hydrogen peroxide or superoxide attack and Fenton reaction generation radicals.

11.4.2.4

Reactive Oxygen Species (ROS)

It has been accounted for that the rate of photocatalytic degradation is totally relied upon surface scope of the photocatalyst by the pollutants being examined (Vinodgopal and Kamat 1992). With a specific end goal to have a viable rate of photoinactivation the microorganism must have surface connections with the photocatalyst (Mailander and Landfester 2009). The receptive oxygen species (ROS) created during photocatalysis will in the first place harm the cell mass of the microorganisms (Jiang et al. 2008). The ROS first separates the lipopolysaccharide layer of the cell wall; this at that point prompts the attack of the peptidoglycan layer, peroxidation of the lipid film and oxidation on the proteins membrane. The harm to these layers causes the potassium particles from the bacterial cells to spill, which influences the cell feasibility. This leakage prompts the loss of pivotal cell capacities and at last cell death. The principal course delineated is the immediate communication of the composite with the bacterial cell. This could occur by the surface oxidation which will bring about the disintegration of the Ag and Zn particles. The conceivable communication may likewise be achieved by coordinate contact of the composite particles with the bacterial cell divider by electrostatic collaboration. The second conceivable technique is by the disturbance of the bacterial cell wall by potential ions or attack by the ROS produced. The third method for inactivation is the adjustment/hindrance of the DNA replication by the association of the particles or the ROS with sugar-phosphate bunches causing quality adjustment, in this manner modifying the protein articulation functioning of cell working. At last, the potential disturbance of the membrane prompts the arrival of the intracellular materials, which in the end prompts cell lysis (Matai et al. 2014).

11.5

Future Perspectives

More study is required to reassess the eco-toxicity potential for each new development in catalyst and for existing materials. Furthermore, life cycle appraisals of the nanomaterials to be addressed. In addition, additional work is required on building up a cost-effective method for synthesizing nanomaterials. For example, there are relatively few reports concerning treating wastewater to test the adequacy of ZnO or testing for long-term execution of these nanotechnologies. Research is additionally required especially on advancements to hold or immobilize the nanomaterials. Future advancement faces a few obstacles—specialized, cost, and social strategies. The substantial scale execution of immobilized catalysts can help in decreasing the hardness of lethal effluents from industries, destructive chemicals and

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pharmaceuticals from local sewage, etc. The mechanism between photocatalytically prompted lipid peroxidation and cell passing need to be resolved. Future work should quantitatively relate lipid peroxidation and photocatalytic vulnerability to decrease the cell feasibility in order to be efficient in treatment process.

11.6

Conclusion

In present situation, there is an urgent requirement for advanced wastewater treatment technologies to guarantee high-quality water with eliminated synthetic and natural contaminants. Nanotechnology is one of the perfect candidates to propel wastewater treatment processes. Different nanomaterials have been produced and examined effectively for wastewater treatment. These incorporates nano-adsorbents (Fe, MnO, ZnO, MgO, CNT), photocatalysts (ZnO, TiO2, CdS, ZnS:Cu, CdS:Eu, CdS:Mn), and electrocatalysts (Pt, Pd). The nano-adsorbents can possibly evacuate heavy metals like Cr, As, Hg, Zn, Cu, Ni, Pb, and Vd from wastewater. Nanophotocatalysts can be utilized for treatment of both lethal contaminants and heavy metals. Photocatalysis was a promising and green mechanism for degradation of contaminants that pollute water assets. Recently numerous specialists have begun to study photocatalysts-based wastewater degradation process. The utilization of nanophotocatalysts was a phenomenal choice for wastewater treatment since wastewater can be reused and waste generation were decreased. The overall treatment process is basic, simple, and it is practicable in day-to-day industrial process.

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Kubacka A, Ferrer M, Cerrada M-L, Serrano C, Sánchez-Chaves M, Fernández García M (2009) Boosting TiO2-anatase antimicrobial activity: polymer-oxide thin films. Appl Catal B 89:441–447 Kühn K-P, Chaberny I-F, Massholder K, Stickler M, Benz V-W, Sonntag H-G (2003) Disinfection of surfaces by photocatalytic oxidation with titanium dioxide and UVA light. Chemosphere 53:71–77 Lade H, Govindwar S, Paul D (2015) Mineralization and detoxification of the carcinogenic Azo dye Congo red and real textile effluent by a polyurethane foam immobilized microbial consortium in an upflow column bioreactor. Int J Environ Res Public Health 12:6894–6918 Li W, Zhang Y, Tian G, Xie S, Xu Q, Wang L, Tian J, Bu Y (2016) Fabrication of graphenemodified nano-sized red phosphorus for enhanced photocatalytic performance. J Mol Catal A Chem 423:356–364 Lin S-T, Thirumavalavan M, Jiang T-Y, Lee J-F (2014) Synthesis of ZnO/Zn nano photocatalyst using modified polysaccharides for photodegradation of dyes. Carbohydr Polym 105:1–9 Lu Z-X, Zhou L, Zhang Z-L, Shi W-L, Xie Z-X, Xie H-Y (2003) Cell damage induced by photocatalysis of TiO2 thin films. Langmuir 19:8765–8768 Magdigan M-T, Martinko J-M (2006) Brock biology of microorganisms, 11th edn. Pearson Education, Inc., Upper Saddle River, NJ Magnuson K, Jackowski S, Rock C-O, Cronan J-E Jr (1993) Regulation of fatty acid biosynthesis in Escherichia coli. Microbiol Mol Biol Rev 57:522–542 Mailander V, Landfester K (2009) Interaction of nanoparticles with cells. Biomacromolecules 10:2379–2400 Matai I, Sachdev A, Dubey P, Kumar SU, Bhushan B, Gopinath P (2014) Antibacterial activity and mechanism of Ag–ZnO nanocomposite on S. aureus and GFP-expressing antibiotic resistant E. coli. Colloids Surf B Biointerfaces 115:359–367 Matsunaga T, Tomoda R, Nakajima T, Wake H (1985) Photoelectrochemical sterilization of microbial cells by semiconductor powders. FEMS Microbiol Lett 29:211–214 Mills A, Le Hunte S (1997) An overview of semiconductor photocatalysis. J Photochem Photobiol A 108:1–35 Ni M, Leung M-K, Leung D-Y, Sumathy K (2007) A review and recent developments in photocatalytic water-splitting using TiO2 for hydrogen production. Renew Sust Energ Rev 11:401–425 Nieuwenhuijsen M-J, Toledano M-B, Eaton N-E, Fawell J, Elliott P (2000) Chlorination disinfection byproducts in water and their association with adverse reproductive outcomes: a review. Occup Environ Med 57:73–85 Oleary W-M (1962) The fatty acids of bacteria. Bacteriol Rev 26:421–435 Pal A, Pehkonen S-O, Yu L-E, Ray M-B (2007) Photocatalytic inactivation of gram-positive and gram-negative bacteria using fluorescent light. J Photochem Photobiol A 186:335–341 Qu X, Alvarez P-J, Li Q (2013) Applications of nanotechnology in water and wastewater treatment. Water Res 47:3931–3946 Ramírez-Montoya L-A, Hernández-Montoya V, Montes-Morán M-A, Jáuregui-Rincón J, Cervantes F-J (2015) Decolorization of dyes with different molecular properties using free and immobilized laccases from Trametes versicolor. J Mol Liq 212:30–37 Rao M-A, Scelza R, Acevedo F, Diez M-C, Gianfreda L (2014) Enzymes as useful tools for environmental purposes. Chemosphere 107:145–162 Reddy M-P, Venugopal A, Subrahmanyam M (2007) Hydroxyapatite-supported Ag-TiO2 as Escherichia coli disinfection photocatalyst. Water Res 41:379–386 Robinson T, McMullan G, Marchant R, Nigam P (2001) Remediation of dyes in textile effluent: a critical review on current treatment technologies with a proposed alternative. Bioresour Technol 77:247–255 Sathishkumar P, Kamala-Kannan S, Cho M, Kim J-S, Hadibarata T, Salim M-R, Oh B-T (2014) Laccase immobilization on cellulose nanofiber: the catalytic efficiency and recyclic application for simulated dye effluent treatment. J Mol Catal B Enzym 100:111–120

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Stefan I-L, Irwin F (1999) Superoxide and iron: partners in crime. IUBMB Life 48:157–161 Sunnotel O, Verdoold R, Dunlop P-S-M, Snelling W-J, Lowery C-J, Dooley J-S-G (2010) Photocatalytic inactivation of Cryptosporidium parvum on nanostructured titanium dioxide films. J Water Health 8:83–91 Tran T-H, Nosaka A-Y, Nosaka Y (2006) Fourier transform reflection absorption IR spectroscopy study of Formate adsorption on TiO2. J Phys Chem B 110:25525–25531 Vinodgopal K, Kamat P-V (1992) Photochemistry on surfaces: photodegradation of 1,3diphenylisobenzofuran over metal oxide particles. J Phys Chem 96:5053–5059 Yang X, Wang Y (2008) Photocatalytic effect on plasmid DNA damage under different UV irradiation time. Build Environ 43:253–257 Yurdakal S, Loddo V, Bayarri Ferrer B, Palmisano G, Augugliaro V, Gimenez Farreras J (2007) Optical properties of TiO2 suspensions: influence of pH and powder concentration on mean particle size. Ind Eng Chem Res 46:7620–7626 Zhu H, Jiang R, Xiao L, Chang Y, Guan Y, Li X, Zeng G (2009) Photocatalytic decolorization and degradation of Congo red on innovative crosslinked chitosan/nano CdS composite catalyst under visible light irradiation. J Hazard Mater 169:933–940 Zucca P, Neves C, Simões M-M, Neves M-D-G-P, Cocco G, Sanjust E (2016) Immobilized lignin peroxidase-like metalloporphyrins as reusable catalysts in oxidative bleaching of industrial dyes. Molecules 21:964

Chapter 12

Degradation of Emerging Contaminants Using Fe-Doped TiO2 Under UV and Visible Radiation Irwing M. Ramírez-Sánchez, Oscar D. Máynez-Navarro, and Erick R. Bandala

Contents 12.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.2 Methodology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.3 Photocatalytic Activity of Fe-Doped TiO2 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.4 Effects of Fe Load in TiO2 Photocatalytic Activity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.5 Synthesis Method for Fe-Doped TiO2 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.6 Co- and Multi-Doping of Fe-Doped TiO2 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.7 Simulated Solar and Solar Stimulation of Fe-Doped TiO2 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.8 Emerging Contaminants Degradation Using Fe-Doped TiO2 . . . . . . . . . . . . . . . . . . . . . . . . . . 12.9 Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

12.1

263 266 266 275 276 278 278 279 280 281

Introduction

Emerging contaminants (EC) are chemical compounds not currently included by existing local or international water quality regulation, and with known or potentially threats to the environmental ecosystem, human health and safety (Dulio et al. 2018). The most common ECs include algae toxins, illegal drugs, industrial I. M. Ramírez-Sánchez (*) Department of Civil, Architectural and Environmental Engineering, The University of Texas at Austin, Austin, TX, USA O. D. Máynez-Navarro Universidad de las Américas Puebla (UDLAP), Ex-Hacienda Santa Catarina Mártir, Cholula, Puebla, Mexico E. R. Bandala Desert Research Institute (DRI), Las Vegas, NV, USA e-mail: [email protected] © Springer Nature Switzerland AG 2019 R. Prasad, K. Thirugnanasambandham (eds.), Advanced Research in Nanosciences for Water Technology, Nanotechnology in the Life Sciences, https://doi.org/10.1007/978-3-030-02381-2_12

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compounds, flame retardant, food additives, nanoparticles, pharmaceuticals (human and vet), personal care products, pesticides, biocides, steroid, synthetic and natural hormones, and surfactants (Mandaric et al. 2016). ECs possess undesirable biological effects in living organisms, some of them even at trace concentration, and it has been reported that natural attenuation, conventional drinking water or wastewater treatment processes can either only partially remove them or are incapable at all of removing ECs from water. Moreover, despite all these undesirable characteristics, most ECs remains only under scrutiny of scientific community but not regulated, so there is an urgent need for the development of treatment processes capable of their removal in water (Geissen et al. 2015). Different water treatment technologies, including Advanced Oxidation Processes (AOPs), have been studied to remove and mineralize ECs from water. In particular, the photocatalytic processes using titanium dioxide (TiO2) semiconductor as photocatalyst (as described in Fig. 12.1) has been widely used by the scientific community for ECs degradation in aqueous samples. TiO2 is one of the most commonly used photocatalysts due to its reasonable optical and electronic properties, fair photocatalytic activity, insolubility in water, chemical and photochemical stability, non-toxicity, low cost, and high efficiency in pollutant mineralization (Hashimoto et al. 2005; Fujishima et al. 2008; Tong et al. 2012; Cassaignon et al. 2013). Despite all these interesting positive features, the scientific community has also recognized the need for TiO2 improvement to become a sustainable photocatalyst (Yu et al. 2010). One of the TiO2 major drawbacks is its large band gap (Eg) more than 3.2 eV to initiate the photocatalytic reaction, limiting the photocatalytic activity of TiO2 only when it is irradiated with short frequency radiation. At the present time, several researchers have reported the synthesis of novel TiO2-based materials to substitute the ultraviolet irradiation (UV) for visible light (Visible), LED light, solar radiation, or laser irradiation as the source of energy to activate TiO2 photocatalytic processes. Since early works in the 1970s (Ghosh and Maruska 1977; Schrauzer and Guth 1977), different techniques have been tested to enhance photocatalytic activity of TiO2 and have demonstrated TiO2 is able to use photons coming from sources with visible light. In general, the strategies to control the photocatalytic activity of TiO2 include modifying its surface, coupling TiO2 with other semiconductor(s), and introducing defects into the TiO2 lattice (also known as doping). From all of these strategies, doping has attracted the attention of the scientific community because of its high potential to generate cost-effective materials. The two main techniques to introduce dopants into TiO2 catalyst are by incorporating dopants ions at the oxide surface (e.g., impregnation or ion-implantation) and inside the oxide lattice (e.g., sol–gel method). Most frequently used doping materials are transition-metal cations (e.g., Cr, V, Fe, Ni) at Ti sites and anions (e.g., N, S, C) at O sites. Choi et al. conducted a systematic study on the photocatalytic activity of TiO2-doped nanoparticles with 21 transition metal elements and found that doping with Fe3+, Mo5+, Ru3+, Os3+, Re5+, V4+, and Rh 3+ significantly increased the photocatalytic activity in the degradation of

12

Degradation of Emerging Contaminants Using Fe-Doped. . .

265

Fig. 12.1 Schematic diagram for the photocatalytic mechanism of TiO2 and ˙OH generation under photons with energy higher than 3.2 eV. According to Density Functional Theory (DFT) computations, the valence band (VB) and conduction band (CB) of pure TiO2 are mainly composed of O 2p orbitals and Ti 3d orbitals, respectively, while the Fermi level (EF) is located in the middle of the band gap (Eg), indicating that VB is full filled while CB is empty (Wen et al. 2012). Thus, when using photons with energy higher than 3.2 eV, photoexcitation of the semiconductor promote VB electrons to the CB creating a charge vacancy in the VB. The vacancy (also known as hole) in the VB can react with hydroxide ions to form ˙OH or can also be filled by absorbed organic molecules serving as electron donors. Also, photogenerated electrons in the CB can be transferred to electron acceptors and bring about ˙OH. (Eg—band gap energy, E—photon energy, OMads—adsorbed organic molecule, OMoxi—oxidized organic molecule)

trichloromethane (chloroform), while Co3+ and Al3+ dopants decreased TiO2 photoactivity (Choi et al. 1994). Among anions and cation dopants, ferric ion (Fe3+) is one of the most frequently employed. Fe3+ is considered an appropriate candidate for doping TiO2 because its ionic radius (0.69 A) is similar to Ti4+ (0.745 A), so Fe3+ can be easily incorporated into the TiO2 crystal lattice (Choi et al. 1994) demonstrated for XPS analysis (Li et al. 2008). Early studies on Fe-doped TiO2 (Fe-TiO2) were related on the

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mechanism of Fe3+ effect on TiO2 photocatalytic activity (Litter and Navío 1996), on Fe-TiO2 particle size (Zhang et al. 1998), and on Fe-TiO2 structure evolution during heat treatment (Wang et al. 2001). In this work, we summarized the results from research reports where Fe-TiO2 nanoparticles were used to degrade or mineralize water contaminants. Also, we compared the main mechanism proposed of photocatalytic activity of Fe-TiO2, and analyzed the potential of Fe-TiO2 to degrade ECs.

12.2

Methodology

Information about Fe-TiO2 was obtained from relevant and updated available scientific databases. The searched keywords were with Boolean terms which were (Fe OR iron) AND (tio2 OR titanium) AND (algae toxins OR illegal drugs OR industrial compounds OR flame retardant OR food additives OR pharmaceuticals OR personal care products OR pesticides OR biocides OR steroid OR synthetic OR hormones OR surfactants) AND (degradation AND photocatalysis). Searches were not limited to articles published after some specific year. However, it was not considered Fe-TiO2 systems for photoelectrocatalyst, gas-phase, water-splitting, photo-Fenton, or magnetic photocatalyst because those topics deserve special attention apart.

12.3

Photocatalytic Activity of Fe-Doped TiO2

Since Johnson et al. began with the characterization of Fe-TiO2 with the early interpretation of Fe narrowing TiO2 band gap (Johnson et al. 1968) and Mizushima and Iida suggested Fe3+ energy level existing near the valence band (VB) in TiO2 (Mizushima and Iida 1971), extensive research efforts have been developed for the Fe-TiO2 system. Schrauzer and Guth suggested Fe doping enhanced photocatalytic activity of TiO2 (Schrauzer and Guth 1977). Then, many studies have been carried out to improve the photocatalytic activity by reducing recombination reaction or increasing the efficiency of photocatalytic reactions (see Table 12.1). Figure 12.2 shows the compound types used to evaluate the photocatalytic activity of Fe-TiO2 where phenol and dyes (methylene blue and methyl orange) have been the most frequent model contaminants tested. Dye wastewater is produced mainly by the textile industry. Several research groups have studied dye degradation, such as acid orange 7 (Butler et al. 2016), rhodamine B (Momeni et al. 2016; Denisov et al. 2017), methylene blue (Shao et al. 2016; Shi et al. 2018), malachite green (Bhatu et al. 2018), and methyl orange (Khang et al. 2016). Although the detailed mechanism of photocatalytic process varies with the different contaminants, most researchers concur that the primary reactions responsible for TiO2 photocatalysis are the interfacial redox reactions of charge carriers (e.g., electrons and holes) generated when the semiconductor catalyst is exposed to

UV (400 W Hg lamp) and visible (200 W Xe lamp)

100:1 mole ratio TiO2: Fe

3.8, 8.2 and 9.1 at. % Fe

Solid-state reaction

Electrochemical anodization and chemical bath deposition

NanotubesN, C, co-doped Fe-TiO2

UV (200–800 nm, UVB lamp)

1, 0.5%

Sol-gel and microwave irradiation

NP N – Co-doped Fe-TiO2 Microparticles Fe-TiO2

1, 1.5, 2% Fe

Mild hydrothermal

NPFe-TiO2 Visible (λ > 400, 23 W visible lamp)

UV (15 W LHGL)

Visible (150 W halogen lamp) and UV (160 W HHgL)

3.4 wt. % Fe

Hydrothermal deposition

Fe-TiO2 immobilized on optical fibers

Visible (460 nm) and UV (235 nm)

UV-visible (sun)

0.2, 1.0, 2.0 at. %

0.3, 0.6, 1.07 wt. % Fe

UV-visible (300 W XEL)

Fe-Doping 0, 1.16, 2.01, 3.08, 4.87, 6.68 wt.% Fe 0, 5, 10, 20 mg Fe

Source radiation Visible (λ > 420 nm, 65 W FLR)

Acidolysishydrothermal

Sol-gel and Hydrothermal Sol-gel and Hydrothermal

Synthesis method Microemulsion

NPFe-TiO2

Particles type NP C co-doped Fe-TiO2 Nanospheres TiO2-Fe MicrospheresFeTiO2

Rhodamine B, N. D.

Diazinon, [10– 100], Amoxicillin, streptomycin, diclofenac, [30] Acid orange 7, [34]

Ibuprofen, carbamazepin, sulfamethoxazole, [5]

Rhodamine B, [20]

Methylene blue, [10] Basic orange 2, [50]

Contaminant, Initial conc. [mg L
1] Malachite green, [10]

N.D.

0.25, 0.5, 1

0.2, 0.4, 0.6 1

30 SOF’s of 10 cm

0.25

0.2

0.5

Catalyst load g L
1 1

Table 12.1 Photocatalytic activity and experimental setup of removal contaminants of water using Fe-TiO2

First order:: k ¼ 12.3  10
3 min
1 for 1 g L
1 First order:: k ¼ 11.2  10
3 min
1 visible, k ¼ 18.9  10
3 min
1 UV

N.D.

Langmuir Hinshelwood: k ¼ 111.2  10
3 min
1 visible, k ¼ 5.37  10
3 min
1 UV Langmuir Hinshelwood: k ¼ 1.8  10
3 min
1 for IBU, k ¼ 1.5  10
3 min
1 for CRB, k ¼ 1.4  10
3 min
1 for SMA N.D.

N.D.

N.D.

Kinetic constant N.D.

Degradation of Emerging Contaminants Using Fe-Doped. . . (continued)

Momeni et al. (2016)

Tabasideh et al. (2017) Aba-Guevara et al. (2017) Butler et al. (2016)

Lin et al. (2017)

Shi et al. (2018) VelázquezMartínez et al. (2018) Sui et al. (2018)

References Bhatu et al. (2018)

12 267

Hydrothermal

Sol-gel and impregnation

Sol-gel and impregnation

Sol–gel

NP Fe-TiO2

NPFe-TiO2

NPFe-TiO2

NPLa co-doped Fe-TiO2

Hydrothermal

Impregnation

Macroporous Fe–Ti–Oxide-M NPFe-TiO2

4% Fe-TiO2, 2% La, 2% Fe-TiO2

0.5, 1, 2, 3, 5, 10 wt. % Fe

0, 1, 2, 3 mol% Fe

1, 2, 5, 10, 20, 50% 0, 0.1, 0.3, 0.5, 0.8, 1, 2, 5 at. % 1.84, 2.35 at %

1, 3, 5, 10%

1%, 2%, 3% and 4% 0.006, 0.034%

Impregnation of Aeroxide TiO2 P25 Sol-gel

Sol-Gel

1, 3, 6, 9 wt. %

Fe-Doping 5.8%

Synthesis method Two step modified sol-gel Hydrolysis

NPFe-TiO2

NanohybridsFeTiO2/carbon nanotube NP Fe-TiO2 NPFe doped TiO2

Particles type NPFe-TiO2

Table 12.1 (continued)

UV (365 nm, 300 W Hg lamp) and visible (300 W metal halide)

UV (254 nm, 125 W LHgL) and UV-visible (365–366 nm, 125 W MHgL) Visible (λ > 420 nm, 1000 W Xe lamp)

UV (25 W MHgL)

Visible (λ > 420 nm, 150 W Xe lamp) UV (365 nm, FLR)

Visible, 3 W + LED bulb + WBH-BL-001

UV-visible (Sun) and UV lamp UV (λ < 380 nm, 27 W low pressure lamp) and visible (λ > 380, 27 W lamp)

Source radiation UV-visible, Sun (11 a.m.–3 p. m.) Visible (λ > 420 nm,150 HXeL)

5

1

Phenol, [50]

2

4

1

1

1

0.5, 1, 1.5, 2 0.25, 0.5, 1

0.5

Catalyst load g L
1 0.8

Phenol, [19.7]

Rhodamine 6G, [4.8]

Methyl orange, [10] Bromocresol green, [6.9]

Methylene blue, methyl orange, [20] Tetracycline, [20]

Phenol, [5–500]

Contaminant, Initial conc. [mg L
1] Methylene blue, [31.9] Methylene blue, methyl orange, [10] Carbendazim, [8]

Initial velocity: 3.80 μmol dm
3 min
1 UV-visible, 2.0 μmol dm
3 min
1 Visible N.D.

First order: k1 ¼ 9.712  10
3 min
1

N.D.

N.D.

N.D.

Pseudo first order: k ¼ 1.2  10
3 min
1 UV, k ¼ 8  10
3 min
1 visible N.D.

N.D.

Kinetic constant First order: k ¼ 2.8  10
3 min
1 N.D.

Shi et al. (2011)

Lezner et al. (2012)

MedinaRamírez et al. (2014) Fan et al. (2014) Wen et al. (2012) Goswami and Ganguli (2012) Delekar et al. (2012)

Kaur et al. (2016) Hemmati Borji et al. (2014)

References Shao et al. (2016) Khang et al. (2016)

268 I. M. Ramírez-Sánchez et al.

Sol-gel

Impregnation

Wet impregnation

NPS co-doped Fe-TiO2

NPFe-TiO2

NP Fe-TiO2 NPB, N co-doped Fe-TiO2 NPN co-doped Fe-TiO2

Sol-gel method with hydrothermal

Sol-gel

NP Au co-doped Fe-TiO2

Nanorods Fe-TiO2 (Rutile) MicrospheresFeTiO2 NPFe-TiO2

NPFe-TiO2

Hydrothermal with collagen fiber

Nanofibers Fe-TiO2

Acid catalyzed sol– gel

Hydro-alcohol thermal Co-precipitation

Sol-gel

NPFe-TiO2

Nonhydrolytic sol-gel

Hydrothermal method

Sol-gel

NPFe-TiO2

1.8–0.45 at % Fe

0.03–1.70 at % Fe

0.5–2.9 mole %

0.16, 0.33, 0.49and 0.65 at. % 0.002, 0.00, 0.01, 0.02, 0.03, 0.05 Fe/TiO2 molar ratio 0.57 Fe-TiO2, 2.0Au/0.57FeTiO2, 3.0Au/ 0.57Fe-TiO2 3, 2 at % Fe

0.91, 0.87, 0.85 at %

0.05, 0.1, 0.25, 0.5, 1 wt % 0.2, 0.5, 1.0 at% of Fe/Ti

1, 3, 5, 8 wt.% Fe

2, 5 mole %

2 mol%

2,4dichlorophenol, [50]

Visible (λ > 420 nm, 1000 W tungsten halogen) and UV (365 nm, 300 W HHgL)

UV (250 W HHgL)

UV-visible (254 nm, 420 nm, 8 W FLR) UV (125 W Hg lamp)

Salicylic acid, [81.5]

Mecoprop, [579.5]

Phenol, [60]

Congo red, [17.4]

Basic orange 2, [248.7]

Visible (λ > 420 nm, 150 W)

UV-vis (Sun)

Methylene blue, [10]

Phenol, [9.4]

4-nitrophenol, [1391.1] Methyl orange, [20]

Methyl orange, methylene blue, [10] 4-nitrophenol, [20]

Basic orange 2, [20]

Visible (300 W Xe lamp)

Visible (λ > 420 nm, 1000 W Tungsten halogen lamp) and UV (365 nm, 300-W HHgL) Solar radiation

UV-visible (8 W BFL)

UV (340 nm, UV lamp)

UV-visible (400 W)

Visible (400–600 nm, 160 W Philips ML)

1.66

2

1

1.25

1

1

0.25

1

1

2

0.4

1

1

First order: k ¼ 0.042  10
3 min
1 Pseudo first order: k1 ¼ 14.9  10
3 min
1

N.D.

N.D.

N.D.

N.D.

N.D.

N.D.

First order: k ¼ 28.496  10
3 min
1 N.D.

N.D.

First order reaction: k1 ¼ 3.6  10
3 min
1, k1 ¼ 1.3  10
3 min
1 N.D.

Degradation of Emerging Contaminants Using Fe-Doped. . . (continued)

Melghit et al. (2009) Li et al. (2009) Abazović et al. (2009) Popa et al. (2008)

Wu et al. (2009)

Cai et al. (2010)

Naik and Parida (2010) Fàbrega et al. (2010)

Zhao et al. (2010) Yalçın et al. (2010) Xing et al. (2010)

Hamadanian et al. (2011)

Seabra et al. (2011)

12 269

Hydrolysis and hydrothermal

Co-precipitation and hydrothermal

Impregnation of Aeroxide TiO2 P25 and calcination of Wako TiO2 Precipitationhydrothermal

Microspheres Fe-TiO2

NPFe-TiO2

NPS co-doped Fe-TiO2

2.2, 6.6, 13.2, 22  10
7 Fe ions mol
1 gcat
1 0.25, 0.5, 2.5 at % Fe

Sol-gel dip coating

Sol-gel and metal ion-implantation

Thin films on glass or silica plates NPFe-TiO2

Wet chemical synthesis

2, 4 Fe/TiO2

Hydrothermal

NPFe-TiO2

NPFe-TiO2

0.4–5.1%

Microemulsion

0.03–0.15% Fe

0.05–30 at % Fe

Sol–gel

0.5 w/w % Fe and FexTi1-xO2, x ¼ 0.005 Fe/Ti molar 0.05–2.00 mole % Fe

0.1–10 at. %

0.042–0.46 wt. %

Fe-Doping 0.1–10% M/M

Thin films on glass slidesFeTiO2 NPFe-TiO2

NPN, co-doped Fe-TiO2

Synthesis method Hydrothermal

Particles type NPFe-TiO2

Table 12.1 (continued)

1

Rhodamine B, [20]

2-Propanol, [156.3] Methanol, N.D.

UV(λ ¼310–390 nm, Xe-lamp

Active yellow XRG, [100] Methyl orange O, [100]

Phenol, [50]

Basic blue 41, [4.8]

0.5

Phenol, [1.88]

0.5

N.D.

20 g silica rings

1

0.5

Glass slide

N.D.

N.D.

N.D

N.D.

N.D

N.D.

Initial velocity: r0 ¼ 3.29  10
8 M s
1 UV-vis, r0 ¼ 0.207  10
8 M s
1 visible First order: kobs ¼ 3.5  10
3 min
1 visible

N.D.

1

1

Kinetic constant N.D.

Catalyst load g L
1 0.5

Contaminant, Initial conc. [mg L
1] Methylene blue, [100] Methyl orange, [30] (UV) and [20] (visible) Phenol, [47.1]

Visible (λ >450 nm, 100 W HHgL)

UV (λ >365 nm, 500 W HHgL) Visible (λ >380 nm, 300 W HHgL) UV-visible (Sun)

UV (365 nm, 500 W ultra HHgL) Visible (405, 436 nm, 500 W ultra HHgL) UV (365 nm, 300 W HHgL) and visible (λ 400 nm, 150 W Xe lamp) UV (450 W Osram XBO)

CHCl3, [1432.6]

Cyclohexane, N. D. EDTA and Cr(VI), [1461] 0.5

1

2.5

N.D

N.D

N.D.

Zhang et al. (1998)

Li et al. (2003) Navío et al. (1999)

N.D. No Data, Fe-TiO2 Fe-doped TiO2, NP Nanoparticles, SOF Side- glowing optical fibers, Visible Visible light, UV Ultraviolet radiation, Sun Solar light, FLR Compact fluorescent lamp, BFL Blacklight fluorescent lamps, XeL Xenon lamp, HHgL High pressure Hg vapor lamp, LHgL Low pressure Hg lamp, MHgL Medium pressure Hg lamp, HXeL Highpressure Xe lamp

NPFe-TiO2

Sol–gel method

NP Fe-TiO2

12 Degradation of Emerging Contaminants Using Fe-Doped. . . 271

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Fig. 12.2 Model compounds used to evaluate the photocatalytic activity of Fe-TiO2

radiation with enough energy, as suggested by Eqs. (12.1) and (12.2) (Wang et al. 2003) and showed schematically in Fig. 12.1. Charge carrier generation
TiO2 þ hv  Eg ! hVB þ þ eCB


ð12:1Þ

Charge carrier recombination hþ þ e
! TiO2 þ heat

ð12:2Þ

Despite the intensive research on Fe-TiO2, the mechanism involved in the Fe-TiO2 photocatalytic process remains unclear and controversial (Liu et al. 2013). It has been widely suggested that Fe doping directly influences the intrinsic properties of TiO2 by introducing additional electron energy states inside the TiO2 Eg, modifying the charge carrier recombination rates, the particle size, and interfacial electron-transfer rates, and extending its absorbance in the visible part of the spectrum. In general, the main mechanisms suggested for Fe performance in TiO2 are: 1. Insertion of two energy states (e.g., the Fe3+/Fe2+ energy level just below CB and Fe3+/Fe4+ energy level slightly above VB) (Mizushima et al. 1972; Litter and

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Navío 1996; Wang et al. 2003; Tong et al. 2008; Li et al. 2009; Yu et al. 2009; Liu et al. 2011; Elghniji et al. 2012; Medina-Ramírez et al. 2014; Shi et al. 2018). 2. Introduction of just one additional energy level below the CB (Yu et al. 2009; Naik and Parida 2010; Xing et al. 2010). According to proposed mechanism 1, pure TiO2 with Eg ¼ 3.2 eV cannot effectively produce enough photo-induced charge carriers under visible radiation (λ > 400 nm), but when Fe3+ doping is included into TiO2, iron form bulk Fe–O–Ti bond and the overlapped d orbital of Fe3+ decreases the Fe–TiO2 Eg to less than 3.2 eV (Li et al. 2009). Shi et al. reported a shift in Eg value for Fe-TiO2 from 3.24 to 2.99 eV, so Fe3+ increases visible radiation absorption of TiO2 (Shi et al. 2018). Alternatively and based on density functional theory calculations, it has been suggested that Fe3+ just introduces one single energy level below the CB (Liu et al. 2013; Sui et al. 2018), as depicted and described in Fig. 12.3. These authors suggested a hybridized band with Ti 3d and Fe 3d reduced 0.3–0.5 eV (Xing et al. 2010), 0.2–0.34 eV (Yu et al. 2009) or 1.2 eV (Naik and Parida 2010) for the Fe-TiO2. Although abovementioned mechanisms predict the improvement in the photocatalytic activity of Fe-TiO2, photocatalytic activity will depend not only on reduction of Eg, but also it is depending on several other variables, such as the catalyst crystalline array, particle size, and shape; differences in the density of hydroxyl groups on the particle surface and the number of water molecules hydrating the surface; surface area and surface charge; differences in the number and nature of trap sites; dopant concentration, localization and chemical state of the dopant ions; intensity of radiation; particle aggregation; or scavenger species in media (Serpone 1997; Abazović et al. 2009; Fàbrega et al. 2010). Consequently, different research groups have experimentally showed the degradation of contaminants using Fe-TiO2, chronologically arranged in Table 12.1. Although comparing the results reported in Table 12.1 for Fe-TiO2 is not easy because preparation methods of the samples and the experimental conditions (e.g., irradiation source and intensity) were considerably different, the reported improvements are mainly using both UV and visible radiation. For example, Choi et al. reported the highest chloroform degradation efficiency under UV irradiation for a sample containing Fe3+ and the substitution of Ti4+ ions in the TiO2 lattice (Choi et al. 1994). Zhou et al. studied the effect of Fe3+ doping concentration on the photo activity of yellow XRG dye (Zhou et al. 2006). Fe-doped TiO2 nanorods were prepared by impregnating-calcination and the photocatalytic activity was found higher than Degussa P25 and pure TiO2 nanorods (Yu et al. 2009). Fex/TiO2 nanofibers exhibited excellent visible radiation photocatalytic activity for degradation of orange II and it was observed to increase with the increase of Fe3+ content (Fex/TiO2 molar ratio 0.02–0.002) under visible radiation (Cai et al. 2010). Nanorods of 0.5% Fe-TiO2 were observed capable to increase the photocatalytic activity over twice compared to Degussa P25 (Yu et al. 2009). Overall, the enhancement in photocatalytic activity could be mainly explained by the extended lifetimes of the charge carriers and Fe3+ role as an electron trap at or near the particle surface, as suggested by Eqs. (12.3) to (12.6). The potential capability of Fe dopant to reduce charge carriers recombination is relevant because

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Fig. 12.3 Schematic diagram for the photocatalytic mechanism of Fe-TiO2 and ˙OH generation under photons with energy less than 3.2 eV. When TiO2 contains a Fe3+ ion, the Fe 3d orbitals split into two bands, one is a hybrid band (A2g) and one is midgap band (T2g) which induce a new localized Eg state (Wen et al. 2012). Thus, when TiO2 absorb photons with energy less than 3.2 eV, photoexcitation of the semiconductor promote an electron from the VB to the midgap band (T2 g), also called shallow trap, creating a pair electron/hole. The hole in the VB can react with hydroxide ions to form ˙OH, absorbed organic molecules or trap for Fe3+. Also, photogenerated electrons in the midgap band (T2g) can be transferred to Fe3+ and subsequently trapped photogenerated electron can transfer to acceptor of electrons and bring about ˙OH. (Eg—BG Energy, E—photon energy, OMads—adsorbed organic molecule, OMoxi—oxidized organic molecule)

spectroscopic studies revealed that most of the photogenerated e
/h+ pairs (c.a. 90%) recombine rapidly after excitation (Colombo and Bowman 1995), being that a significant reason for control recombination rates in TiO2. Electron trap Fe3þ þ eCB
! Fe2þ

Migration

ð12:3Þ

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Fe2þ þ Ti4þ ! Fe3þ þ Ti3þ

ð12:4Þ

Fe3þ þ hVB þ ! Fe4þ

ð12:5Þ

Fe4þ þ OH
! Fe3þ þ OH

ð12:6Þ

Hole trap

Migration

12.4

Effects of Fe Load in TiO2 Photocatalytic Activity

Studies mentioned in Table 12.1 show an appropriate amount of Fe3+ ions can control the photocatalytic activity under UV or visible radiation, so regulating the load of Fe seems crucial to obtain good results. In general, the photocatalytic activity of Fe-TiO2 rapidly raises with increasing Fe doping, reaches a maximum value, and decrease with further increases of Fe content as depicted in Fig. 12.4a (Zhang et al. 1998; Li et al. 2003; Yamashita et al. 2003; Adán et al. 2007; Cong et al. 2007; Ambrus et al. 2008; Tong et al. 2008; Li et al. 2009; Cai et al. 2010; Yalçın et al. 2010; Zhao et al. 2010; Wen et al. 2012; Kaur et al. 2016). The effect of Fe3+ doping act not only as an intermediate for photoinduced holes and electrons transfer but also as a shallow capturing site of photoinduced holes, which will efficiently separate the photoexcited charge carriers and increase their lifetime (Cai et al. 2010). Otherwise, the Fe3+ effects in TiO2 has been reported that only decrease photocatalytic activity as increasing Fe content under UV and visible light irradiation due to recombination of the generated charge carriers as depicted in Fig. 12.4b (Radford and Francis 1983; Bouras et al. 2007; Li et al. 2008; Fàbrega et al. 2010; Seabra et al. 2011). Another author reported an initial reduction of photocatalytic activity with increase in Fe content and then a rapid increase on photocatalytic activity followed by its reduction with further increasing Fe doping, as shown in Fig. 12.4c (Zhu et al. 2004). The Fe optimal value may depend on a variety of parameters, such as the synthesis method, thermal treatment, and on the synthesis precursor itself (Ambrus et al. 2008). For example, optimal Fe concentration reported were 0.034 at. % (Hemmati Borji et al. 2014), 0.2 at. % (Velázquez-Martínez et al. 2018), 0.2 at. % (Zhang et al. 1998), 0.5 at. % (Xing et al. 2010), 1 at. % (Li et al. 2003), 1.2 at. % (Ambrus et al. 2008), 1.8 at. % (Popa et al. 2008), 1.84 at. % (Goswami and Ganguli 2012), and 2 at. % (Melghit et al. 2009). Also, impair effects have been frequently observed when an excess of Fe3+ dopant is added because more Fe3+ in the TiO2 sample act as recombination sites and decrease the photocatalytic efficiency (Navıo et al. 1999; Bouras et al. 2007; Li

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Fig. 12.4 Tendency of the effect reported of Fe load in TiO2 photocatalytic activity which overload effect (a), impair effect (b), and bottom-up effect (c)

et al. 2008; Abazović et al. 2009; Cai et al. 2010; Fàbrega et al. 2010; MedinaRamírez et al. 2014). Fe3+ ions inhibit extended lifetime of charge carriers acting as electro-hole pair recombination center as stated in Eqs. (12.7) to (12.10) (Fàbrega et al. 2010). Recombination Fe2þ þ hVB þ ! Fe3þ

ð12:7Þ

Fe4þ þ eCB
! Fe3þ

ð12:8Þ

Fe4þ þ Fe2þ ! 2Fe3þ

ð12:9Þ

Fe4þ þ Ti3þ ! Fe3þ þ Ti4þ

ð12:10Þ

Recombination

Recombination

Recombination

12.5

Synthesis Method for Fe-Doped TiO2

Surface area, structure, and crystalline state strongly affect photocatalytic activity, so proper morphology control of nanostructured TiO2 materials play a significant role, which could be controlled by synthesis method. The specific surface area and surface-to-volume ratio facilitate reaction and interaction between the photocatalyst and the aqueous media (Abazović et al. 2009). Temperature treatment has a crucial effect on phases, for example, iron cations tend to segregate to the surface as

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calcination temperature increases (Fàbrega et al. 2010). Some studies pointed out that, even at trace amounts, doped Fe3+ can lead to the transition of anatase to rutile phase (Li et al. 2009). So far, several methods such as sol-gel, hydrothermal, wet impregnation, or ion-implantation has been developed for the preparation of Fe-TiO2, as mentioned in Table 12.1. The most frequently used technique for the synthesis of Fe-TiO2 nanoparticles (NP) is sol-gel with significant technological advantages such as the higher yield of isolated additive ions introduced and the significantly lower processing cost associated to the reduced number of processing steps involved (Fàbrega et al. 2010). Hemmati Borji et al. synthesized nanoparticles using the sol-gel method with two Fe/Ti atomic ratios, 0.006 and 0.034%, and materials’ photoactivity were tested by degradation of phenol (Hemmati Borji et al. 2014). The specific surface area (SBET) of Fe(III)-doped TiO2 aerogels obtained using different molar ratios of the reactants were three times greater than Degussa P25 powder (Popa et al. 2008). Impregnation technique seeks for the direct dopant introduction at the oxide surface. In this strategy, the additional charges are photogenerated already at the oxide surface where the catalytic reaction takes place. For example, impregnation of Fe3+ ions onto Aeroxide TiO2 P25 increased UV and visible activity for the calcined catalyst (Nahar et al. 2007). Nevertheless, impregnation usually results in accumulation of oxide clusters on the surface that could promote heterogeneous photo Fenton process on TiO2 surface (Zhao et al. 2010). Utilizing hydrothermal method after using acid pretreatment on Ti-bearing tailings is another way to obtain metal-doped TiO2. Sui et al. used this method to exploit the Ti-bearing tailings, obtaining nanoparticles with an average size 20 nm, having enhanced photocatalytic properties as examined in the degradation of rhodamine B using UV and visible radiation (Sui et al. 2018). In metal ion-implantation, Fe is accelerated enough to have the highest kinetic energy (150 keV) and to be implanted into the bulk of TiO2. Fe-implanted TiO2 exhibited an effective photocatalytic reactivity for the liquid-phase degradation of 2-Propanol diluted in water at 295 K under visible (λ > 450 nm) radiation (Yamashita et al. 2003). It was found that the photocatalytic reactivity under visible radiation increased as increasing the amount of implanted Fe ions. Another type of structure was used by Khang et al., being Fe-doped TiO2 and carbon nanotube nanohybrids synthesized by hydrolysis method (Khang et al. 2016). HR-TEM revealed that Fe-doped TiO2 nanoparticles (~8 nm) were attached to the carbon nanotube sidewalls. Unlike bare TiO2, the nanohybrid materials exhibited red-shift with increased absorbance in the visible range. The photocatalytic performance was tested on the degradation of methylene blue and methyl orange under visible irradiation, having the highest activity with 6% Fe-doped TiO2.

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Co- and Multi-Doping of Fe-Doped TiO2

Fe-TiO2 could be doped with one or more than one element (called co-dopants and multi-dopants, respectively). Co-doping of Fe-TiO2 has been studied with C, N, S, Au, and La, and multi co-doped Fe-TiO2 has been studied with C-N and C-B. In general, synergistic effect on photocatalytic activity was observed when mono and multi co-doped Fe-TiO2 used as dopants under visible radiation, so co-doped and multi-dopants of Fe-TiO2 have gained great attention. C co-doped Fe-TiO2 synthesized by microemulsion enhanced the photocatalytic activity with 2.01 wt% Fe under visible radiation (λ > 420 nm) with respect to 1.16, 3.08, 4.87, 6.68% Fe (Bhatu et al. 2018). N co-doped Fe-TiO2 under solar light enhanced photocatalytic degradation of phenol compared to Fe-TiO2 or N-doped TiO2 (Naik and Parida 2010). S co-doped Fe-TiO2 demonstrated dramatic enhanced photocatalytic activity under UV and visible radiation (Hamadanian et al. 2011). The deposition of Au on the Fe-TiO2 surface had synergistic effects and acted as a sink for photo-induced charge carriers that promotes interfacial charge-transfer processes under visible and UV radiation (Wu et al. 2009). La co-doped (2%) Fe-TiO2 increase photocatalytic degradation of phenol under visible radiation compared to undoped, 4% Fe and Aeroxide TiO2 P25 (Shi et al. 2011). N, C multi-doped Fe-TiO2 nanotubes (inner diameter 35–60 nm) were used for degradation of rhodamine B with best results obtained by the elemental composition 8.2 Fe%, 4.1 C%, and 3.4 N % (Momeni et al. 2016). N, B multi-doped Fe-TiO2 at 0.5 at. % Fe/Ti showed high photocatalytic activity under visible radiation and the photocatalytic degradation process increased with increasing Fe/Ti ratios (Xing et al. 2010).

12.7

Simulated Solar and Solar Stimulation of Fe-Doped TiO2

One of the main objectives of doping TiO2 is the reduced energy requirement for a photocatalytic process that could be achieved using artificial or solar visible radiation. Photocatalytic activity could be improved in different energy photons ranges: (1) between 387.5 nm < λ  400 nm; (2) using strictly visible radiation or photons with λ > 400 nm (Cong et al. 2007; Li et al. 2009; Cai et al. 2010); (3) simulated solar radiation (Fàbrega et al. 2010), or (4) direct solar visible radiation (Yamashita et al. 2003; Sonawane et al. 2004; Melghit et al. 2009; Naik and Parida 2010; Shao et al. 2016). Using visible radiation only, enhanced photocatalytic activity has been reported for 0.5% Fe-doped TiO2 obtained by grinding (Lezner et al. 2012), and 4% Fe-doped TiO2 synthesized by the sol-gel method (Shi et al. 2011). Shao et al. used titanium oxychloride, a rarely used precursor of Ti, by a versatile sol-gel approach in the absence of surfactants to investigate the photocatalytic degradation of methylene

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blue using natural sunlight (Shao et al. 2016). Photocatalytic degradation of aqueous congo red solution under sunlight was demonstrated using Fe-doped TiO2 nanorods with 2 at.% Fe (Melghit et al. 2009).

12.8

Emerging Contaminants Degradation Using Fe-Doped TiO2

Due to the relatively new introduction or detection of ECs, knowledge gaps exist on their fate, behaviors, and effects, as well as on treatment technologies for their efficient removal. Conventional wastewater treatment is not efficient for the elimination of ECs and their metabolites, so they are released into rivers or other natural streams (Gogoi et al. 2018). AOPs are efficient in the degradation of ECs, but less information related to their industrial development, process scaling-up, and research for the energetically green process is available (Rodriguez-Narvaez et al. 2017). As a consequence, many researchers have synthesized new materials to substitute the UV radiation for sunlight, visible light, LED light or laser irradiation for ECs degradation using Fe-TiO2. ECs underwent to degradation using Fe-doped TiO2 under UV and visible radiation were biocide (4-nitrophenol, 2,4-dichlorophenol, diazinon, mecoprop, carbedazim), pharmaceuticals (amoxicillin, carbamazepine, diclofenac, ibuprofen, sulfamethoxazole, streptomycin, tetracycline), and industrial chemicals (EDTA, 2-propanol) such as depicted in Fig. 12.2 and summarized in Table 12.1. All biocides were degraded using Fe-TiO2 nanoparticles under UV, meanwhile 2,4-dichlorophenol and carbedazim were also degraded under visible and Solar light, respectively. The best condition found to overcome 4-nitrophenol was using TiO2 nanoparticles impregnated with Fe with 1 wt. % coupled with H2O2 (Zhao et al. 2010) and 0.25 wt. % (Yalçın et al. 2010). 2,4-dichlorophenol was effectively degraded using Au co-doped Fe-TiO2 nanoparticles synthesized by sol-gel method under visible light (Wu et al. 2009). Sonophotocatalytic process were used to enhance degradation of Diazinon which optimal conditions were frequency of 37 kHz, power of 100 W, 0.4 g L
1 catalyst load and of 1.5% Fe content (Tabasideh et al. 2017). Although doping with Fe ions influenced on optical characteristics of TiO2, no increase of TiO2 photocatalytic activity after doping was observed for degradation of the herbicide mecoprop (RS-2-(4-chloro-otolyloxy) propionic acid, C10H11ClO3) (Abazović et al. 2009). Under Solar light, Aeroxide TiO2 P25 impregnated with 2 wt. % degraded carbedazim (Kaur et al. 2016). Although, the systems used to degrade pharmaceuticals reported in Table 12.1 were completely different (e.g., Co doped Fe-TiO2 nanoparticles, Fe-TiO2 immobilized on optical fibers and macroporous TiO2 oxides), all catalyst degraded pharmaceuticals under visible light. The best condition to degrade amoxicillin and streptomycin was pH 3.5 and pH 8 using N and Fe co-doped TiO2 under visible light (λ > 400) (Aba-Guevara et al. 2017). Carbamazepine, ibuprofen, and

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sulfamethoxazole were effectively removed using 3.4 wt. % Fe under visible light with respect to Aeroxide TiO2 P25 (Lin et al. 2017). 3D macroporous Fe-Ti-OxideM with 10% of Fe-doping achieved fairly good degradation and mineralization of tetracycline under visible radiation, suggesting also that transition metal ions are capable to introduce new energy levels into TiO2 Eg without changing its VB edge (Fan et al. 2014). Although the degradation of contaminants is relevant to the effectiveness of a water treatment process, the risk that harmful products could remain even after complete contaminant degradation has been reached. For this reason, mineralization of organic pollutants (e.g., the complete conversion of organic carbon into water and CO2) using TiO2 has attracted attention and been well documented (Herrmann 2005; Chong et al. 2010; Tong et al. 2012; Das 2014). One of the most significant species formed during photocatalytic process (as depicted Figs. 12.1 and 12.3) is free radicals that can participate in further reaction resulting in total mineralization of contaminants because of their nonselective nature and strong redox potential. Particularly, Fe-TiO2 has demonstrated the advantage of ECs mineralization under visible or solar light. For example, amoxicillin mineralization achieved was 41.5%, 34.7% for streptomycin after 240 min under visible radiation (Aba-Guevara et al. 2017), as well as 86% of tetracycline after 420 min under visible (Fan et al. 2014), and 4-nitrophenol after 60 min under UV (Zhao et al. 2010). Advance oxidation/Fe-TiO2 coupled systems have been proposed to overcome ECs. Bansal and Verma reported the scaling-up of the photocatalytic process coupled with a photo-Fenton process using Fe-TiO2 composite for degrading pentoxifylline, a vasodilator drug used for the treatment of atherosclerotic disease—using a non-concentrating solar reactor and demonstrated almost complete mineralization to nitrite, nitrate ammonium ions, and CO2 (Bansal and Verma 2018). Tabasideh et al. studied the sonophotocatalytic degradation of diazinon, a persistent organophosphorus pesticide used for insect control in household and agricultural applications—using Fe-doped TiO2, and showed Fe-doped TiO2 under UV and ultrasound irradiation had a synergetic effect on diazinon degradation (Tabasideh et al. 2017).

12.9

Conclusion

ECs in water bodies are definitively a major concern. The potential risks associated with their release and presence in surface, ground, and drinking water are not fully known, which could thread the environmental equilibrium or human health. As a result, there is a demanding need for new cost-effective methodologies able to degrade these compounds. Fe-TiO2 has been studied on the photocatalytic degradation of various dyes and ECs with significant findings. Fe-TiO2 has demonstrated to be synthesized using different methodologies, able to degrade a wide arrange of pollutants, and showing versatility to be coupled with other materials to enhance its photocatalytic capabilities.

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The catalytic role of iron during photoactivation of TiO2 still remains controversial, because of the different synthetic methods followed, the various post-synthesis processes, catalyst load and/or intensity of radiation used. This diversification on the processing methods and different experimental setups result in a large variety of iron chemical states, localization of active sites and/or target contaminant degradation routes that make reliable comparison unfeasible. However, the discrepancy among photocatalytic mechanisms of Fe-TiO2 does not diminish the relevant evidence that Fe3+ doping effectively improves degradation of organic compounds, in particular ECs, under UV and visible radiation. We suggest that the most viable mechanisms of Fe into TiO2 is the generation of one additional energy level below the CB, but it should further be theoretically and experimentally supported. Although research on the photocatalytic activity of Fe-TiO2 have been mainly focused on Fe dopant concentration, preparation methods, and thermal and reductive treatment. Further researches should take account of the spectrum and intensity of the irradiation source, the catalyst’ surface charge and ionic strength to understand the whole Fe-TiO2 photocatalytic mechanism. Some research is currently focusing on coupling Fe-TiO2 with other materials such as carbon nanotubes and graphene or other AOPs like photo-Fenton or sonophotocatalysis, in order to enhance the enhance the capability of photocatalytic processes in the removal of ECs in aqueous phase. On the other hand, other investigations aim for co-doping or multi-doping agents, which improve photocatalysis in some way. These studies should be continued in order to benefit Fe-TiO2 performance as photocatalyst for pollutant removal, as the results shown so far seem promising.

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Chapter 13

Oxide Nanomaterials for Efficient Water Treatment Alagappan Subramaniyan

Contents 13.1

Introduction: Some Interesting Facts on Depleting Water and Water Treatment Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.2 Nanomaterials for Water Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.3 Aluminum Oxide (Al2O3) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.4 Zinc Oxide (ZnO) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.5 Titanium Oxide (TiO2) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.6 Iron Oxide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.7 Cerium Oxide (CeO2) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.8 Magnesium Oxide (MgO) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.9 Graphene Oxide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.10 Copper Oxide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.11 Oxide Nanomaterial Versus Carbon . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.12 Challenges and Issues in NMWT . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.13 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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There is an acute shortage of drinking water with close to 42% of the total world population lacking access to clean and safe drinking water since 2005. According to WHO/UNICEF recent reports nearly 663 million people around the world do not

A. Subramaniyan (*) Department of Physics, Thiagarajar College of Engineering, Madurai, Tamil Nadu, India e-mail: [email protected] © Springer Nature Switzerland AG 2019 R. Prasad, K. Thirugnanasambandham (eds.), Advanced Research in Nanosciences for Water Technology, Nanotechnology in the Life Sciences, https://doi.org/10.1007/978-3-030-02381-2_13

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have access to safe drinking water (Prathna et al. 2018). It is important to note that an over a billion of the estimated 6.2 billion people relying on improved water sources continue to use water that is unsafe (Bartram et al. 2014). It is reported that on an average 842,000 deaths occur annually due to water-related problems including insufficient water. Nearly 48% occur in children below 5 years (Pruss-Ustun et al. 2014). The growing population global warming heavy industrial usage along with poor measures to preserve rain water also contribute to the water crisis. In the presence of adequate ground water a preferred alternative solution is water treatment. The quality of ground water varies across different geographical locations. The presence of fluoride more than 1.5 mg/L makes the water unsafe for drinking (Teutli-Sequeira et al. 2014). Available ground water in Asia and Africa are more than 25 mg/L (Mondal 2015). Hence defluoridation is an essential process in making the ground water safe for drinking. Defluoridation can be achieved by a variety of ways like adsorption ion exchange electrocoagulation nano-filtration and membrane separation (Suja and Pandit 2010). Adsorption is the most popular technique till now for water defluoridation as it is an economical route for generating high quality of treated effluent and also offers flexibility with respect to design and scaling. Metals, metal oxides, carbon, and polymers are used as adsorbents. The efficiency of adsorbents in defluoridation depends on its area and reactivity. This could contribute to high fluoride removal from ground water and also speed up the water treatment process. The advent of nanotechnology has provided facilities for production of nanoadsorbents with larger surface area enhanced reactivity and self-assembly which can give efficient and quick water treatment methods. There are a variety of water treatment methods available which function on different principles. Some methods are focused on microbial reduction and others are concentrated on removing hazardous elements. The water treatment methods have been in accordance with latest technological developments. Probably Stone Age concentrated on sand filtration and today’s filtrations are drifting towards NMWT. The sand filters work on growth of biofilm surrounding the sand particles that are rich in bacterial population and hence improve the filtration power of the media (Johnson 1914). Some other conventionally used techniques include chemical oxidation chemical precipitation coagulation and ion exchange. The most common chemical oxidant used is chlorine which provides an effective and robust barrier to pathogens (Fawell and Nieuwhuijsen 2003). Chemical precipitation works by the addition of counter ions to reduce the solubility of ionic contaminants. However adsorption is the most feasible and economically attractive treatment method for removal of arsenic fluoride and heavy metals from water (Petrusevski et al. 2008). The past 10 years have witnessed increased interest on the utilization of nanoparticles as adsorbents in water treatment. Nanotechnology has been proved as the best way to treat permanent and emerging contaminants (Brame et al. 2011). Nanomaterial based adsorbents offer attractive alternatives to conventional adsorbents (bulk adsorbents) because of their higher aspect ratio enhanced reactivity which could translate to higher adsorption capacity (Das et al. 2014).

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Nanomaterials for Water Treatment

Nanotechnologies for water treatment processes are based on nanoadsorbents, photocatalysts, microbial disinfectants, and nano-membranes. High and effective implementation of nanotechnology for water treatment demands overcoming the high cost of the nanomaterials. This could be achieved by recycling along with regeneration. Nanoadsorbents offer additional possibilities at household level in powder form, thin film coated on a substrate, and in a membrane filter. Nanoparticles are small particles with big impact on technology with respect to their device size and efficiency. The smaller size of the particles provides the advantage of constructing compact systems. Reports have shown that nanoparticles can be tailored to target multiple contaminants at the same time (Qiao et al. 2014), thereby possibly reducing the treatment cost. However, there are some emerging concerns over the safe disposal of nanomaterials and their potential risk to consumer health and ecosystem due to their toxic nature. A second major issue is the relatively higher cost compared to bulk counterpart. Finally, the operation and management of the treatment system is also a major limitations for the application of nanotechnology based water treatment. The development and use of novel nanomaterials for environmental protection and water treatment has received elaborate attention in recent years (Zhong et al. 2007; Prathna et al. 2011). An extensive search on the database (Google Scholar) for nanoparticle application for drinking water treatment shows approximately 47,200 entries with approximately 7000 entries (15%) in 2017 alone. This highlights the increased interest in investigating application of nanoparticles for water treatment. Different types of nanoparticles have been used for adsorptive removal studies. Some of them include iron based nanoparticles for arsenic removal, carbon and aluminum based nanomaterials for fluoride removal, etc. (Jadhav et al. 2015). Carbon nanotubes have been extensively employed in combination with other nanomaterials for the removal of fluoride from water (Li et al. 2001). Modified multi-walled carbon nanotubes have been used extensively for the removal of a number of heavy metals from water (Li et al. 2011). Graphene based nanomaterials have been used to remediate Pb (II) ions from water (Zhao et al. 2005). Other nanomaterials such as alumina are used extensively for the removal of fluoride while magnetic nanomaterials are used for the removal of arsenic from water (Jadhav et al. 2015). Studies have been carried out to effectively remove organic contaminants using semiconductor ZnS and TiO2 nanomaterials (Hu et al. 2005) while zerovalent iron and bimetallic Fe nanoparticles have been used to remediate both organic and inorganic contaminants. Furthermore significant work is being done on the development of suitable nano-adsorbent for the simultaneous removal of multiple contaminants. 3D flower iron oxide nanostructures have been developed for the simultaneous removal of As, Cr, and azodye Orange II (Zhong et al. 2006). Ceria 3D flower-like nanostructures were used for the effective removal of As and Cr from contaminated waters (Zhong et al. 2007). Simultaneous removal of the most toxic contaminants in drinking water, namely arsenic and fluoride, has been studied by Li et al. (2011) using mesoporous alumina. Nano-enabled filtration membranes

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Nanotechnology is being utilized in developing innovative polymeric and ceramic membranes to improve the performance of membrane filtration systems (Pendergast and Hoek 2011). In theory incorporation of nanoparticles in membranes provides fouling resistance as well as additional benefits of disinfection and contaminant degradation depending on the nanomaterial used (Li et al. 2008).

13.3

Aluminum Oxide (Al2O3)

Al2O3 has gained importance as a catalyst component/absorbent and ceramic material in various industrial processes. There are seven Al2O3 polymorphs although only four, called α, δ, θ, and γ, are typically involved in most of the industrial processes. Theoretical studies of (Al2O3)n (n
15) yielded certain structural, electronic, and chemical behavior against adsorbates. The γ-Al2O3 is nanostructured phase commonly obtained by most synthetic methods but also the α-Al2O3 polymorph is synthesized having high surface area. γ-Al2O3 has a lower surface energy and becomes energetically stable at size below a point close to 10 nm.

13.4

Zinc Oxide (ZnO)

Zinc oxide is an organic semiconductor white solid that is insoluble in water. Zinc Oxide is wide-band gap (3.3 eV) semiconductor of II–IV semiconductor group and has large exciton binding energy of 60 meV. ZnO has several favorable properties including good transparency, high electron mobility, and strong room-temperature luminescence. These properties are used in emerging applications for transparent electrodes, in liquid crystal displays, in energy-saving or heat-protecting windows, and in electronics as thin-film transistors and light-emitting diodes. It is widely used as an additive in numerous materials and products including rubbers, plastics, ceramics, glass, cement, lubricants, paints, ointments, adhesives, sealants, pigments, foods, batteries, ferrites, fire retardants, and first-aid tapes. ZnO is a unique material that exhibits semiconducting piezoelectric and pyroelectric multiple properties. Using a solid-vapor phase thermal sublimation technique, nanocombs, nanorings, nanohelixes/nanosprings, nanobows, nanobelts, nanowires, and nanocages of ZnO have been synthesized under specific growth conditions. The nanostructures could have novel applications in optoelectronics, sensors, transducers, and biomedical due to its nontoxic nature. Zinc oxide is considered to be the best semiconductor at nano-dimensions. The dependence of properties on the size of particles has led to many interesting applications especially by tuning the band gap of semiconductors. One of the main advantages of ZnO is its high doping ability. The band gap of ZnO can further be tuned to 3–4 eV by its alloying with magnesium oxide or cadmium oxide. ZnO has n-type character even in the absence of intentional doping.

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13.5

291

Titanium Oxide (TiO2)

TiO2 is the naturally occurring oxide of titanium. When used as a pigment it is called titanium white. It has a wide range of applications from paint to sunscreen to food coloring. Advantages of TiO2: • • • • • • •

Easily fabricated. Causes no pollution. Avoids the diffusion between the phase change material and the bottom electrode. Non-toxicity. Can be used as a catalyst. Low price. It is used as sunscreen and UV blocking pigments in the industry.

Titanium dioxide (TiO2) is one of the most prominent oxide materials for performing various kinds of industrial applications related to catalysis. It is used as a white pigment in paintings, as part of photovoltaic devices or electrochromic devices, sensors, as a food additive in cosmetics, and as a potential tool in cancer treatment. In TiO2 materials the so-called “quantum confinement” or “quantum-size effect” is restricted to very low sizes, below 10 nm, due to their rather low exciton Bohr radii. This would mean that a significant part of the potential novel chemical or physical applications needs to be carefully explored in the range of a few nanometers. TiO2 occurs in nature in three different polymorphs which, in order of abundance, are rutile, anatase, and brookite. Additional synthetic phases and several high pressure polymorphs have been also reported. At high temperatures TiO2 can be easily reduced and this decisively influences conductivity. The grain boundaries of TiO2 can be altered to strongly influence its electrical properties. Defects can also be introduced in TiO2 to tune its properties.

13.6

Iron Oxide

Fe and O form a number of phases, e.g., FeO (wustite), Fe3O4 (magnetite), α- Fe2O3 (hematite), γ-Fe2O3 (maghemite), and ε-Fe2O3.309. The latter two phases are synthetic while remaining oxides occur in nature. The magnetic properties of the nano Fe oxides have been extensively studied, in particular the enhancing magnetic recording properties of magnetite.

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Cerium Oxide (CeO2)

Cerium(IV) oxide, also known as ceric oxide, ceric dioxide, ceria, cerium oxide, or cerium dioxide, is an oxide of the rare-earth metal cerium. It is a pale yellow-white powder with the chemical formula CeO2. It is an important commercial product and an intermediate in the purification of the element from the ores. At the wastewater treatment plant about 95% of CeO2 NPs undergo biosolid sludge accumulation in wastewater treatment plants. CeO2 discharge into the aquatic environment is fairly limited as compared to accumulation levels in the terrestrial environment including landfills.

13.8

Magnesium Oxide (MgO)

Magnesium oxide or magnesia is a white hygroscopic solid mineral that occurs naturally as periclase and is a source of magnesiumIt has an empirical formula of MgO and consists of a lattice of Mg2+ ions and O2 ions held together by ionic bonding. Magnesium hydroxide forms in the presence of water (MgO + H2O ! Mg (OH)2) but it can be reversed by heating it to separate moisture. MgO finds application in refractory materials and in medical industry as a relief for heart burn and sour stomach. One of the major applications of MgO is its use in the wastewater treatment industry. This is due to its efficiency and cost-effectiveness compared to other chemicals. An advantage of using MgO is in the effective removal of phosphorous and its minimal environmental impact due to absence of harmful by-product during its application. When MgO is applied to the wastewater it then reacts with the solution to become magnesium hydroxide Mg(OH)2. The natural characteristics of Mg(OH)2 suspension offers the best performance in wastewater treatment. The benefits are listed below. • Due to its high alkalinity, it is easy to control the pH without the risk of exceeding the Maximum Allowed Limitation of 9. • By applying the suspension in the wastewater, the pH increases quickly while maintaining the optimum bacterial growth conditions for biological treatment. • In the event that heavy metals are present in the wastewater, the suspension will be able to allow them to precipitate effectively to be extracted from the water. • The suspension is also able to remove phosphorous effectively compared to sodium hydroxide and calcium hydroxide.

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13.9

293

Graphene Oxide

Graphite is a 3D-carbon based material which can be thought of as being made from many layers of graphene. Graphite oxide is slightly different from graphite. When strong oxidizing agents are used to oxidize graphite, oxygenated functionalities are introduced in the structure of the graphite which makes the material hydrophilic and also expands the layer separation. This feature makes it possible to exfoliate the graphite oxide in water through sonication and this eventually produces monolayer or few-layer oxygen-functionalized graphene called graphene oxide or GO. The number of layers is the major difference between graphene oxide and graphite oxide. Although graphite oxide is a multilayer system, monolayer flakes and few-layer flakes can be found in a graphene oxide dispersion. Graphite oxide is a compound of carbon oxygen and hydrogen in variable ratios obtained by treating graphite with strong oxidizers. The maximally oxidized bulk product is a yellow solid with C:O ratio between 2.1 and 2.9 that retains the layer structure of graphite but with a much larger and irregular spacing. The graphene oxide membrane, the modified graphene oxide membrane, and the oxidized grapheme hybrid membrane have the advantages of simple preparation process and good separation performance; it has great potential in the field of water treatment. In the view of the ion transport, principle in graphene membrane is not yet clear and it needs a deep understanding of the mechanism when being applied to desalination sodium filtration pervaporation and other fields. At the same time, for the new separation membrane of enriched graphene group how to improve the strength of the separation membrane and put it into practical application is an urgent problem to be solved (Yong chen Liu 2017 IOP Conf. Ser).

13.10

Copper Oxide

Copper oxide is a compound from the two elements, copper and oxygen. Copper oxide may refer to: • • • •

Copper(I) oxide (cuprous oxide Cu2O). Copper(II) oxide (cupric oxide CuO). Copper peroxide (CuO2). Copper(III) oxide (Cu2O3). One of the most important applications of copper and copper compounds is bacteria disinfectant due its versatility low cost essential for humans at low levels and biocide activity properties. Copper nanoparticles have been used to purify Godavari river water.

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13.11

A. Subramaniyan

Oxide Nanomaterial Versus Carbon

Lab-scale studies have been carried out to determine the potential of carbon nanotube membranes in water treatment. These nanotube membranes have been studied to be effective against bacteria, virus, turbidity, and also towards the removal of organic contaminants. The membranes were identified as promising for desalination and as an alternative to reverse osmosis membranes since these filters require less energy due to their high permeability. Carbon nanotube membranes in addition are expected to be more durable and easier to clean and reuse (Srivastava et al. 2004). The cost of carbon nanotube adsorbents could be reduced by large-scale manufacturing. Porous nanoparticles are introduced to membranes to increase their efficiency. A technique called direct contact membrane distillation is developed with buckyball paper. This technique is an alternative to reverse osmosis and other desalination techniques particularly when the concentration of solutes is high. The Buckypaper is used as a highly hydrophobic membrane to separate a feed of hot sea or brackish water from a permeate of cold fresh water. While liquid cannot cross the air gap formed by the membrane, water vapor is able to pass through the pores from the hot feed to the cold permeate driven by the difference in partial vapor pressure. This vapor then condenses on the permeate side creating fresh water. The inherent hydrophobicity of the nanotubes and high Buckypaper porosity of nearly 90% lend them to water vapor permeabilities of up to 3.3  1012 kg/m sPa on a smallscale rig. However cracking of the Buckypapers with time is a problem as salt water can penetrate into the relatively large cracks and breach the Buckypaper membrane. This leads to a gradual reduction in permeate quality over time.

13.12

Challenges and Issues in NMWT

Paper based filters coated with nanoparticles are easy to produce and distribute to remote locations and do not require energy inputs for water treatment. A Point of Use (POU) water treatment system consisting of bactericidal paper impregnated with silver nanoparticles to deactivate pathogenic bacteria was developed by Dankovich and Gray. The silver loss from the nanoparticle impregnated sheets was minimal and had the potential to be an effective emergency water treatment system. The same group has also developed POU paper filters with copper nanoparticles effective against E. coli with a log reduction of 8.8 when a concentration of 65 mg Cu/g was used. The authors reported that the amount of copper in each paper amounts to less than a cent and can be potentially used in emergency response and disaster relief rural household filters and as backpacking filters. However studies are still ongoing regarding the volume of water that can be treated using a single filter paper (Dankovich and Smith 2014). Systematic studies need to be carried out testing the efficacy of commercial nanotechnology based filters in the market. In a related study, Mwabi et al. (2011) evaluated the design and working of four household water

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treatment devices, namely a bios and filter, a bucket filter, ceramic candle filter, and silver impregnated porous pot filter. The performance was evaluated in terms of flow rate physicochemical contaminant and microbial contaminant removals. The silver coated ceramic pot demonstrated the highest bacterial reduction compared to other filters. However a cost comparison study was not performed. A challenge associated with the use of nano-adsorbents is the likely release of metal ions from the nanoparticles into the treated water. For example, Stumm and Wollast observed the dissolution of Al3+ ions from alumina nanoparticles to be a two-step process; the first step involved the interaction of the nanoparticle surface with the matrix followed by detachment of metal ions from Al-O complex. Aluminum ions have been observed to cause neurological and gastro intestinal toxicity (Krewski et al. 2007) though the WHO has not established a firm limit on aluminum ion toxicity due to inadequate data. Similarly, WHO states that the known risk of silver ingestion is a condition called argyria which discolors skin and hair. To prevent this WHO recommends that silver concentrations of 0.1 mg/L can be tolerated for 70 years without any health risk. Hence studies should also focus on the potential release of metal ions from the nanoparticles when used for water treatment. The study conducted by Mellor et al. (2015) observed that water treated by nano-silver coated ceramic pots had silver concentrations less than the WHO permissible limits over a period of 12 months as compared to using conventional treatment involving chlorination. The WHO limit for chlorine is 5 mg/L and elevated levels can increase the risk of bladder cancer. More studies in this direction can be helpful in determining the long-term performance of filters and can be used to provide recommendations to small communities for household level treatment of drinking water.

13.13

Conclusions

Though several water treatment methods are being practiced since Stone Age, the vast research in easy synthesis of oxide nanopowders has driven the technology towards NMWT. Research still continues in improving the efficiency of water treatment with respect to composition and morphology of oxide nonmaterial. This chapter describes the properties and efficiency of water treatment of oxide nonmaterials like Alumina, Zinc oxide, Titanium dioxide, Silica, Magnesium oxide, Cerium oxide, Graphene oxide, and Copper oxide. Easy synthesis of the oxide nanopowders in a large scale would reduce the cost of the nanodevice. Nanotechnology has contributed to small and efficient water treatment devices.

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References Bartram J, Brocklehurst C, Fisher MB, Luyendijk R, Hossain R, Wardlaw T, Gordon B (2014) Global monitoring of water supply and sanitation: history methods and future challenges. Int J Environ Res Public Health 11(8):8137–8165 Brame J, Li Q, Alvarez PJJ (2011) Nanotechnology enabled water treatment and reuse: emerging opportunities and challenges for developing countries. Trends Food Sci Technol 22:618–624 Dankovich TA, Smith JA (2014) Incorporation of copper nanoparticles into paper for point of use water purification. Water Res 63:245–251 Das R, Ali Md E, Hamid SBA, Ramakrishna S, Chowdhury ZZ (2014) Carbon nanotube membranes for water purification: a bright future in water desalination. Desalina 336:97–109 Fawell J, Nieuwhuijsen MJ (2003) Contaminants in drinking water. Br Med Bull 68:199–208 Hu J, Chen G, Lo IM (2005) Removal and recovery of Cr(VI) from wastewater by maghemite nanoparticles. Water Res 39:4528–4536 Jadhav SV, Bringas E, Yadav GD, Rathod VK, Ortiz I, Marathe KV (2015) Arsenic and fluoride contaminated groundwaters: a review of current technologies for contaminants removal. J Environ Manag 162:306–325 Johnson GA (1914) Present day water filtration practice. J Am Water Work Assoc 1:31–80 Krewski D, Yokel RA, Nieboer E, Borchelt D, Cohen J, Harry J, Kacew S, Lindsay J, Mahfouz AM, Rondeau V (2007) Human health risk assessment for aluminium oxide and aluminium hydroxide. J Toxicol Environ Health B Crit Rev 10:1–69 Li Q, Mahendra S, Lyon DY, Brunet L, Liga MV, Li D, Alvarez PJJ (2008) Antimicrobial nanomaterials for water disinfection and microbial control: potential applications and implications. Water Res 42:4591–4602 Li YH, Wang S, Cao A, Zhao D, Zhang X, Xu C, Luan Z, Ruan D, Liang J, Wu D, Wei B (2001) Adsorption of fluoride from water by amorphous alumina supported on carbon nanotubes. Chem Phys Lett 350:412–416 Li C, Zhang Y, Wang X, Zhao J, Chen W (2011) Removal and recovery of lead (II) ions from contaminated licorice extracts using oxidized multi-walled carbon nanotubes. J Nanosci Nanotech 11:9731–9736 Mellor JE, Kallman E, Craver VO, Smith JA (2015) Comparison of three household water treatment technologies in San Mateo Ixtatan Guatemala. J Environ Eng 141(5):04014085. https://doi.org/ 10.1061/(ASCE)EE.1943-7870.0000914 Mondal SG (2015) A review on adsorbents used for defluoridation of drinking water. Rev Environ Sci Biotechnol 14(2):195–210 Mwabi JK, Adeyemo FE, Mahlangu TO, Mamba BB, Brouckaert BM, Swartz CD, Offringa G, Mpenyana-Monyatsi L, Momba MNB (2011) Household water treatment systems: a solution to the production of safe drinking water by the low-income communities of Southern Africa. Phys Chem Earth 36:1120–1128 Pendergast MTM, Hoek EMV (2011) A review of water treatment membrane nanotechnologies. Energy Environ Sci 4:1946–1971 Petrusevski B, Sharma S, Van der Meer WG, Kruis F, Khan M, Barua M, Schippers JC (2008) Four years of development and field-testing of IHE arsenic removal family filter in rural Bangladesh. Water Sci Tech 58(1):53–58 Prathna TC, Chandrasekaran N, Mukherjee A (2011) Studies on aggregation behaviour of silver nanoparticles in aqueous matrices: effect of surface functionalization and matrix composition. Colloids Surf A Physicochem Eng Aspects 390:216–224 Prathna TC, Sharma S, Kennedy M (2018) Nanoparticles in household level water treatment: an overview. Sep Purif Technol 199:260. https://doi.org/10.1016/j.seppur.2018.01.061 Pruss-Ustun A, Bartram J, Clasen T, Colford JM, Cumming O, Curtis V, Bonjour S, De Dangour AD, France J, Fewtrell L, Freeman MC (2014) Burden of disease from inadequate water sanitation and hygiene in low and middle income settings: a retrospective analysis of data from 145 countries. Trop Med Int Health 19:894–905

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Qiao J, Cui Z, Sun Y, Hu Q Guan X (2014) Simultaneous removal of arsenate and fluoride from water by Al-Fe hydroxides. Front Environ Sci Eng 8:169–179 Srivastava A, Srivastava ON, Talapatra S, Vajtai R, Ajayan PM (2004) Carbon nanotube filters. Nat Mater 3:610–614 Suja GP, Pandit AB, Process modeling for estimation of colloidal aluminium in alum treated water during defluoridation. In: International Conference on Environmental Engineering and Applications (ICEEA 2010). p 103–106 Teutli-Sequeira A, Riose MM, Linares H (2014) Comparison of aluminum modified natural material in the removal of fluoride ions. J Colloid Interface Sci 418:254–260 Zhao G, Ren X, Gao X, Tan X, Li J, Chen C, Huang Y, Wang X (2005) Removal of Pb(II) ions from aqueous solutions on few-layered graphene oxide nano sheets. Dalton Trans 40:10945–10952 Zhong LS, Hu JS, Cao AM, Liu Q, Song WG, Wan LJ (2007) 3D flower like ceria micro/ nanocomposite structure and its application for water treatment and CO removal. Chem Mater 19:1648–1655 Zhong LS, Hu JS, Liang HP, Cao AM, Song WG, Wan LJ (2006) Self- assembled 3D flowerlike Iron oxide nanostructures and their application in water treatment. Adv Mater 18:2426–2431

Chapter 14

Nanotechnology for Oil-Water Separation Prakash M. Gore, Anukrishna Purushothaman, Minoo Naebe, Xungai Wang, and Balasubramanian Kandasubramanian

Contents 14.1 14.2 14.3 14.4

14.5

14.6

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Superwettable Janus Materials . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Superwettable Janus Membrane Based on Polycarbonate and Nano-Copper Phthalocyanine . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Results and Discussion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.4.1 Morphological Analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.4.2 Separation Performance . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.4.3 Capillary Pressure Study . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.4.4 Breakthrough Pressure Analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Superwettable Janus Fabric Based on Polyester and Nano-PTFE Particles . . . . . . . . . . . . 14.5.1 Results and Discussion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.5.2 Wetting Behavior of Janus Fabric . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.5.3 Absorption Study . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.5.4 Separation Analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.5.5 Permeation Flux Analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.5.6 Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Superwettable Sponges and Foams for Effective Oil-Water Separation . . . . . . . . . . . . . . . .

300 302 307 309 310 314 314 316 318 320 321 324 326 327 328 328

P. M. Gore Nano Surface Texturing Lab, Department of Metallurgical & Materials Engineering, DIAT (DU), Ministry of Defence, Pune, Girinagar, India Institute for Frontier Materials, Deakin University, Geelong, VIC, Australia A. Purushothaman Centre for Biopolymer Science and Technology, Central Institute of Plastics Engineering and Technology, Eloor, Udyogmandal, Kochi, India M. Naebe · X. Wang Institute for Frontier Materials, Deakin University, Geelong, VIC, Australia B. Kandasubramanian (*) Nano Surface Texturing Lab, Department of Metallurgical & Materials Engineering, DIAT (DU), Ministry of Defence, Pune, Girinagar, India © Springer Nature Switzerland AG 2019 R. Prasad, K. Thirugnanasambandham (eds.), Advanced Research in Nanosciences for Water Technology, Nanotechnology in the Life Sciences, https://doi.org/10.1007/978-3-030-02381-2_14

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14.7 Conclusion and Future Direction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 333 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 334

14.1

Introduction

The increased consumption of oils and industrial solvents in recent decades for fulfilling the ever-increasing energy needs has accelerated the development of associated industries and world economies (Fingas 2012, 2015). The massive transportation of these oils and solvents to various industrial sectors is mainly accomplished via sea-based roots, which involves frequent oil-spills and accidents, e.g., Gulf of Mexico in 2011 (Fingas 2015; Hansen 2016). These oil seepages cause enormous amount of oils, i.e., ~1.7 to 8.8 million tons every year, being liberated in oceans, which accommodate highly sensitive marine ecosystem (Gore et al. 2016a; Gupta and Kandasubramanian 2017a). These oil-spills traumatize the marine organisms and ecosystems, and further cause chemical toxicity and oxygen and light deficiency, thus threatening their survival (Fingas 2012, 2015). Researchers have explored various chemical, biological, and physical techniques for accomplishing the effective oil-water separation for countering the oil-seepages in such situations. Recently, researchers have been exploring nanomaterials for efficient oil/solventwater separation, as they render highly active surface area, improved functionality with ability to tailor the properties, and nano-scale dispersion (Xue et al. 2014; Arora and Balasubramanian 2014; Padaki et al. 2015; Yadav et al. 2016; Breitwieser et al. 2018; Simon et al. 2018; Walker et al. 2018; Padhi et al. 2018). The literature analysis done via Web of Science reveals nearly 868 articles being published in last 10 years, i.e., from year 2008 to 2018, on oil-water separation using nanomaterials, stating their rapidly increasing importance. Further, the literature analysis shows that the average number of research papers published on oil-water separation using nanomaterials have increased at an average rate of 145% (Fig. 14.1a). Considering, the abovementioned techniques, the physical absorption method involves utilization of foams (Mishra and Balasubramanian 2014; Arora and Balasubramanian 2014), activated carbon (Ma et al. 2016), textiles substrates (Gore et al. 2016b), metal fine mesh (Matsubayashi et al. 2017), and fibers (Xue et al. 2014), which have been widely explored by the researchers.. The recently developed techniques such as electrospinning (Cheng et al. 2017; Gore and Kandasubramanian 2018), layer-by-layer assembly, selective oil extraction, and air-flotation have also gained wide attention by researchers for effective oil-water separation, due to their facile fabrication (Xue et al. 2014; Wang et al. 2015a; Ma et al. 2016). In order to achieve the efficient oil-water separation, the materials exhibiting high superwettability (Sahoo and Kandasubramanian 2014a; Sahoo and Balasubramanian 2014) absorption performance towards oil and/or water is required, which is mainly dependent on their surface characteristics such as superhydrophobicity/oleophobicity (water/oil contact angle >150
), superhydrophilicity/oleophilicity (water/oil contact angle ~ 0
), and hierarchical surface roughness at nano/-micro scale, thus making

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Fig. 14.1 (a) Scopus literature analysis, (b) Janus Effect, and its utilization in Oil-Water separation (Saini and Kandasubramanian 2018). (Reprinted with permission. Copyright 2018, Taylor and Francis)

them effective superwettable material (Xue et al. 2014). Various plants and animal surfaces manifest the superhydrophobicity with water contact angle (WCA) greater than 150
and sliding angle less than 10
which inspires the bionic creations having unique wettability structures, such as Nelumbo nucifera (Lotus) leaf surface (Sahoo and Kandasubramanian 2014b; Sahoo and Balasubramanian 2014; Sahoo et al. 2015). The liquids with lesser surface tension than the Lotus leaf show the immediate spreading and penetration of liquid. The surfaces with surface tension less than water (72.8 mN m1), greater than oil/organic solvents ( 90
, then the negative pressure gets activated, where it restricts the entry of liquid in the pores. In present study, the petroleum oils instantly wet the Janus microfiber membrane with even spreading; thus, the contact angle of the oils with pore wall can be presumed as 0
; hence, considering this, theoretical capillary pressures of petroleum oils for Janus microfiber membrane have been calculated, and are tabulated in Table 14.1. The capillary pressures of water could not be calculated for Pristine PC and Janus microfibers, due to limitation in analyzing the water contact angle with pore wall of inter-connected microfibers. The approximate average pore radius of interconnected microfibers of Pristine PC and Janus microfibers was found to be 20 μm and 6 μm, respectively (determined via FE-SEM analysis).

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Table 14.1 Capillary pressures for various petroleum oils and organic solvents Petroleum oils Diesel Petrol

Surface tension (mN/m) 29.5 29

Capillary pressure (N/m2) Pristine PC microfibers 2950 2900

Janus microfibers 9833 9670

Table 14.1 shows that the theoretically calculated capillary pressures for Janus microfibers are lower, i.e., 9833 and 9670 N/m2 for Diesel and Petrol, respectively, which implies their maximum affinity towards oils, thereby facilitating the higher permeation of the oils through the Janus microfiber membrane. The theoretical capillary pressures for Pristine PC microfibers are found to be 2950 and 2900 N/m2. The noticeable differences observed in the theoretically calculated capillary pressures is attributed to their pores radius, i.e., 20 μm for PC microfibers and 6 μm for Janus microfibers, which is related to the distance between the interconnected microfibers. Further, Table 14.1 also reveals that the capillary pressures decrease as the surface tension of oil decreases; this is attributed to the increased viscosity of oils, i.e., Diesel, having high surface tensions, which causes adhesion with Janus microfibers (Wang et al. 2015d; Yu et al. 2017b; Gore and Kandasubramanian 2018). This analysis is also reinforced by the lower permeation flux of Diesel, i.e., 6903 Lm2h1 (for 3 wt% loaded CuPc), and the higher flux petrol, i.e., 2829 Lm2h1 (for 3 wt% of CuPc).

14.4.4 Breakthrough Pressure Analysis In order to prevent the failure of the Janus microfiber membrane due to cross-flow of water (which might occur under external force), hence, its breakthrough pressure analysis is important for determining its service performance. Therefore, the breakthrough pressure can be calculated theoretically for Janus microfiber membrane by presuming the microfiber membrane as interconnected web with predominant cylindrically shaped structures (microfibers), via equation (Choi et al. 2009; Chhatre et al. 2010; Tian et al. 2012b; Gupta and Kandasubramanian 2017): lcap ð1  cos θÞ 
 Pref PBreakthrough ¼
R D∗ D∗ cylinder  1 cylinder  1 þ 2 sin θ

ð14:6Þ

where θ ¼ contact angle between oil and microfiber membrane surface, D*cylinder ¼ estimation of porosity of surface porosity by presuming spacing and structure manifesting in the pores (dimensionless quantity), given by following equation (Choi et al. 2009; Chhatre et al. 2010; Tian et al. 2012b; Gupta and Kandasubramanian 2017):

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D∗ cylinder ¼

RþD R

ð14:7Þ

where R ¼ radius of cylindrically shaped structures, i.e., microfibers, and 2D ¼ inter-cylinder spaces, i.e., inter-microfiber spaces. Pref is a least possible differential pressure for a liquid, i.e., oil, given by following equation (Chhatre et al. 2010): Pref ¼

2γ lcap

ð14:8Þ

where γ ¼ surface tension, lcap ¼ capillary length of liquid, calculated via following equation (Chhatre et al. 2010): lcap ¼

rffiffiffiffiffi γ ρg

ð14:9Þ

where ρ ¼ liquid density, g ¼ acceleration due to gravity. The Janus microfiber used in the present study revealed an approximate average diameter of 2.1 μm, and an average inter-cylinder (inter-microfiber) spacing of 6 μm (analyzed via FE-SEM and ImageJ application). The theoretically calculated breakthrough pressure for cross-flow of water through a surface of Janus microfiber (for WCA ~ 158
) membrane is found as 1.260 kPa, and for Pristine PC microfibers (for WCA ~ 113
) it is found to be 0.01978 kPa. The petroleum oils instantly get absorbed and transported via Janus microfiber membrane, which is also attributed to their varying surface tension, i.e., (γwater > γJanus microfiber membrane > γpetroleum  oil), and the water exhibiting high surface tension rest on membrane surface (Gupta and Kandasubramanian 2017). Summary. In summary, superhydrophobic copper phthalocyanine pigmented nanofibers of polycarbonate were fabricated via electrospinning technique. The entrenched nanofibers exhibit an exceptional oil/water separation property. The pigmented fibrous material shows higher contact angle than the pristine electrospun membrane. The polycarbonate–copper phthalocyanine fibers possess high separation efficiency and high permeation flux (7073 Lm2 h1) for petrol.

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Superwettable Janus Fabric Based on Polyester and Nano-PTFE Particles

In our previously reported work, we described fabrication of Janus membrane using surfactant stabilized nano-PTFE dispersion onto cotton fabric via Meyer Rod coating technique, which demonstrated improved self-cleaning and anti-icing characteristic, flame-retardancy, water permeation breakthrough pressure, and comparable separation efficiency (Gupta and Kandasubramanian 2017a). The environmental catastrophes arising due to Oil-spill often involves harsh working condition, and hence the applied materials used during the oil-water separation need to be efficient, and mechanically robust during the service operations. Present study reports thermally stimulated superhydrophobic/oleophilic Janus fabric, exhibiting selective directional fluid gating effect for Oil-Water mixture, fabricated by impregnating nano-sized dispersion of Poly (1,1,2,2-tetrafluoroethylene) onto Poly (ethyl benzene-1,4dicarboxylate) based fabric via scalable and automized padding technique. The engineered fabric exhibited superhydrophobicity-oleophilicity (WCA ¼ 172
and OCA ¼ 0
), low ice-adhesion, and self-cleaning characteristic which is attributed to its hierarchical surface texture having closely spaced randomly oriented nano-needle forest morphology generated on microfiber based fabric as revealed by FE-SEM micrograph. The as-prepared Janus fabric exhibited selective directional fluid gating with excellent separation efficiency up to 98% for various petroleum oils/solvents, with stable repeatability (till 30 recurrences), flame-retardancy, anti-icing characteristic, and a maximum permeation fluxes for n-Hexane, Toluene, and Petrol, i.e., 14597.65 Lm2h1, 5016.87 Lm2h1, and 7501.904 Lm2h1, respectively, for Janus fabric. The reported Janus fabric also retained its intrinsic properties under extreme environmental conditions such as UV-irradiation (254 nm), hypersaline solution, extreme acidic and alkaline solutions (pH 1 and 14, respectively), high temperature (180
C), and subzero temperature (20
C). The commercial polyester fabric based on Poly (ethyl benzene-1,4-dicarboxylate) (plain weaved, 180 g/m2, thickness ¼ 190 μm, warp ¼ 38/cm, weft ¼ 38/cm) was procured from Ahmedabad Textile Industry’s Research Association, India. Fluorosystem dispersion, i.e., Poly (1,1,2,2-tetrafluoroethylene), was acquired from the Chemours Company, India (Teflon™ PTFE DISP 33LX, PTFE content—61%, Density—1.52 g/cm3, particle size ~ 0.220 μm, aided with wetting agent) (Chemours 2016). The Janus fabric was fabricated using a stable fluorosystem dispersion onto the commercial Polyester fabric via padding technique (Fig. 14.9a). The padding rate and roller squeezing pressure were maintained at 1 m/min and 200 kN/m, respectively, to get the uniformity in coated fabric. For ensuring the complete penetration and deposition of dispersion inside the microfiber pores of the fabric, three padding cycles were carried out sequentially (Fig. 14.9a). The rotation of the squeezing rolls in the padding system creates tangential speed at the roller nips, due to which a drag force is generated on the dispersion in forward direction during coating on the fabric (Birley et al. 1992). The uniform distribution of pressure, i.e., 200 kN, from the

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Fig. 14.9 (a) Padding Process, FE-SEM micrograph of (b) pristine polyester fabric, and (c) dispersion-coated fabric, (d) FE-SEM micrograph of dispersion-coated fabric at 300 nm scale, (e) simulated 3D projection of FE-SEM micrograph

squeezing rollers removes excess quantity of dispersion from the fabric, whereas the forward drag force ensures the uniform velocity distribution during padding (Birley et al. 1992; Makowski et al. 2016) (Fig. 14.9a, b). Therefore, the synergistic effect of these two parameters helps in maintaining the constant thickness, i.e., 30 μm, and controls the shrinkage and stretch growth in the coated Janus fabric (Birley et al. 1992; Makowski et al. 2016; Sarwar et al. 2017). Polyester fabric samples of dimension 12  12 cm2 were cleaned with Acetone for 20 min and then with DI water for 15 min in an ultrasonic bath (20 kHz frequency) for removing any traces of impurities. After coating fluorosystem dispersion onto the cleaned polyester fabric, it was kept at room temperature for 2 h. Then, the coated fabric was passed in an air-oven at 150
C for 6 h for the complete removal of water phase and wetting agent from the dispersion (Chemours 2016). The engineered Janus fabric showed a thickness of 30 μm (measured by digital thickness gauge meter), then the fabric samples were taken for further characterizations.

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The morphology of the fabric was analyzed via Field Emission Scanning Electron Microscope (Carl-Zeiss AG, JSM-6700F, Germany) with 3 kV accelerating voltage. The wetting study was done using static contact angle goniometer (DSA100, Krüss GmbH, Germany) using DI water (8 μL volume) at room temperature. The water contact angles (WCA) were measured after resting the droplet on the surface for 2 s, then the average of five reading was taken. The anti-icing study was performed by sprinkling DI water droplets of 20 μL volume (cooled at 5
C) using a micropipette (Tarsons 2–20 μL (microliter) T20 Accupipette) on the pre-cooled surface of Janus fabric samples, which were subsequently placed in a NEWTRONIC deep-freezer system for 15 min (at 20
C) to get the frozen droplets. Then, the pressurized airstream was progressively passed on the fabric samples at an approximate rate of 10 kPa/s via air-compressor until the frozen water droplets detached and rolled-off. The distance between the air gun and Janus fabric was maintained at 30 cm. The abrasion resistance study was done using a wear test machine (Pin-On-Disc Friction and Wear Test Rig, Magnum Engineers, India). The tensile testing of the fabric samples was done using a customized tensile test machine (Load capacity ¼ 50 N, specimen dimension ¼ 5  3 mm). The FT-IR study was done using a Perkin-Elmer Spectrum BX FTIR system (Perkin-Elmer Inc., USA) at room temperature in the range 4000–600 cm1.

14.5.1 Results and Discussion The FE-SEM micrograph of pristine polyester fabric depicted in Fig. 14.9a indicates the smooth surface morphology of the weaved microfibers having average diameter of 7 μm (calculated via ImageJ software). Figure 14.9b shows the surface of the coated polyester fabric, where, after the addition of dispersion narrowly spaced randomly oriented Nano-needles (average diameter 220 nm, calculated via ImageJ software) forest like morphology is observed on the microfibers of fabric as depicted in Fig. 14.9d, e (Qi et al. 2002; Zhang and Seeger 2011; Wu et al. 2014). A hierarchical surface morphology with closely spaced Nano-needles observed in Fig. 14.9d has been simulated three-dimensionally via Scanning Probe Image Processor software (SPIP, Image Metrology) as shown in Fig. 14.9e. Threedimensional simulation reveals convex-shaped Nano-needles with rough surface texture, which is essentially required in superhydrophobic surfaces (Cheng et al. 2010; Sahoo and Kandasubramanian 2014b; Sahoo et al. 2015). The hierarchical surface morphology (Fig. 14.9d, e) of Janus fabric resembles the surface-texture on the wings of Chremistica maculata, i.e., Cicada insect, which has non-wetting property to protect itself in living environment from rain/water sources (Sun et al. 2009; Darmanin and Guittard 2015). The analytical comparison of the SEM micrographs and simulated surface of the Janus fabric revealed analogous multi-scaled hierarchical morphology on both surfaces, demonstrating successful bio-mimicking of hydrophobic Cicada wings. The bio-mimicked hierarchical surface of Janus fabric

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exhibited superhydrophobic-superoleophilic nature, low ice-adhesion and selfcleaning ability. After passing the Fluorosystem dispersion-coated Janus fabric in an air-oven at 150
C for 6 h, the water phase and wetting agent evaporate and the Janus fabric undergoes thermal transition by exhibiting superhydrophobicity. Superhydrophobicity in Janus fabric could be attributed to chemical interactions between low surface energy fluoro molecules/particles, i.e., ―CF2― and ―CF3, and Polyester end groups, i.e., ―COOH and ―OH. The Fluoro molecules in the form of ―CF2― and ―CF3 present in Fluorosystem dispersion are able to migrate to the surface of the Polyester fabric at low temperature conditions, where the macromolecules of polymer are practically immiscible and compete for surface exposure. The polar end groups, i.e., “―COOH” and “―OH,” present in Polyester groups, are capable of forming Hydrogen bonding, which possibly interact with ―CF2 and ―CF3 molecules at a thermal transition temperature of 150
C, where surfactants and stabilizing agents get evaporated (Lebold et al. 2000; Qi et al. 2002; Demir et al. 2017). This thermal transition temperature helps in promoting the migration of Fluoro particles/molecules to migrate to the surface of the Polyester fabric, where they interact with polar ―COOH and ―OH end groups via Hydrogen bonding, thereby decreasing the surface energy and imparting superhydrophobicity in the Fluorosystem-coated polyester fabric (Smith 1980; Lebold et al. 2000; Qi et al. 2002; Demir et al. 2017). FT-IR characterization of fabric samples revealed small peak at 2866 and 1341 cm1 corresponds to stretching of ―C―H― bond in methylene groups, whereas peaks at 714 and 875 cm1 correspond to in-ring bond stretching of ―C―H― groups, present in ester linkage of polyester chain (Cecen et al. 2008; Parvinzadeh and Ebrahimi 2011). The characteristic peak at 1706 cm1 belongs to ―C―O― bond stretching of ester group in polyester, whereas other distinct peaks at 1239, 1079, and 1009 cm1 correspond to ―C―O― bond stretching of ester group (Parvinzadeh and Ebrahimi 2011). The small peak at 1402 cm1 corresponds to ―C―C― bond stretch present in aromatic ring (Cecen et al. 2008; Parvinzadeh and Ebrahimi 2011). The characteristic peaks at 1217 and 1149 cm1 correspond to ―C―F― bond stretching, and the small peak at 631 cm1 corresponds to wagging of ―F―C―F― bond, i.e., to and fro motions in CF2 group, present in Fluorosystem (Huang et al. 2011; Zhang and Seeger 2011).

14.5.2 Wetting Behavior of Janus Fabric The wetting behavior of Janus fabric was extensively studied using water and oils/ solvents. The superhydrophobic nature attained at 150
C in Janus fabric is accredited to the evaporation of water phase from the dispersion (Chemours 2016). However, the improvement in superhydrophobic nature from 120 to 150
C may be attributed to the increase in surface protuberance height leading to elevated surface roughness as demonstrated by Sahoo et al. (Sahoo and Kandasubramanian

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Fig. 14.10 (a) Superhydrophobicity of Janus fabric, (b) Wettability of Janus fabric, (c) Separation efficiency, (d) Permeation flux analysis, (e) Coefficient of friction, (f) Tensile strength of Janus fabric

2014b; Sahoo et al. 2015). The Janus fabric did not show enhancement in water contact angle (WCA) after 150
C; however, beyond a heat treatment of 280
C, polyester fabric starts losing its intrinsic property. The maximum WCA achieved was 172
 3
at a heat treatment of 150
C (Fig. 14.10a). Various wetting theories have been proposed for describing the wettability of surfaces, but Wenzel and Cassie-Baxter are widely considered for explaining the superhydrophobicity in surfaces (Wenzel 1936; Cassie and Baxter 1944; Sahoo and Kandasubramanian 2014b; Sahoo and Balasubramanian 2014; Sahoo et al. 2015; Gupta and Kandasubramanian 2017a). According to Wenzel’s theory, the roughness in homogeneous surface is main driving factor for improving the hydrophobicity,

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which is explained in previous section (Wenzel 1936; Sahoo and Kandasubramanian 2014b; Gore and Kandasubramanian 2018). Cassie-Baxter’s theory explained the superhydrophobicity of a heterogeneous rough surfaced Janus fabric having porosity, and it is given by equation mentioned in previous section (Sahoo and Kandasubramanian 2014b; Arora and Balasubramanian 2014; Sahoo et al. 2015; Gupta and Kandasubramanian 2017a). The Cassie-Baxter’s regime has a porosity on its surface which entraps the air in between liquid and solid phase, thereby creating small air-pockets; this subsequently leads to the enhancement in hydrophobic nature of the surface (Cassie and Baxter 1944; Sahoo and Kandasubramanian 2014b; Sahoo et al. 2015; Gupta and Kandasubramanian 2017a; Gore and Kandasubramanian 2018). Sahoo et al. have reported that elevation in WCA of a superhydrophobic surface mainly depends on the Cassie-Baxter’s model, where the porous structure and roughness are the main contributing features, which leads to the formation of air-pockets, thereby reducing the adhesion between liquid and solid phase (Sahoo and Kandasubramanian 2014b; Sahoo and Balasubramanian 2014; Sahoo et al. 2015). Superhydrophobic surface possesses self-cleaning ability as demonstrated by the “Lotus effect,” where a space between hierarchical structure (in our case convexshaped Nano-needles) and surface is filled by air bubbles, thereby facilitating the water droplet to easily roll-off, which takes away any contaminant particles present on the surface along with them leading to a “self-cleaning action” (Bhushan et al. 2009; Fernández et al. 2017). The self-cleaning ability of developed Janus fabric has been demonstrated in a movie. The Janus fabric revealed superoleophilicity, i.e., WCA ~ 0
, for various petroleum products (petrol, diesel, and engine oil) and organic solvents as shown in Fig. 14.10b. This superoleophilic nature of Janus fabric is attributed to the synergistic effect of low surface energy (imparted by fluorosystem) and hierarchical multiscaled texture (Burkarter et al. 2007). The surface tensions of petrol (γp ¼ 29 mN/m), engine oil (γe ¼ 65 mN/m), and diesel (γp ¼ 29.5 mN/m) are lower than water (γw ¼ 72.8 mN/m), and therefore they get easily absorbed on the surface of Janus fabric. The large difference in surface tension between water and oil imparts the superhydrophobic and superoleophilic nature to the Janus fabric at room temperature, which can be utilized for oil-water separation (Zhou et al. 2013; Wang et al. 2015d; Gu et al. 2017; Gore and Kandasubramanian 2018). Superhydrophobic material should be able to withstand an extreme environmental conditions like subzero temperature (0 to 20
C), saline water, UV-irradiation, thermal stability, and chemical durability; considering this, the wettability of developed Janus fabric was comprehensively investigated under abovementioned environmental conditions. Wettability of Janus fabric was evaluated in hypersaline, i.e., Sodium Chloride, solution for different concentrations from 10 to 40%. Study shows that Janus fabric maintains its superhydrophobicity till 30% saline concentration with a WCA of 152
 3
, but slightly decreases to 148
 3
in 40% saline concentration, which is attributed to the increased ionic nature of saline solution at higher concentration leading to reduction in surface tension of solid/liquid phase (Serrano-Saldaña et al.

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2004). The study demonstrates that the engineered Janus fabric maintains superhydrophobicity in hypersaline solution from 0 to 30% concentrations. The wetting behavior of Janus fabric was studied for various subzero temperatures ranging from 20 to 0
C, where it showed a decrease in WCA with decreasing temperature. The WCAs measured were 165
at 0
C and 139
at 20
C; this reduction in hydrophobicity could be ascribed to the transition from Cassie-Baxter state at 0
C to Wenzel state at 20
C. This transition might take place due to the synergistic effect of reduced temperature and condensed water phase in the space between nano-needles and rough surface of Janus fabric (Dorrer and Rühe 2007; Sahoo et al. 2015; Gupta and Kandasubramanian 2017a; Gore and Kandasubramanian 2018). The Janus fabric possessing frozen water droplets on its surface was bombarded with a pressurized airstream at a progressively increasing rate of 10 kPa/s, the droplets started to roll off after 5 s and detached within 8 s, after reaching an air pressure of 90 kPa. The anti-icing property of Janus fabric has been shown in movie. The Janus fabric was bombarded with UV-irradiation of 254 nm wavelength for 72 h, which showed reduction in WCA from 172
 3
to 162
 3
, still maintaining superhydrophobic nature. The superhydrophobic performance of Janus fabric under UV-irradiation could be attributed to the presence of fluorosystem possessing a stable C―F― bond which has very short bond length of 1.35 Å and a bond energy of 504 kJ/mole (Xiu et al. 2008; Zhang and Seeger 2011; Wu et al. 2014; Lim 2016). The chemical durability Janus fabric was investigated for establishing its robustness in strong acidic and alkaline conditions. In this study, the developed Janus fabric was immersed into the acidic (pH ¼ 1) and alkaline (pH ¼ 14) solutions for 72 h. The Janus fabric being exposed to strong corrosive environment showed some deterioration in WCA, but still maintained its superhydrophobicity. As mentioned earlier, the stability of the Janus fabric could be attributed to the presence of fluorosystem which contains a stable ―C―F― bond which is also linked to its chemical resistance (Xiu et al. 2008; Zhang and Seeger 2011; Wu et al. 2014). The decrease in WCA could be ascribed to the oxidation/degradation in polyester polymer which has low chemical resistance against highly acidic (pH ¼ 1) and alkaline (pH ¼ 14) solutions (Buxbaum 1968). The significant reduction of WCA in alkaline environment might be associated with the release of Terephthalic Acid and Ethylene Glycol due to the oxidation/degradation in polyester, which is also evidenced by the discoloration in Janus fabric (Buxbaum 1968; Brueckner et al. 2008). Considering, the abovementioned studies, the Janus fabric has demonstrated a stable performance under various extreme environmental conditions.

14.5.3 Absorption Study In order to establish the potential of Janus fabric for oil/water separation, the oil absorption study was performed. The Janus fabric efficiently absorbed and separated

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Diesel from a dyed water within 2 s, without any staining. The Janus fabric demonstrated high absorption capacity of 1.5 (wt/wt) for engine oil within 15 s. The absorption capacity of Janus fabric with Diesel and Petrol was 0.8 (wt/wt) and 0.7 (wt/wt), respectively, the absorption time for both oils was 2 s. The organic solvents used in the study were n-hexane and Toluene for which the absorption capacities were 0.4 (wt/wt) and 0.6 (wt/wt), respectively, with an absorption time of 2 s each. The absorption capacity of Janus fabric was calculated by following equation (Bastani et al. 2006; Mishra and Balasubramanian 2014; Arora and Balasubramanian 2014; Gupta and Kandasubramanian 2017a; Gore and Kandasubramanian 2018): Absorption Capacity ¼

W2  W1 W1

ð14:10Þ

where W1 ¼ weight of fabric before immersion, W2 ¼ weight of fabric after absorption. The recycling ability and durability of Janus fabric are the main factors for considering practical application in oil-water separation, which was evaluated by the number of washing cycles and absorption/combustion cycles. After each absorption cycle, the oily Janus fabric was rinsed as per the standard procedure mentioned in previous section, and further the cleaned Janus fabric was dried in an air heating oven at 150
C for 30 min to retain its superhydrophobicity. The Janus fabric demonstrated absorption capacity of 0.8 (wt/wt) for diesel till 30 washing cycles, demonstrating robustness for continuous oil-water separation. The same Janus fabric was evaluated for its superhydrophobicity after every five absorption cycles. The Janus fabric effectively retained its superhydrophobicity till 30 washing cycles with WCA from 172
 3
to 160
 3
. Similar oil absorption capacities were also investigated for Petrol, Engine oil, and organic solvents (Fig. 14.10b). The absorption capacity was also evaluated for flame tested Janus fabric, where it showed a linear decreasing trend from 0.8 (wt/wt) to 0.1 (wt/wt) till 10 absorption cycles. The same combusted Janus fabric exhibited superhydrophobicity till 10 absorption cycles with WCA from 172
 3
to 150
 3
. The ability of Janus fabric to retain superhydrophobicity even after burning is attributed to the presence of Fluorosystem (Huang et al. 2011; Zhang and Seeger 2011; Chemours 2016; Drobny 2016; Lim 2016). The absorption kinetic study of all the petroleum oils and organic solvents absorbed by the Janus fabric can be done by Fractal Like-Linear Driving Force model (FL-LDF) (Mishra and Balasubramanian 2014; Khosravi and Azizian 2017):   mt ln 1  ¼ D0 t α mmax

ð14:11Þ

where mt ¼ Mass of Oil absorbed by the Janus Fabric at time t, mmax ¼ Maximum Oil absorption capacity of Janus Fabric per unit mass,

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D0 ¼ Mass transfer coefficient, (proportional to the diffusion coefficient of oil diffusion into pores, i.e., in terms of pore size and oil viscosity, t ¼ Absorption time, and, α ¼ constant, (denotes different pores with different sizes present in fabric.

14.5.4 Separation Analysis The potency of Janus fabric as an oil-absorbing material has been investigated via oil-water separation study (Pan et al. 2008). The separation efficiency of Janus fabric was studied for various oils/solvents with water (ratio of 1:1). The oil-water mixture was poured on Janus fabric, where only oil passed through the Janus fabric to container and the water was retained on the surface, due to the difference in the surface tensions of oil/water and Janus fabric (Pan et al. 2008; Sahoo and Kandasubramanian 2014b; Xue et al. 2014; Ma et al. 2016; Gupta and Kandasubramanian 2017a; Gore and Kandasubramanian 2018). The separation cycles were repeated for 30 times as shown in Fig. 14.10c, where the Janus fabric maintained the separation efficiency more than 95% for all washing cycles. The stability of Janus fabric was evaluated by measuring WCA after every five washing cycles as depicted in Fig. 14.10c. The Janus fabric retained superhydrophobicity within 172
 3
to 165
 3
till 30 washing cycles. Similar separation efficiencies were also investigated for Petrol and Engine oil (Fig. 14.10b). The directional movement of fluid through the capillary channels of fabric or membrane are principally governed by the porosity and surface characteristic (Gu et al. 2017). Therefore, the breakthrough pressure for the engineered Janus fabric for possible penetration of water through its cross-section can be calculated by considering the fabric as an interwoven mesh with predominant cylindrical texture as described in previous section (Tuteja et al. 2008a, 2008b; Choi et al. 2009; Chhatre et al. 2010; Tian et al. 2014; Mates et al. 2014; Song et al. 2014). The polyester fabric used in the study exhibited an average radius of cylinders 105 μm and an average inter-cylinder spacing of 275 μm as calculated from FE-SEM micrograph (Fig. 14.9b) using ImageJ software. The petrol, engine oil, and diesel spontaneously pass through the Janus fabric owing to the difference in their surface tensions (γJanus fabric > γoil), but, for water (γJanus fabric < γwater) to pass through the Janus fabric needs to cross the breakthrough pressure of 1.33 kPa, which has been calculated from Eq. (14.11). Similar results were reported by Zhou et al., where the breakthrough pressure for water was theoretically calculated and compared with experimental results (Zhou et al. 2013; Wang et al. 2015d).

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14.5.5 Permeation Flux Analysis The permeation flux study of Oils and Solvents is important for establishing its permeation rate. Therefore, the permeation flux study was performed on Petrol, Diesel, n-Hexane, Toluene, and Water for Janus fabric and Pristine Polyester samples. Previously described equation was used for calculating permeation fluxes (Gu et al. 2017; Liu et al. 2017; Yu et al. 2017b; Gore and Kandasubramanian 2018). The results (Fig. 14.10d) demonstrated that n-Hexane solvent exhibited highest permeation flux for Pristine, i.e., 85277.46 Lm2h1, Janus fabric, i.e., 14597.65 Lm2h1, due to its low surface tension, as compared to other used oils and solvents. Petrol and Toluene have demonstrated permeation fluxes around 48892.41 Lm2h1 and 67906.12 Lm2h1 for Pristine Polyester and 5016.87 Lm2h1 and 7501.904 Lm2h1 for Janus fabric, respectively. Diesel has showed a permeation flux of around 33641.57 Lm2h1 and 3771.011 Lm2h1 for Pristine Polyester and Janus Fabric, respectively. As observed from the permeation flux results, only Pristine Polyester fabric allowed the Water to percolate and pass through its cross-sectional area, whereas the Janus fabric restricted the percolation, and thus the permeation of Water through it, due to superhydrophobic nature, i.e., WCA ~ 172
. The noticeable low permeation flux of oils and solvents for Janus fabric could be attributed to the strong adhesive and capillary forces coupled with acting surface tension arising due to action of interacting Fluoro molecules (Qi et al. 2002; Wang et al. 2015a, c; Brown and Bhushan 2015; Ma et al. 2016, 2017; Yu et al. 2017b). The high permeation rate of Pristine Polyester could be attributed to availability of unfilled pores (Fig. 14.9b), thereby allowing the easy transportation of oils and solvents; however, these pores get filled-up after coating with Fluorosystem dispersion as shown in Fig. 14.9c. The viscosity of Engine oil was high among all oils and solvents, thereby decreasing its permeation rate; therefore, the permeation flux for Engine Oil could not be calculated. Frictional Performance. The potency of the developed Janus Fabric for practical applications as oil-water separation material also related to its mechanical stability. The surface mechanical stability of the Janus fabric was investigated by analyzing its friction resistance and tensile strength. The friction resistance of Janus fabric was evaluated via pin on disc friction and wear test rig by sliding the fabric against Wear disc EN 31 (58–60 HRC) with a speed of 4 cm/s under 5 N load for 10 min, which corresponds to a sliding distance of 24 m. Figure 14.10e depicts the variation of coefficient of friction (COF) with sliding time for pristine polyester fabric and Janus fabric demonstrating larger COF value for Janus fabric (0.38) compared with pristine polyester fabric (0.16) suggesting enhanced friction resistance owing to the presence of mild PTFE particles on stress dissipation suggesting that the PTFE lubrication effect is not sufficient to harmonize the effect of enhanced surface roughness induced by PTFE nanoparticles (Ramalho and Miranda 2005). Although, WCA of Janus fabric reduced from 172
 3
to 142
 3
after abrasion, it still retained its

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hydrophobic characteristic, indicating the robustness of the Janus fabric against mechanical forces (Ramalho and Miranda 2005; Si et al. 2015). The effect of tensile loading on the superhydrophobicity of Janus fabric was investigated via customized strain sensing system with elongation rate of 50 μm/ min. The tensile strength of the Janus fabric (40.3 MPa) is higher than pristine polyester fabric (31.8 MPa) as depicted in Fig. 14.10f and is attributed to the shear response of robust PTFE coating (Testa and Yu 1987). Although Janus fabric suffers fracture as the axial tensile stress reaches 40.3 MPa, nevertheless, the broken fabric sustains still exhibits superhydrophobicity, which concludes that mechanical stretching does not destroy the fabric surface asperities (Tang et al. 2013; Gu et al. 2017). The successive results demonstrate the mechanical stability of the Janus fabric, thereby making it a proficient candidate for marine oil-spill clean-ups (Gore et al. 2018b; Tang et al. 2013).

14.5.6 Summary The Janus fabric demonstrated superhydrophobicity-oleophilicity, and possessed an oil-water separation efficiency of 98% for various oils/solvents, with stable repeatability (till 30 washing cycles). The Janus fabric also demonstrated self-cleaning and anti-icing characteristic, flame-retardancy, high thermal stability (180
C), subzero temperature stability (20
C), and a maximum permeation fluxes for n-Hexane, Toluene, and Petrol, i.e., 14597.65 Lm2h1, 5016.87 Lm2h1, and 7501.904 Lm2h1, respectively, for Janus fabric. The Janus fabric retained intrinsic properties in harsh environments like UV-irradiation (254 nm), hypersaline solution, and acidic-alkaline solutions (pH ¼ 1–14). Subsequently, theoretically calculated breakthrough pressure (1.33 kPa) for water permeation through Janus fabric establishes it as a proficient oil-absorbing material for effective cleaning of massive oil-seepages.

14.6

Superwettable Sponges and Foams for Effective Oil-Water Separation

Sponges and foams are one of the readily available and economical materials exhibiting porosity and primary wetting characteristics. These materials generally absorb liquids with abundant water phase, which is attributed to the oxygen-rich groups which manifest in their structure, thus limiting their applicability mostly for the water absorption, due to lower selectivity. However, by doing the chemical modification of these materials with low surface energy materials, e.g., nanoparticles, their intrinsic properties can be tailored, i.e., superhydrophobicity/ oleophilicity, for the effective utilization in the selective absorption of oils and

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solvents from the oily wastewater. A wide range of methods have been reported for the fabrication of superhydrophobic fabric materials, which can also be utilized for the functionalization of the foams/sponges, e.g., dip-coating, in situ method, and block co-polymer grafting (Liang and Guo 2013; Xue et al. 2013, 2014; Mishra and Balasubramanian 2014; Arora and Balasubramanian 2014; Wang et al. 2015a; Ma et al. 2016; Yu et al. 2017a). In one of the studies, Arora and Balasubramanian have fabricated a foam based on polyvinylidene fluoride (PVDF) and nano Silicon Carbide (nano-SiC) (5 wt%) (diameter  100 nm) (Arora and Balasubramanian 2014). They fabricated a foam based on the separation of the solid–liquid phases by utilizing the differential solubilities of the solvent, i.e., Dimethyl Formamide (DMF), and water (non-solvent). The developed PVDF/nano-SiC foam revealed superhydrophobicity (WCA ~ 152
) and oleophilicity (OCA ~ 0
), and further showed an absorption efficiency up to 20.5 and 21.5 times its weight, for paraffin oil and engine oil, respectively (Fig. 14.11). The superhydrophobicity and the mechanical strength of the developed PVDF/nano-SiC foam was attributed to the dispersed nano-SiC

Fig. 14.11 PVDF/nano-SiC foam (a) Superhydrophobicity, (b) FESEM Analysis, (c) Absorption performance of oils and solvents, and (d) Time dependent Absorption (Arora and Balasubramanian 2014). (Reprinted with permission. Copyright 2014, Royal Society of Chemistry)

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Fig. 14.12 PVDF/CS composite foam demonstrating (a) superhydrophobicity, (b) practical oil-water separation, (c) FESEM analysis, and (d) absorption performance of Camphor soot and PVDF/CS composite foam (Mishra and Balasubramanian 2014). (Reprinted with permission. Copyright 2014, Royal Society of Chemistry)

particles which exhibited the intrinsic hydrophobicity (WCA ~ 130
), and further it helped in improving the strength of the matrix (Arora and Balasubramanian 2014). Further, the PVDF/nano-SiC foam showed reusability up to 4 cycles with mechanical squeezing, along with a minor decrease in absorption performance. They claimed that the incorporation of nano-SiC improved the mechanical strength of the foam, which helped in reducing its withering during service operation. The FESEM analysis of the PVDF/nano-SiC foam revealed a hierarchical morphology with interconnected pores in the structure, with a pore diameter varying in the range of nanometers to micrometers. Further, their FESEM analysis revealed a diversified dispersion and adhesion of the agglomerated nano-SiC nanoparticles on the periphery of the pores of the foam, thus increasing the volume of the inside pores of the foam (Arora and Balasubramanian 2014). In another study, Mishra and Balasubramanian have fabricated a foam based on nano Camphor Soot (diameter ~ 25–60 nm) particles, and intrinsically hydrophobic PVDF polymer via non-solvent induced phase separation method (Mishra and Balasubramanian 2014). The developed PVDF/CS foam revealed superhydrophobicity (WCA ~ 170
), and superoleophilicity (OCA ~ 0
) (Fig. 14.12a, c).

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The 0.5 wt % loading of CS into the PVDF/CS composite foam, improved the oil absorption capacity of up to 2500%, which was mainly attributed to the intrinsic oleophilic nature of the CS particles (Fig. 14.11b, d). The nano CS particles revealed progressively improved absorption capacity till 25 burning cycles, where it effectively absorbed the oil (Mishra and Balasubramanian 2014). The FESEM analysis revealed a nano/-micro porous hierarchical morphology (Fig. 14.12c), which helped in enhancing oleophilicity, and simultaneously (OCA ~ 0
) improving the superhydrophobicity (WCA ~ 170
) in the PVDF/CS composite foam. They claimed that the CS particles exhibited interconnected fractal network structure of nanotubes, thus revealing a fractal-like morphology of carbon nanospheres, where 2 to 3 nanotubes emanate from the nanoparticles, and further the single nanotube help in holding 4 to 8 nanoparticle, thus forming a three-dimensional network which directly enhanced the oil absorption capacity of the PVDF/CS composite foam (Fig. 14.12c). Further, they reported that a single nanotube connects 4–8 nanoparticles, with 2–3 nanotubes growing from a single nanoparticle, thus denoting a threedimensional network. This three-dimensional network of porous CS particles is analogous to a three-dimensional capillary network, which has a propensity for absorption, owing to capillary pressure (Mishra and Balasubramanian 2014). In another study Li et al. have developed a melamine based sponge engineered with Copper Oxide (Cu2O) with different crystal structures of the materials, i.e., cubic, octahedral, and cubo-octahedral, using dip-coating of Cu2O onto melamine sponge (Fig. 14.13) (Li et al. 2018). The developed Melamine/Cu2O Sponge demonstrated a hydrophobicity (WCA ~ 149
) and oleophilicity (OCA ~ 0
). The Density Function Theory (DFT) calculations showed a correlation between surface energy and the hydrophobicity of the developed sponge, where Cu2O with (111) and Cu2O with (100) crystals showed surface energies of 0.73 and 1.29 J/m2, respectively. Further, their DFT calculations also revealed that sponge with octahedral Cu2O (111) showed maximum hydrophobicity (WCA ~ 149
). The Melamine/Cu2O (111) sponge showed maximum separation efficiency of 97.3% for Silicone Oil-Water separation, along with a stable absorption recycling performance till 8 cycles. The SEM analysis of the Sponge revealed a microporous morphology responsible for the enhanced separation and absorption efficacy of the oil-water mixture, along with the nano/-micro scaled dispersion of the Cu2O particles on the surface of the foam. The FT-IR analysis of the Melamine/Cu2O sponge showed a characteristic peak for Cu―O bond stretch at 697 cm1, and the peaks at 2275 and 2360 cm1 revealed asymmetric bond stretching vibration of ―NCO group present in the melamine (Li et al. 2018). The developed Melamine/Cu2O Sponge demonstrated a hydrophobicity (WCA ~ 149
) and oleophilicity (OCA ~ 0
). The Density Function Theory (DFT) calculations showed a correlation between surface energy and the hydrophobicity of the developed sponge, where Cu2O with (111) and Cu2O with (100) crystals showed surface energies of 0.73 and 1.29 J/m2, respectively. Further, their DFT calculations also revealed that sponge with octahedral Cu2O (111) showed maximum hydrophobicity (WCA ~ 149
). The Melamine/Cu2O (111) sponge showed maximum separation efficiency of 97.3% for Silicone Oil-Water separation,

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Fig. 14.13 Melamine/Cu2O Sponge (a) SEM analysis, (b) contact angle measurement, (c) Separation efficiency, (d) Practical oil-water separation (Li et al. 2018). (Reprinted with permission. Copyright 2018, Springer Nature)

along with a stable absorption recycling performance till 8 cycles. The SEM analysis of the Sponge revealed a microporous morphology responsible for the enhanced separation and absorption efficacy of the oil-water mixture, along with the nano/micro scaled dispersion of the Cu2O particles on the surface of the foam (Fig. 14.13). The FT-IR analysis of the Melamine/Cu2O sponge showed a characteristic peak for Cu―O bond stretch at 697 cm1, and the peaks at 2275 and 2360 cm1 revealed asymmetric bond stretching vibration of ―NCO group present in the melamine (Li et al. 2018). In one more study, Wang et al. (2015a-d) have fabricated foam based on singlestep copolymerization technique using organic-coated three-dimensional material based on dopamine and octadecylamine (ODA) onto commercially available nickel foam. The developed foam based on nickel revealed superhydrophobicity (WCA ~ 154
) (water tilting angle ~4  1
) along with superoleophilicity (OCA ~ 0
) (Fig. 14.14) (Wang et al. 2015b). The SEM analysis revealed a macroporous structure in the foam along with nano/-micro scaled protrusions of the dopamine and ODA on the foam surface. Further, the surface engineered Nickel foam exhibited permeation flux of 2823, 2337, and 1844 Lm2h1 for Lubricating oil, motor oil, and engine oil respectively. FT-IR analysis of the developed foam

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Fig. 14.14 (a) SEM analysis of Nickel foam, (b) Permeation flux analysis (Wang et al. 2015b). (Reprinted with permission. Copyright 2018, Springer Nature)

revealed peaks at 1945 and 610 cm1, which they attributed to the benzene ring present in the dopamine structure. Further, they reported peaks at 3331, 2918, and 2850 cm1 associated with the stretch vibrations of the N―H, ―CH3, and ―CH2― of octadecylamine (ODA), which confirmed the copolymerization of the coated substrates onto the surface engineered foam. The developed Nickel foam exhibited ten absorption cycles, and still maintained hydrophobicity, i.e., WCA > 142
(Wang et al. 2015b). Though the Sponge and Foam based superwettable materials have been widely explored by researchers, their poor absorption performance after some usable cycles due to filling up of the pores is one of the major issues, and hence limits their applicability for long-term service applications (Mishra and Balasubramanian 2014; Arora and Balasubramanian 2014).

14.7

Conclusion and Future Direction

The contamination of oceanic and ground water sources due to oil seepages and industrial waste solvents has emerged as a global issue urging for immediate counter measures to epitomize the catastrophic repercussions on sensitive ecological system. Considering this, various superwettable materials have been explored by the researchers for the efficient oil-water separation. In this context, we have extensively discussed the nanoparticle engineered superwettable, i.e., superhydrophobic/ superoleophilic, Janus materials like fabrics/textiles, membranes, electrospun nanofibers, and their practical applicability for the effective oil-water separation. Further, we have also discussed the nanoparticle decorated sponges and foams, and their pertinence for mitigating the oil-water separation challenges. In connection with superwettability of the materials, we have also explained the main driving

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factors and the principle wetting theories such as Wenzel and Cassie-Baxter in the present book chapter. Though the current research works have led to the development of highly efficient superwettable materials for practical oil-water separation, still their cost-efficiency, life cycle performance during the service period, and environment friendliness are some of the limiting factors. Considering, the rising environmental concerns, and the awareness towards the global warming, the current research focus has shifted towards developing the environment friendly, biodegradable, and robust superwettable materials for efficient oil-water separation (Sahoo and Kandasubramanian 2014b; Gore and Kandasubramanian 2018).

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Chapter 15

Nanotechnology for Wastewater Treatment and Bioenergy Generation in Microbial Fuel Cells M. J. Salar-García and V. M. Ortiz-Martínez

Contents 15.1 15.2 15.3 15.4

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Microbial Fuel Cells for Wastewater Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Nanomaterials for MFC Anode Electrodes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . MFC Membrane Composites Including Nanomaterials . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.4.1 PES-Based Membranes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.4.2 Sulfonated Poly Ether Ether Ketone-Based Membranes . . . . . . . . . . . . . . . . . . . . . . 15.4.3 Polyvinyl Derived-Based Membranes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.4.4 Sulfonated Polystyrene-Ethylene-Butylene-Polystyrene-Based Membranes . 15.4.5 Nafion-Based Membranes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.5 Nanomaterials for MFC Cathode Electrodes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.6 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

15.1

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Introduction

Environment protection and energy crisis are two major concerns that need to be addressed for sustainable economic growth. Water contamination proves to be one of the most alarming human effects on the environment. Industry, urbanization, and agriculture often introduce several types of pollutants including heavy metals, agrochemicals, and drugs into water environments. Recently, bio-electrochemical M. J. Salar-García (*) · V. M. Ortiz-Martínez (*) Department of Chemical and Environmental Engineering, Campus Muralla del Mar, Technical University of Cartagena, Cartagena, Spain Department of Chemical Engineering, Campus Espinardo, University of Murcia, Murcia, Spain e-mail: [email protected]; [email protected] © Springer Nature Switzerland AG 2019 R. Prasad, K. Thirugnanasambandham (eds.), Advanced Research in Nanosciences for Water Technology, Nanotechnology in the Life Sciences, https://doi.org/10.1007/978-3-030-02381-2_15

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Fig. 15.1 Schematic representation of MFC

systems (BES) have emerged as a preeminent technology for environmental management. In particular, Microbial Fuel Cells (MFCs) attract huge interest due to their capacity to produce bioenergy from multiple substrates, including complex sources such as urban and industrial wastewater and biomass wastes (Santoro et al. 2017). MFC devices, usually formed by respective anode and cathode compartments, exploit microbial metabolism to oxidize organic and inorganic matters (see Fig. 15.1). Theoretically, electrons and protons are produced by inoculated or naturally occurring bacteria in a given substrate. Electrons are transferred to the anode electrode and externally led through a conductive material containing a resistor for electricity generation, while protons go through the membrane that separates anodic and cathodic compartments. In practice, other positive and negative ions are involved in the ion-exchange process through the membrane. At the cathode, oxygen is often used as oxidant because of its high reduction potential. This way oxygen, one of the main cathodic reactions consist of the reduction of oxygen into water by the combination with the electrons and protons coming from the oxidation process in the anode chamber (Hernandez-Fernandez et al. 2015). MFC technology offers enormous potential for waste reuse. Moreover, MFC systems provide an electrochemical framework to perform and integrate several processes for pollution degradation, including persistent industrial contaminants such as xenobiotic compounds, processes designed to reduce carbon footprint by fixing carbon dioxide, separation processes (e.g., metal removal) and product valorization. The development of MFC technology can be approached by its up-scaling

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for large-scale application of wastewater effluents or by its miniaturization for portable or more-scale applications (Santoro et al. 2017). For instance, MFCs can be micro-scaled to be employed for powering portable electronic devices. MFCs are complex systems involving bioelectrochemical reactions and physicalchemical processes. The technology has been intensively investigated through experimental works in the lasts two decades in order to enhance its efficiency in terms of energy generation and wastewater treatment capacity. The optimization and analysis of MFCs require the knowledge of different scientific and engineering backgrounds, ranging from electrochemistry and microbiology to material, environmental, and chemical engineering. Despite the numerous advantages offered by MFCs, this technology still shows some limitations that hinder its up-scaling and the spreading for practical implementation. In this sense, many efforts have been addressed to enhance MFC efficiency through the development of new materials and design and the study of biodegradability of waste effluents in the anodic chamber of these devices. Recent advances on MFC technology include intensive research on new materials to be used in the main components of the systems (anodic/cathodic electrodes and separator) in order to overcome the bottlenecks of the systems, namely low power density levels from wastewater substrates. Among them, nanostructured materials have drawn high interest for the fabrication of advanced electrode and separators due to their enhanced properties such as high specific surface, increased transfer rates and in many cases low costs and ease of fabrication (Zhao et al. 2017). The use of nanotechnology aims at: (1) increasing electron transference between microbes and anode electrode surface, (2) improving ion exchange through the separator, and (3) enhancing oxygen reduction reaction at the cathode. The improvement of all these key factors leads to increased overall MFC performance, implying higher power generation and wastewater treatment capacity. This chapter deals with the recent advances made on the application of nanotechnology in MFC systems as an effective strategy to improve the performance of the technology. Next sections address nanostructured materials for the three main components that form MFC systems, divided into anode electrode, separator, and cathode electrode materials.

15.2

Microbial Fuel Cells for Wastewater Treatment

Global warming and water scarcity have promoted the interest in reusing wastewater for clean water and energy production, as well as nutrients extraction in order to synthesize agricultural fertilizers. Conventional methods such as aerobic or anaerobic treatments still show some limitations. In the case of aerobic treatment, it has high energy requirements and produces large amounts of wastes. Moreover, this type of procedure is not able to benefit from the wide variety of compounds contained in wastewaters. Regarding anaerobic treatments, they are capable to producing methane from organic matter oxidation, which can be directly transformed into electricity.

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Direct conversion of chemical energy into electricity

Generation of low amount of wastes

No gas treatment requirements

MFC ADVANTAGES

Self-suficient

Operation under environment condition

Fig. 15.2 Advantages of MFCs over conventional wastewater treatments

However, this technique is not able to benefit the whole energy potential of wastewater and the treated water still need a posttreatment stage in order to increase its quality and fit the legal requirements for reuse (He et al. 2017; Hernandez-Fernandez et al. 2015). Microbial fuel cells have demonstrated to be a promising technology capable to address two of the most concerning environmental issues: (1) fossil fuel depletion and (2) water scarcity. MFCs allow us to produce bioenergy and treat wastewater simultaneously. This technology offers several advantages in comparison with conventional wastewater treatment (see Fig. 15.2) (Hernandez-Fernandez et al. 2015; Kim et al. 2008; Logan and Regan 2006). Pure substrates such as glucose, fructose, or sodium acetate have been widely employed as fuels in MFCs. However, the potential of this technology lies on using complex substrates such as real wastewater. Table 15.1 contains some of the most common types of wastewater used as source of energy in MFCs and the performance of the system in terms of wastewater treatment capacity. MFCs have demonstrated to be a suitable technology for the removal of different types of compounds such as nutrients, azo dyes, or even heavy metals. Nitrogen and phosphorous-based compounds can be efficiently removed in MFCs, especially in biocathode chambers. They can be recovered as ammonia or struvite (MgNH4PO4.6H20). Other potential application of MFC is the removal of metal ions such as Cr (VI), Cu (II), Zn (II), or Fe(III), among others. The high redox potential of some of these heavy metals makes possible that they can act an electron

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Table 15.1 Types of wastewater sources used in MFCs Complex substrate Domestic wastewater Brewery industry wastewater Food processing wastewater Confectionary industry wastewater Dairy industry wastewater Paper recycling industry wastewater Textile industry wastewater

COD removal (%) 83 20.7 86 75 90.46 70 71

References Cusick et al. (2010) Wen et al. (2010) Mansoorian et al. (2013) Technology et al. (2016) Mansoorian et al. (2016) Cheng et al. (2011) Kalathil et al. (2012)

acceptors, allowing their reduction and precipitation. Other type of pollutants such as azo dyes have also been successfully removal from wastewater by using MFCs (Gude 2016; Hernandez-Fernandez et al. 2015; Pandey et al. 2016). Although MFC technology offers several advantages over the conventional wastewater treatment, they still need further improvement in order to facilitate their large-scale application. Nanotechnology is one of the most promising fields for the design of advanced and high efficiency materials. Their application in different research fields including MFCs has proven to be a success. In the case of MFCs, nanotechnology has been employed to improve their performance in term of power output and wastewater treatment capacity. Nanomaterials have been used to increase the activity of the catalyst for the oxygen reduction reaction, to facilitate the growth of biofilm around the anode electrode or even to improve the selectivity of the separator.

15.3

Nanomaterials for MFC Anode Electrodes

Anode performance is a key factor in MFC systems that can limit the power level achieved. Generally, the anode reaction is greatly influenced by the surface properties of the anode material employed. Thus, the improvement of anode electrode through new materials and design is critical to enhance the efficiency of this technology for practical implementation. One of the aims in using nanomaterials for anode fabrication is to enhance the electron transfer mechanisms between microorganism, which act as biocatalysts in the anode chamber, and the material forming the anode electrode itself, to increase current generation. Nanostructured materials can be used for the modification of the surface of a given electrode made of other material types or as base electrode material. Commonly used anode nanomaterials include carbon and metal compounds with high conductivity, high specific surface areas, chemical stability, and biocompatibility. This last property is fundamental since they must ensure the affinity connection between the microorganisms and the electrode for efficient electron transference. Carbonaceous materials are inexpensive and are among the most widely used materials for MFC anode fabrication. Carbon cloth, carbon fiber, carbon felt, or carbon brush is commonly

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used as base anode electrodes. These materials, in comparison with nanostructured electrodes, offer limited surface area (Erbay et al. 2015). Carbon nanotubes (CNTs) have been greatly studied in MFC systems in the last years. Many works have shown that CNT can greatly improve anode performance due to their high conductivity and high specific surface area, by increasing the transference rate of electrons produced by bacterial metabolism. CNTs are usually employed as contribute materials for anode modification, both as raw CNTs and doped with other chemical elements. In the last case, doped CNTs can ever offer higher power densities in comparison with non-doped nanotubes. For instance, multiwall (MW) carbon nanotubes have been employed for the modification of electrodes made of carbon cloth and carbon paper, increasing the performance of the systems. The methods for the modification of carbon anodes with CNTs can be simple, e.g., submersion of the electrode material in a solution containing CNTs. For example, Sun et al. (2010a, b) synthesized carbon nanotubes (in the presence of polyelectrolyte polyethyleneimine, PEI) for the construction of carbon paper electrodes. When comparing the bare anode based on carbon paper and the anodes including the CNTs, the modified anodes are capable of producing up to 20% more of power output. This increase in power output have also been observed when using CNTs for the modification of carbon cloth electrodes in comparison with bare carbon cloth electrodes, both in terms of power density and coulombic efficiency (Tsai et al. 2009). CNTs-textile composites have been studied as promising electrode material in MFCs (Xie et al. 2011). Carbon textile fibers modified with CNTs provide high performance because of the resulting porous structure, with a 3-D network formed that enhances substrate transference and efficient biofilm development due to high surface area availability. Specifically, Xie et al. (2011) observed an increase of 68% in power density as compared to conventional carbon cloth anodes. The higher performance of CNT-textile anodes also leads to significantly higher current densities (158%) and energy recovery rates. Nevertheless, when employing CNTs as anode materials, the biocompatibility of these materials with microorganism present in the anodic chamber need to be comprised, since some works have reported that CTNs can pose several toxicity issues, e.g., limiting cellular proliferation (Yu et al. 2018). As aforementioned, the doping of CNTs with chemical elements can improve the performance provided by non-doped carbon nanotubes. Research works on nitrogen-doped CNTs (Ci et al. 2012) have shown that the addition of nitrogen greatly contributes to enhancing the efficiency of the anode electrode by creating large number of actives sites and improving biocompatibility with bacteria populations. As study case, a maximum power density of 1.04 W
m2 was reported by using N-doped CNTs in dual-chamber MFCs by Ci et al. (2012), while the power densities for non-doped CTNs and bare carbon cloth anodes in the same systems reached only 0.71 and 0.47 W
m2, respectively. Conducting polymers can also be combined with carbon nanomaterials to enhance electron conductive and stability. These types of polymers include polyaniline, polyethylenimine, and polypyrrole, among others. Polyaniline or PANI has been specially studied for the modification of anode electrodes in MFC systems in combination with CNTs (PANI-CNT composites). In such case, the

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positive charges of PANI can interact with the negative charges of the bacterial membrane increasing the attachment of the biofilm formed to the anode electrode and thus the electron transfer rate. For example, Qiao et al. (2007) developed PANI/ CNTs composites for anode modification, obtaining higher efficiencies in MFC systems versus devices equipped with plain PANI anodes. The combination of polyaniline and carbon nanotubes provides larger specific surface areas. Similarly, and more recently, polypyrrole-CNTs composites were employed in carbon felt anodes (Roh and Woo 2015), offering 38% more power (287 mW
m2) versus non-modified carbon felt anodes. In this case, the polypyrrole-CNTs composites were further treated with acid (mixture of nitric and sulfuric acid) and organic germanium (Ge-132). As seen, the nanomaterials commented until now have been tested on several support materials such as carbon paper, felt, or textile fibers. According to Erbay et al. (2015), the orders of magnitude of the surface areas in these types of composites are generally lower than 1 m2
g1. In contrast, 3-D nanostructured electrodes offers high surface areas and boost microbial electron transfer. In this regard, 3-D porous CNT-based sponges have been recently developed for the direct construction of MFC anodes (Erbay et al. 2015). These materials, which are formed by the interconnection of carbon nanotubes, are capable of producing power densities of up to 2150 W
m3 (anode volume). Moreover, the method reported in the bibliography for the fabrication of the said nanostructured sponge-type anodes is relatively simple and offer low costs. In the last years, graphene has attracted growing attention as advanced anode material in MFC devices. In addition to the outstanding properties of graphene in terms of electrical conductivity and chemical stability, the use of graphene prevents possible toxicity issues posed by other materials such as carbon nanotubes. Graphene can be synthesized in different arrangement types (graphene sheets, graphene aerogel, composites, etc.). Among the first attempts to apply graphene as electrocatalyst in the anode electrode is the work by Zhang et al. (2011), in which a stainless steel mesh modified with graphene is analyzed as anode material. The results show that MFCs working with graphene-based anodes are capable of producing power densities of 2668 mW
m2. This level of power output is significantly higher than the power density obtained with anode electrodes lacking graphene (18-fold higher). Other relevant works reported in the bibliography include the study of graphene aerogel doped with nitrogen (Yang et al. 2016a, b). This type of material presents a (hierarchical) porous structure facilitating the bacterial cell diffusion and electron transfers. This material provides significant power densities in milliliter-scaled MFCs, with maximum values of 750 W
m3 (referred to anode volume). More recently, 3-D porous graphene aerogels prepared by hydrothermal reduction methods has been reported as enhanced electrodes with high specific capacitance values (>3600 F
m2) and improved stability in terms of electricity generation. The maximum power density obtained with this anode type is in the order of 2400 mW
m3 (anode volume). Graphene has also been combined with other metallic materials, for instance, nickel foam. Wang et al. (2013) developed 3-D graphene oxide-nickel nanomaterials

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consisting of the deposition of graphene oxide onto nickel foam. In this case, the material structure can be analyzed from both macro and micro perspectives. While nanostructure provides large surface area promoting bacterial growth and electron transference, this type of anode material also offers a microporous structure that boosts mass diffusion. This anode material has shown to be effective for MFC systems, offering power densities of 660 W
m3 (anode volume material). This power density is higher in comparison with that obtained with plan nickel foam electrodes and other common carbon electrodes such as carbon cloth. Graphene oxide has also been recently studied for the degradation of dyes in the anode chamber of MFCs. Dyes are xenobiotic compounds frequently present in wastewater effluents that are regarded as persistent and toxic pollutants. Enhanced anode materials based on graphene oxide have shown to be effective at removing this type of contaminants with simultaneous power generation (Khalid et al. 2018). Other material options include graphene-based sponges. As in the case of CNTs, MFC anodes based on graphene sponges were developed and studied by Xie et al. (2012). Apart from the graphene sponges, the anodes comprised stainless steel as material for current collectors. This type of anode material can be considered low cost, offering great advantages such as enhanced conductivity. Another important group of materials for anode construction are metal nanostructured materials. For example, several works have studied the influence of gold nanoparticles as an effective way to increase MFC performance (Alatraktchi et al. 2014; Guo et al. 2012; Sun et al. 2010a). Au nanoparticles are used to decorate the anode electrode, showing that higher current outputs can be obtained when employed along with other material substrate (e.g., carbon paper). Alatraktchi et al. (2014) deposited Au nanoparticles (with particle size from 60 to 100 nm), obtaining almost twofold power density comparing with carbon papers lacking Au. Although gold nanoparticles could enhance MFC performance, the high price of this noble metal make it unfeasible the spreading of its use for practical implementations, and thus alternative materials based on transition metals could be preferable to noble materials. Other innovative and noble nanostructure materials employed in MFC anodes are bio-palladium nanoparticles reported by Quan et al. (2015). These Pd nanoparticles are synthesized by microorganisms (Shewanella oneidensis) from a solution of Palladium(II), with the advantage of less amounts of chemicals needed for the nanoparticle synthesis compared with conventional methods and improved biocompatibility materials. When decorating carbon cloth electrodes (support material) with bio-Pd nanoparticles, the obtained power output maximum power and coulombic efficiency are, respectively, improved by 14 and 31% in comparison with bare carbon cloth anodes. Nano spinel materials such as Ni-ferrite (NiFe2O4) are an example of low cost nanostructure materials that can be employed for MFC anode decorating (Peng et al. 2017). This non-noble material acts as promotor of the transference between the bacteria and the anode electrode. By the addition of a low amount of Ni-ferrite (5%), an increase in power density of almost 30% higher is obtained in comparison with non-decorated carbon electrodes.

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Other transition metal-type nanomaterials that have drawn interest for anode fabrication in MFC are TiO2 nanotubes (Feng et al. 2016; Wen et al. 2013). TiO2 is a well-known photocatalyst employed in electrochemical energy systems. Although this compound can pose some biocompatibility issues, several authors have developed methods to successfully apply these TiO2 nanotubes in MFC anodes (Feng et al. 2016). With this type of nanostructure material, high current densities can be obtained in the order of 12 A
m2, which is up to 190-fold higher values in comparison with non-nanostructured titanium electrode and also significantly higher than in the case of using bare carbon felt-type anodes. The TiO2 nanotubes also showed to provide resistance against corrosion and high conductivity.

15.4

MFC Membrane Composites Including Nanomaterials

Proton exchange membranes (PEM) play an important role in MFC performance. They have to selectively transport protons from the anode to the cathode and simultaneously avoid both the transfer of substrate from the anode to the cathode and oxygen crossover from the cathode to the anode. Commercial membranes such as Ultrex or Nafion have been widely used as separator in MFCs, Nafion being the most common. However, these materials show some limitations such as their high cost and in the case of Nafion, both oxygen and substrate transfer, cation transport and biofouling among others. Due to these drawbacks, in recent years big efforts have been made in order to synthesize alternative materials to overcome the limitation of the commercial membranes. Polymeric/inorganic nanoparticle-based membranes have gained much attention in recent years due to their promising applications in biomedicine, environmental science, and energy production. The presence of nanoparticles in membranes generates preferential permeation paths, which improve the separation process. Moreover, they avoid the permeation of undesired species in addition to increasing thermal and mechanical stability. Therefore, the presence of nanoparticles in the membrane structure modifies the chemical and physical properties of the separator (Jadav and Singh 2009; Mahreni et al. 2009; Pan et al. 2010).

15.4.1 PES-Based Membranes Poly(ether)sulfone is one of the polymers reported in literature, which has been used as supporting material for nanoparticle-type membranes. In 2012, Rahimnejad et al. (2012) synthesized a novel type of nanocomposite membranes based on Fe3O4 nanoparticles. Their work focuses on the effect of different amounts of nanoparticle/ polymer on the performance of a double chamber MFC using Saccharomyces cerevisiae as anodic biocatalyst. The system was fed with glucose and neutral red as mediator. Membranes with 10, 15, and 20% of Fe3O4 nanoparticles were synthesized

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by casting method. The iron-based nanoparticles were synthetized by using FeCl2, FeCl3, and sodium hydroxide solutions as precursors. Atomic force microscopy (AFM) images show that the higher the amount of Fe3O4 nanoparticles, the larger the porous size. However, for amounts of Fe3O4 nanoparticles higher than 15%, the porous size is not uniform and the roughness of the surface is very high due to the appearance of some aggregations. Membranes containing 15% of Fe3O4 nanoparticles allow MFC to increase the power output by 29% over the value obtained by using a Nafion 117-based membrane (15.4 mW
m2). MFCs working with membranes prepared with the highest content of nanoparticles (20%) show lower value of power output than those with membranes containing only 10% of nanoparticles. The results demonstrate that the unique properties of ferric nanoparticles such as their conductivity, magnetism, or high catalytic activity, clearly favor the performance of MFCs. More recently, melt extrusion technique was employed for synthesizing Fe3O4 nanoparticles-type membranes using PES as polymeric matrix. Di Palma et al. (2018) prepared Fe3O4 nanoparticles by co-precipitating Fe+2 and Fe+3 using NH4OH as precipitating agent. They used a microextruder in order to obtain nanocomposite sheets of PES and two different amounts of magnetite nanoparticles, 5 and 20%. The membranes were tested as separators in a double chamber MFC fed with a synthetic solution containing sodium acetate as carbon source. The results reported in this work are in line with those obtained by Rahimnejad et al. (2012). The roughness of the nanocomposite membranes increases as the amount of nanoparticles also increases, being maximum for PES-based membranes containing 20% of magnetite nanoparticles (1.215 μm). Regarding the MFC performance, nanocomposite membranes containing 20% of Fe3O4 allow MFCs to reach a maximum power output of 9.59 mW
m2, which is similar to that obtained by using the commercial membrane CMI-7000 (12.58 mW
m2). The results show that the higher the presence of magnetite nanoparticles in the membranes, the higher the MFC performance in terms of power output. Regarding the wastewater treatment capacity and coulombic efficiency, MFCs working with this type of membranes exhibit a total organic carbon (TOC) removal of 75% and a coulombic efficiency of 11.36%. Among the material synthesized, nanocomposite membranes containing 20% of magnetite seem to exhibit a balance between electrochemical and chemical performance, being suitable for both bioenergy production and wastewater treatment.

15.4.2 Sulfonated Poly Ether Ether Ketone-Based Membranes Other polymer commonly used to prepare nanocomposite membranes is sulfonated poly ether ether ketone (SPEEK). Prabhu and Sangeetha (2014) reported the preparation of Fe2O4 nanoparticles-type membranes using SPEEK as polymeric matrix. The sulfonated polymer was obtained by mixing sulfuric acid and PEEK, and the polymer membranes were prepared by casting method dissolving an appropriate

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amount of SPEEK into N-methyl pyrrolidone. On the other hand, Fe2O4 nanoparticles were synthetized from FeCl3 and FeCl2, as previously commented, and then added to the casting solution. The effect of the amount of magnetite nanoparticles on the membrane properties and the performance of MFCs was investigated. Membranes with a content of 2.5, 5, 7.5, and 10% of iron nanoparticles were used as separator in a single chamber MFCs. E. coli (DH5-α) was used as active bacteria in the anode and glucose as carbon source whereas carbon cloth loaded with platinum as catalytic cathode for the oxygen reduction reaction. The results show that the roughness of the membranes increases as the amount of magnetite nanoparticles also increase. On the other hand, the higher the amount of iron nanoparticles, the lower the oxygen transfer rate for amounts of nanoparticles fixed between 2.5 and 7.5%. However, for amounts of nanoparticles beyond 7.5%, the oxygen transfer increases due to the space in the internal structure. These results are in line with those previously described for PES-based nanocomposite membranes. Regarding the MFC performance, SPEEK-based membranes containing 7.5% of iron nanoparticles allow MFCs to reach a maximum power output of 104 mW
m2, which is higher than SPEEK-based membranes without nanoparticles and also higher than Nafion 117 membranes. A similar trend is observed for the columbic efficiency, being maximum for membranes containing 7.5% of iron nanoparticles (78%). The roughness of the 7.5% based membranes is enough to allow a thin layer of biofilm to develop over the surface of the membrane avoiding the oxygen transfer from the cathode to the anode and increasing the MFC performance. However, despite the high roughness of the 10% membrane, the space in the internal membrane structure negatively affects the MFC performance. Regarding the ion transport, iron nanocomposite membranes exhibit a lower selectivity to cations such as sodium, calcium, potassium, magnesium, and ammonium than Nafion-based membranes, which favor the proton selectivity. These results show that SPEEK/ Fe3O4 nanocomposite membranes are a suitable alternative to commercial membranes as separator in MFCs. Besides magnetite nanoparticles, membranes based on SPEEK have been doped with montmorillonite nanoparticles (MMT). Hasani-Sadrabadi et al. (2014) studied the effect of both the sulfonation degree of SPEEK and the amount of montmorillonite nanoparticles on the performance of single chamber MFCs. Although the maximum ion exchange capacity belongs to a sulfonation degree of 89%, this sample was water soluble. Thus, it was determined that 82% of sulfonation degree is the maximum value which allows the sample to be mechanically and hydrolytically stable. Regarding the liquid uptake and the proton conductivity, both parameters increase as the sulfonated degree also increases. However, the results show that the higher the amount of sulfonate groups, the higher the oxygen transfer rate because the sulfonate groups might affect the crystallinity of the SPEEK structure. The oxygen molecules can diffuse across the ionomeric nanochannels formed, which increase as the sulfonated group also increases. Nevertheless, the oxygen permeability showed by the sample with different sulfonated degrees is lower than that exhibited by Nafion 117. The author considers the selectivity of the membranes as the ratio of proton conductivity to oxygen permeability. According to the sulfonated

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degree, the selectivity of the membrane increases as the number of sulfonate group also increases, being maximum for a sulfonated degree of 70%. Therefore, this was the polymeric matrix selected in order to prepare the nanocomposite membranes based on montmorillonite. In terms of selectivity, membranes with 3% of nanoclay exhibited the highest value. Therefore, SPEEK-based membranes with 70% of sulfonated degree and 3% of nanoclay were tested as separator in single chamber MFCs. In this case, E. coli Top10F was used as bacteria in the anodic chamber fed with glucose, electrodes were made of carbon cloth, and the cathode catalyst consisted of a blend of platinum and carbon black. The results show that the nanocomposite membrane with 70% of SPEEK and 3% of nanoclay allows MFCs to reach about 40% more power than the commercial membrane Nafion 117. The high power output displayed in MFCs, the low cost materials, the simple synthesis process, as well as the low oxygen transfer make SPEEK-based nanoclay membranes suitable alternatives to commercial MFC separators. Sulfonated silica has also been used as nanomaterial in order to improve the properties of SPEEK-based membranes. Sivasankaran and Sangeetha (2015a) studied the effect of the addition of SiO2 and sulfonated SiO2 nanoparticles to SPEEKtype membranes on their proton conductivity as well as their application as separator in MFCs. They added different amounts of sulfonated silica (2.5, 5, 7.5, and 10%) into the SPEEK matrix and the results were compared to those obtained with SPEEK/SiO2-type membranes as well as with Nafion 115. The membranes synthetized were characterized before being used as separator in a single chamber air cathode MFC fed with domestic wastewater in batch mode. Nyquist plots from electrochemical impedance spectroscopy (EIS) showed that the resistance of the membranes decreases as the amount of sulfonated silica also increases, being minimum for membranes containing 7.5%. This might be due to the low conductivity of SiO2 particles, whose presence increases the resistance of the membrane and reduces the conductivity. However, the addition of a more conductive compound such as SiO2-SO3H reduces the resistance of the membrane, and consequently increases its conductivity. All membranes containing sulfonated silica exhibit higher conductivity than both SPEEK-SiO2 and SPEEK membranes, being maximum for membranes containing 7.5% of SiO2-SO3H (1.018 S
cm1). Amounts of sulfonated silica beyond 7.5% increases the resistances of the membrane due to the agglomeration of the sulfonated silica over the SPEEK matrix, reducing the conductivity. Regarding the use of these membranes as MFC separator, membranes prepared with 7.5% of SiO2-SO3H allow MFC to reach the maximum power output (1008 mW
m2), even higher than Nafion 115 (320 mW
m2). The columbic efficiency is about 90%, also higher than the rest of materials tested. This work reports that the addition of SiO2-SO3H to SPEEK-based membranes increases the conductivity, and therefore improves their performance as MFC separator. In comparison to commercial membranes such as Nafion 115, membranes containing 7.5% of SiO2-SO3H exhibit similar conductivity and water uptake; however, the oxygen mass transport is an order of magnitude lower for sulfonated silica SPEEK membranes, which enhances the power performance of MFCs.

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15.4.3 Polyvinyl Derived-Based Membranes Polyvinylidene fluoride (PVDF) has commonly been used as polymeric matrix for different kinds of membranes due to its thermal and chemical stability as well as its biocompatibility. This material has shown promising results in different research fields such as separation processes, bioenergy production in direct methanol fuel cell or for biological applications. Considering all these advantages, in 2014 Shahgaldi et al. (2014) employed electrospinning method in order to produce PVDF nanofibers. The aim of their work is to synthesize PDVF nanofiber/Nafion-based membranes to be used as separator in double chamber MFCs, which use baker’s yeast as biocatalyst in the anode and glucose as fuel. Moreover, they studied the effect of different combination of PDVF nanofibers and Nafion (10 wt%/0.2 g, 18 wt%/0.4 g and 29 wt%/ 0.6 g) on MFC performance. The results showed that the membranes containing 0.4 g of Nafion allow MFCs to reach the maximum power output (4.9 mW
m2) while MFCs working with membranes prepared with 0.6 g of Nafion exhibit a similar power performance to commercial Nafion 117. This can be explained by the higher proton conductivity of membranes containing 0.4 g of Nafion. Regarding the columbic efficiency, this membrane also exhibits the highest value (12.1%). On the other hand, membranes prepared with a proportion of PVDF nanofiber/Nafion of 29 wt%/ 0.6 g offer both higher power output and higher columbic efficiency than membranes containing 10 wt%/0.2 g of PVDF nanofiber/Nafion. These results allow them to conclude that the amount of PVDF nanofiber is a key factor in Nafion composite membranes. In terms of wastewater treatment capacity, all system studied allow to remove more than 70% of chemical oxygen demand (COD), being all of them suitable for treating wastewater. More recently, polyvinyl alcohol (PVA) and sulfonated styrene (SS) were crosslinked in order to be used as polymeric matrix for nanocomposite membranes. In 2017, Rudra et al. (2017) incorporated different amounts of graphite oxide (0.2 wt%, 0.4 wt% and 0.6 wt%) into a polymer matrix based on PVA and SS for synthesizing nanocomposite materials to be used as separator in single chamber MFCs. The system worked with Lysinibacillus species as active bacteria in the anode fed with synthetic wastewater. In this work, the hydrophilicity of PVA is reduced by its crosslinking with SS as well as the addition of GO. Moreover, the presence of GO reduces the oxygen transfer through the membranes. Regarding the resistance, the Nyquist plots showed that among all the materials synthesized, nanocomposite membranes containing 0.4 wt% of GO exhibit the minimum resistance (5.4 Ω), and therefore the highest conductivity. In terms of power output, MFCs working with this type of nanocomposite membrane are capable of reaching a maximum power of 194 mW
m2, which is higher than the values achieved by the rest of materials tested. In terms of wastewater treatment capacity, all materials synthesized allow MFCs to reach COD removal rates higher than 70% after 25 day working, being maximum for those containing 0.2% of GO (88.97%). These results show the potential application of this low cost nanocomposite material as separator in MFCs for simultaneously producing bioenergy and treating wastewater.

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15.4.4 Sulfonated Polystyrene-Ethylene-ButylenePolystyrene-Based Membranes Sulfonated polystyrene-ethylene-butylene-polystyrene (SPSEBS) has also been employed for synthesizing nanocomposite-type membranes for MFCs. Sivasankaran and Sangeetha (2015b) added nanoparticles of sulfonated TiO2 ranged between 2.5 and 10% to a SPSEBS matrix. The materials elaborated were tested in a single chamber air cathode MFCs working with wastewater as bacteria inoculum and glucose as carbon source. The ionic conductivity of the samples increases as the content of sulfonated TiO2 also increases, being maximum for membranes prepared with 7.5% of titanium nanoparticles (3.35 meq
g1). These results might be due to the increase of the acid sites caused by the presence of sulfonated TiO2. The same trend is observed in terms of proton conductivity, being maximum for nanocomposite membranes prepared with 7.5% of sulfonated TiO2 (3.57  102 S
cm1). However, the oxygen mass transfer decreases as the content in titaniumbased nanoparticles increases, which benefits their use in MFCs. Regarding their application as MFC separator, the maximum power output of 1345 mW
m2 was achieved by using nanocomposite membranes containing 7.5% of sulfonated TiO2, higher than using the rest of membranes prepared and four times higher than using Nafion 117. In respect of columbic efficiency, MFCs working with 7.5% sulfonated TiO2-type membranes achieve a value of 87%, higher than using the rest of materials. Then, the same polymeric matrix was dopped with a range of sulfonated silica nanoparticles from 2.5 wt% up to 10 wt% by the same author (Sivasankaran and Sangeetha 2015b). They tested the materials synthesized as separator in single chamber air cathode MFCs inoculated with wastewater and fed with glucose as main carbon source. The work reports that the higher the amount of sulfonated silica, the higher the ion exchange capacity of the membranes. The maximum value is observed for membranes containing 7.5% of sulfonated silica (3.015 meq
g1). For higher amounts of silica nanoparticles, this parameter decreases due to the agglomeration of the nanoparticles. Regarding the internal resistance, it shows the opposite trend, decreasing as the amount of sulfonated silica increases in the membrane structure. The minimum resistance is exhibited by the membranes containing 7.5% of sulfonated silica nanoparticle (37 Ω), and therefore the maximum conductivity (0.321 S
cm1). In respect of their application as MFC separator, SSEBS-based membranes containing 7.5% of sulfonated silica allow MFC to reach a maximum power output of 1209 mW
m2 and a columbic efficiency of 85%, both values higher than using the rest of membranes. Moreover, the authors report that the selected membranes offer results much better than other composite membranes achieving power densities more than 10 times higher.

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15.4.5 Nafion-Based Membranes Nafion solution has also been tested as supporting matrix for nanocomposite membranes. Ghasemi et al. (2012) compared the use of commercial membranes Nafion 112 and Nafion 117 with both carbon nanofiber/Nafion and activated carbon nanofiber/Nafion as separator in MFCs. They use Nafion solution in order to avoid the agglomeration of carbon nanofiber (CNF) and the activated carbon nanofiber (ACNF). The new membranes were tested in a double chamber MFC fed with glucose as carbon source. The roughness of the membrane decreases as the pore size decreases. For both parameters, the membranes studied show the following trend: Nafion 112 > Nafion 117 > CNF/Nafion > ACNF/Nafion. Small pore sizes reduce the oxygen transfer through the membrane, which benefits the MFC performance. Regarding the porosity, this factor increases as the pore size decreases, being maximum for ACNF/Nafion (47.6%). In terms of power generation, both materials synthesized allow MFCs to reach higher values of power output than commercial membranes Nafion 112 and Nafion 117. Among the carbon-based nanocomposite membranes, ACNF/Nafion are able to produce a maximum power output of 57.64 mW
m2 when they are used as separator in MFCs. Concerning the wastewater treatment capacity, all membranes synthesized allow to reach a maximum COD removal higher than 70%, being maximum for ACNF/Nafion membranes. These results confirm that both types of carbon fiber-based nanocomposite membranes are suitable alternative to commercial membranes for bioenergy production and wastewater treatment in MFCs. A few years later, Bazrgar Bajestani and Abbas Mousavi (2016) also employed Nafion solution in order to elaborate TiO2-based nanocomposite membranes. The authors studied the effect of different solvents such as dimethylformamide (DMF), dimethylacetamide (DMAc), and N-methylpyrrolidone (NMP) on the behavior of the membranes synthesized as MFC separator. In this case, the MFC assembly selected was a double chamber system fed with sludge containing DMF and molasses. Regarding the characterization, TiO2 nanoparticles-based membranes containing DMF show the highest conductivity (12.6 mS
cm1), up to threefold higher than commercial membrane Nafion 112. This membrane also exhibits the highest ion exchange capacity (1.32 meq
g1). In terms of MFC performance, TiO2 nanocomposite membranes prepared with DMF as casting solvent allow these devices to reach an open circuit voltage of 330 mV, higher than the rest of materials tested. These results show that the casting solvent play an important role on the performance of TiO2/Nafion nanocomposite membranes as separator in MFCs.

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Nanomaterials for MFC Cathode Electrodes

At the cathode, oxygen reduction reaction (ORR) can generally follow two different pathways, 4-electron and 2-electron (peroxide) pathways (Mustakeem 2015). The 4-electron pathway implies that oxygen is reduced directly to water and is preferred over the 2-electron pathway, which includes the production of hydrogen peroxide, since in 4-electron pathway, double quantity of electrons is transferred. Generally, the 4-electron pathway appears to be predominant on noble metal catalysts while the peroxide pathway is more common on carbon-based electrodes. In any case, specific reaction mechanisms depend on pH conditions as seen in Table 15.2 (Gajda et al. 2015; Kinoshita 1988; Sawant et al. 2016). As indicated in Table 15.1, the acidic 2-electron pathway involves peroxide as intermediate. In acidic conditions, for both 4- and 2- electron pathways, H2O is a final product (Santoro et al. 2017). In neutral/basic conditions, the produced OH can accumulate at the sites of the catalysts lowering kinetic performance (Sawant et al. 2016). As indicated by Sawant et al. (2016), a combination of two- and fourelectron pathways can even occur in specific types of doped carbon materials employed as catalysts (e.g., N-doped carbon) (Yue et al. 2015). A wide range of nanostructured materials have been investigated for MFC cathodes, including nanomaterial composites based on metal and carbon compounds. As in the case of anode fabrication, carbon nanotubes (CNTs) have been widely studied as ORR catalyst in the presence and in the absence of additional materials. CNTs have also been investigated in the presence of platinum in order to minimize the amount required of this noble catalysts. In this case, CNTs act as cathode support (Ghasemi et al. 2013), showing that PT/CNTs can increase power output up to 32% in comparison with Pt-coated cathodes in the absence of CNTs, while the amount of platinum can be lowered by 25% respect to the conventional amount required (0.5 mg/cm2 cathode area) to maintain the cathode efficiency. In other works (Sanchez et al. 2010), the use of CNTs is aimed to increasing the efficiency of platinum-based cathodes by increasing the total surface area and the ratio of surface/volume. Table 15.2 Reaction pathways for the ORR Acidic conditions 4-electron pathway 2-electron pathways

Neural/Alkaline conditions 4-electron pathway 2-electron pathways

Reactions

Potentials (E)

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

E ¼ 1.229 V E ¼ 0.695 V E ¼ 1.770 V

O2 þ 2H2O þ 4e ! 2OH O2 þ H2O þ 2e ! HO2 þ OH HO2 þ H2O þ 2e ! 3OH 2 HO2 ! 2OH þ O2

E ¼ 0.401 V E ¼ 0.065 V E ¼ 0.867 V

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Noble nanomaterials have been employed in cathode MFC systems. An et al. (2011) analyzed the effect of including silver nanoparticles (AgNP) to coat graphite cathodes and compared their performance with graphite cathodes covered with platinum. The comparison of the performance of respective AgNP and Pt-coated cathodes showed that the cathode based on silver nanoparticles could offer the highest currents (from to 0.04 to 0.12 mA). The different materials were assessed in membraneless systems, and in the case of the use of silver nanoparticles, the materials also posed several limitations related to biofilm growth (biofilm inhibition). This last issue could be overcome by the use of MFC separators systems. More recently, goldpalladium (Au-Pd) nanoparticles with core-shell have been investigated. The nanoparticles consisted of Au cores and thin Pd shells (Yang et al. 2016a, b). This type of cathode can achieve power outputs of generates a. 16 W m3 (volume anode), which in turns showed to be more than double the power obtained with hollow structured-based platinum cathodes (7.1 W m3). Materials with core-shell structures can provide high catalytic properties because of the presence of lattice strain between the shell and the core domains. Despite the good performance of precious catalysts for the ORR, they pose limitations related to high costs and low abundance in nature (Bullock 2017). Due to their lower price, transition metal-based catalysts are of high interest for their application in bioelectrochemical systems (BES) and therefore also in microbial fuel cells. Cobalt-based catalysts have shown promising results as cathode materials to increase the rate of the oxygen reduction reaction. Cheng et al. (2006) tested cobalt tetramethylphenylporphyrin (CoTMPP) in air-cathode single-chamber MFCs obtaining comparable performance to platinum (12% lower in the case of the macro-cyclic compound CoTMPP). Other macrocyclic complexes of cobalt and iron have also proposed to catalyze the ORR. For example, the addition of cobalt oxide into iron phthalocyanine was tested by Ahmed et al. (2012), showing that these material can offer significant power outputs in MFCs (654  32 mW
m2). Yuan et al. (2011) synthesized iron phthalocyanine (FePc) supported onto aminofunctionalized multiwalled CNTs for the construction of MFCs cathodes, which showed high efficiency for the ORR. These materials even enhanced the performance of Pt-based cathodes, offering a power output of up to 601 mW
m2. Manganese oxide (MnO2) has drawn much attention in the last years as low cost and promising catalysts in MFC systems. In terms of nanostructured materials, several works have studied MnO2 nanoparticles prepared by relatively simple methods offering good results. For example, hybrid manganese oxide nanostructures have been reported by Haoran et al. (2014) synthesized by a hydrothermal route. This type of nanomaterial offers significant efficiency towards ORR catalysis, allowing the oxygen reduction reaction to be performed by the four-electron pathway in basic solution. Other interesting works have combined nanostructured MnO2 and graphene as cathode material. Gnana kumar et al. (2014) investigated nanotubular MnO2/GO achieving higher power densities in comparison with cathodes only based on higher neat manganese oxides and nanotubes and nano-rods, respectively.

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Other transition metal-based nanostructured oxide catalysts displaying spinel structure have been successfully studied in MFC systems (Hu et al. 2015; Mahmoud et al. 2011; Ortiz-Martínez et al. 2016). Spinel oxides are compounds that have the general formula AB2O4, in which A and B are metal cations. Depending on the composition, these compounds can display normal or inverse, tetrahedral and cubic structures. The cations are placed in octahedral and tetrahedral sites, with the possibility of one metal cation type displaying multiple valences in the same structure. In terms of catalytic activity towards the ORR, cubic phase spinel oxides offer higher efficiencies in comparison with tetragonal-structured oxides due to a stronger surface binding capacity with oxygen. Cobalt-based spinel oxides haven drawn much attention due to their low cost and the presence of multiple valences of Co cation in the spinel structure. The addition of manganese into cobalt oxide spinels can even offer higher efficiencies. Several spinel compositions with formula MnxCu1-xCo2O4 supported onto carbon were studied by Hu et al. (2015). These compounds were prepared by a relatively simple hydrothermal method. Some of the phases investigated offered performances comparable to Pt (supported onto carbon) as MFC catalysts. Other authors (Ortiz-Martínez et al. 2016) analyzed several cobalt spinel oxides doped with Cu and Ni, synthesized by thermal decomposition at different metal atomic ratios. The synthesized compounds were assessed both in power output densities and removal of COD from real wastewater (industrial effluents) in air-cathode single-chamber systems. Among the phases studied, the Cu-Co cobalt oxide with chemical formula (Cu0.30Co0.70)Co2O4 achieved a significant power density of approximately 570 mW
m2 (87% out of the power generated by platinum). In terms of wastewater treatment, spinel MFCs working with spinel oxidecoated cathodes achieved COD removal values around 56% after 240 h of operation.

15.6

Conclusions

Traditional wastewater treatment technologies still show some limitations related to the energy requirements and the yield of the process. Microbial fuel cells (MFCs) have demonstrated to overcome these drawbacks, becoming a potential alternative to conventional wastewater treatment methods. MFC is a sustainable technology, which mitigates two of the most urgent challenges for human being: global warming and water scarcity. However, despite MFCs are able to simultaneously treat wastewater and produce bioenergy, there are some limitations which difficult their scaledup application. In recent years, nanotechnology has gained attention due to their benefits in different research fields such as membrane technology for energy or separation processes as well as environmentally friendly materials for biomedicine, among others. This chapter addresses the application of nanotechnology in order to improve the wastewater treatment capacity of MFCs. Despite further works are needed for facilitating the MFC commercialization, it has been demonstrated that the use of nanomaterials in anode, cathode, or membrane allows MFC performance to be improved.

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Chapter 16

Nanocomposite Materials Based on TiO2/ Clay for Wastewater Treatment Soulaima Chkirida, Nadia Zari, Abou El Kacem Qaiss, and Rachid Bouhfid

Contents 16.1 16.2 16.3 16.4

Introduction: Bionanocomposites . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Clays . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Clays in Nanocomposites . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Application of Bionanocomposite in Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.4.1 Photocatalysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.4.2 Photocatalysis: Mode of Action and General Mechanism . . . . . . . . . . . . . . . . . . . . . 16.4.3 Dioxide of Titanium TiO2 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.5 Nanomaterials Based on TiO2 for the Water Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.6 Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

16.1

363 365 366 368 371 371 373 376 377 377

Introduction: Bionanocomposites

In the recent past, there has been an increasing interest and tendency to hear the notions composites, nanocomposites, and bionanocomposites. In fact, they stem from the same basic idea; creating new, light, weight, and high performance materials to substitute the conventional non-biodegradable ones. Generally, a composite is determined as a combination of two or more organicinorganic materials with different physical and chemical properties and distinguishable interface. It is composed of two phases: a matrix, which is the continuous phase and a dispersed phase known as reinforced materials. Regarding the nanocomposites, they have come forth to overcome limitations of composites. They have the peculiarity of a nanoscale dimensions (10
9) in at least one of their phases, which leads to diverging S. Chkirida · N. Zari · A. El Kacem Qaiss · R. Bouhfid (*) Laboratory of Polymer Processing, Moroccan Foundation for Advanced Science, Innovation and Research (MAScIR), Institute of Nanomaterials and Nanotechnology (NANOTECH), Rabat, Morocco e-mail: r.bouhfi[email protected] © Springer Nature Switzerland AG 2019 R. Prasad, K. Thirugnanasambandham (eds.), Advanced Research in Nanosciences for Water Technology, Nanotechnology in the Life Sciences, https://doi.org/10.1007/978-3-030-02381-2_16

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characteristics of the composites because they contain a noticeable and an available number of interfaces for interactions between the intermixed phases (Roy et al. 1986). A bionanocomposite has been defined as a nanocomposite involving biopolymers in combination with nanoscale reinforcements to produce composites with improved properties such as biodegradability, biocompatibility, permeability, crystallinity, and thermal stability (Darder et al. 2007). Bionanocomposites is resulting from a combination of several sciences like biology, materials science, chemistry, engineering, and nanotechnology to generate green and eco-friendly materials from renewable resources that can be served instead of conventional materials in numerous applications without being harmful to the environment. The recent great attention and concern of bionanocomposites is related to several reasons. The first is the ability to design and create new flexible materials with the required physical properties. Secondly, the use of a low content of dispersed nanometric reinforcements makes it unnecessary to fully occupy the matrix with fillers to obtain high mechanical properties. The third reason is the employment of cheap and overflowing natural resources as polymers and reinforcement materials. Finally, these materials have a high specific surface area and this is due to the nanoscale size of the particles. Thus, the surface of the reinforcement is very high, so the interfacial region reinforcement/Matrix is even more significant (Schadler et al. 2007). The main biopolymers used in the bionanocomposite matrix are polymers of biological origin, such as polysaccharides (cellulose, chitin, chitosan, alginate, pectin, xanthan, etc.); (Shchipunov 2012) involving proteins (corn, soybean, caseinate, gelatin, whey, silk or wheat gluten); (Kirk-othmer 2007), nucleic acids (DNA, ribonucleic acid, deoxyribonucleic acid), as well as certain polyesters produced by microorganisms (Kirk-othmer 2007). The addition of reinforcements to biological polymers is being required and done because of the limitations of using these biopolymers alone. In fact, they are restricted by some deficient properties, such as low mechanical strength, high gas and water vapor permeability, low heat degradation temperature, etc. (Blackburn and Richard 2005). Indeed, reinforcements come from two main sources: organic or inorganic materials, and in different forms such as fibers, spheres, plates, whiskers, flakes, and leaves to modify and improve the properties of the composite, namely: conductivity, weight, surface properties, durability, visual appearance, shrinkage, etc. (Callister and Rethwisch 2007). The materials used as fillers are diverse and the most common are carbon nanotubes despite their prohibitive cost, followed by clays, metal nanoparticles, spherical silica, nanoclusters, cellulose nanofibrils, and fullerenes (Blackburn and Richard 2005). The following section will discuss in detail the use of clays as reinforcements and their amenities.

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16.2

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Clays

Because of their low cost, abundancy in most continents of the world, clay minerals are well known and recognizable from the earliest days of civilization. They are strong nominees as adsorbents due to their high adsorption properties and potential for ion-exchange (Nayak and Singh 2007). Their composition consists of colloidal fragments of primary silicates, and their structure consists of a gathering of tetrahedral and octahedral sheets of small cations, like aluminum or magnesium coordinated by oxygen atoms. Their classification is based on how these (T) tetrahedral and (O) octahedral sheets have been placed in a package of layers as T:O clays. For instance, 1:1 clays are formed by one tetrahedral and one octahedral group in each layer (Grim 1930). For a better understanding, the following table (Table 16.1) highlights the major clay mineral groups, by indicating their chemical structures composition. Due to the unequal size of interlayer cations with the holes of tetrahedral sheets, the presence of interlayer cations causes to an interlayer spacing. The layers stay near together with a regular gap between them. Among the items of clays, the gap between layers is called interlayer space, and it is caused by the presence of different hydrated cations with unequal size. The most common exchangeable cations are K+, Na+, Ca2+, Mg2+, and H+, and they can undergo exchange reactions with organic as well as inorganic cations (Konta 1995). Another property of clays is their capability to exchange ions; it can be measured by a special parameter known as the cation exchange capacity (CEC). This property is relying upon the nature of the isomorphous substitutions in the tetrahedral and octahedral layers, and hence on the nature of the soil where the clay belongs. This explains why the same type of clay from different origins show distinct CEC (Guggenheim 2001). In addition, clays are endowed with a specific layer space, called the basal spacing (d001), which is defined as the sum of one layer thickness and the interlayer distance (Fig. 16.1). In general, it is the thickness of repeated units in a regular multilayer Table 16.1 Classification and example of clay minerals Structure type 2:1 (TOT)

Group Smectite

Mineral examples Montmorillonite Hectorite Saponite

2:1 (TOT)

Illite

Illite

2:1 (TOT)

Vermiculite

Vermiculite

1:1 (TO)

Kaolinserpentine

Kaolinite, dickite, nacrite

Composition [(Al3.5–2.8Mg0.5–0.2)(Si8) O20(OH)4] [(Mg5.5–4.8Li0.5–1.2)(Si8) O20(OH)4] [(Mg6)(Si7.5–6.8Al0.5–1.2) O20(OH)4] [(Al4)(Si7.5–6.5Al0.5–1.5) O20(OH)4] [(Al4)(Si6.8–6.2Al1.2–1.8) O20(OH)4] Al4Si4O10(OH)8]

Basal spacing (Å) 12.4–17

10 9.3–14 7.14

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Fig. 16.1 Scheme of a 2:1 clay structure (smectite) indicating the basal spacing

structure, which can be measured or calculated from their X-ray diffraction patterns. The basal spacing depends also on the nature of clay, its swelling, and its hydration degree of its interlayer cations (Olad 2011).

16.3

Clays in Nanocomposites

Many efforts have been made for the elaboration of intercalated and exfoliated polymer/clay nanocomposites with improved properties, where the clay is the reinforcement. To create these nanocomposites, we start by fully separating all the clay particles into individual layers by breaking the bonds between them (intercalation), then afterwards all clay layers tend to disperse in a polymer (exfoliation) (Pavlidou and Papaspyrides 2008). This process is depicted in Fig. 16.2. Among the most important points in the preparation of a polymer/clay nanocomposite is the awareness of the degree of intercalation/exfoliation and its effect on the nanocomposite’s properties. For this purpose, an analysis of the microstructure of the prepared nanocomposite is required. Two techniques including transmission electron microscopy (TEM) and X-ray diffraction (XRD) analysis are commonly used to characterize the microstructure of nanocomposite as well as pure clay or pure organo-clay. Due to the regular layered structures of these latter, their analyses show a characteristic peak in XRD analysis indicating the d spacing in clay structure (Olad 2011).

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Fig. 16.2 Exfoliation of a clay in a polymer matrix

According to the literature, it has been deemed that montmorillonite, hectorite, and saponite are the most frequently used smectite type layered silicates. Similarly, Sepiolite and palygorskite are the recurrently used microfibrous clay minerals to render a vast range of advanced nanostructured materials and nanocomposites for the sake of their particular structural, morphological, and textural features (Ruiz-Hitzky et al. 2001). For instance, Montmorillonite has a high surface area in the range of 750 m2/g, a noticeable aptitude to exchange ions, and a very important surface reactivity which makes it very acceptable and usable in polymers (Hussain et al. 2006). It is a hydrated aluminosilicate and the structure consists of two silica tetrahedral sheets sandwiching an edge-shared octahedral sheet. Isomorphous substitutions of Si4+ for Al3+ in the tetrahedral sheet and of Al3+ for Mg2+ in the octahedral sheet cause an excess of negative charges within the MMT layers. These negative charges are balanced by some cations such as Ca2+ and Na+ located between the clay layers (HUSSAIN et al. 2006) (Fig. 16.3). While sepiolite exhibits a fibrous morphology, it is a microcrystalline hydrated magnesium silicate of unit cell formula Si12O30-Mg8(OH,F)4(H2O)4 composed with an alternation of blocks and tunnels that grow up in the form of fibers (Fig. 16.4) (Ruiz-Hitzky et al. 2001). Two layers of tetrahedral silica sandwiching a central magnesium oxide-hydroxide layer A constitute the blocks B where silanol groups (Si-OH) are presented on their boundaries on account of the discontinuity of the silica sheets. As well as the smectites clays, Sepiolite has been depicted as a clay mineral that is able to give polymer-fibrous clay nanocomposites (Bergaya and Lagaly 2006). Polymers not only relate with the superficial surface of this silicate, but they can also pass through the tunnels of the mineral (Inagaki et al. 1995).

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OH

OH OH

Exchangeable Cations nH2O

Oxygens and

OH

Hydroxyls

Aluminum, iron, magnesium

Silicon, occasionally aluminum

Fig. 16.3 Structure of montmorillonite

16.4

Application of Bionanocomposite in Wastewater

As indicated before, bionanocomposite can be used in several fields, but in this section, we will focus on their applications on wastewater and their ability to remove various organic inorganic pollutants and micropollutants from sewage. The prime cause behind that assorted types of bionanocomposites, which are currently being studied as a tool for water purification and decontamination, lies to their unique properties like their high surface area, adsorption and catalyst behaviors,

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Fig. 16.4 Schematic model representing the sepiolite structure

small size, mechanical properties, and their affinity for several pollutants and ions (Thomas and Stephen 2010). Before proceeding in discussing the efficiency of bionanocomposites, it is a crucial matter to hint at the performance of the clay alone in purifying water. For instance, the montmorillonite clay can be used as an adsorbent for the uptake of a wide variety of contaminants from water. Modified montmorillonite was investigated to remove several toxic pollutants in a batch reactor like phenols (2-chlorophenol, 3-cyanophenol, and 4-nitrophenol) (Lee et al. 1997a). Due to this modification, there was a change in the surface character of montmorillonite from hydrophilic to organophilic. Moreover, montmorillonite has been directly used for the removal of heavy metal cations (Abollino et al. 2003) and cationic dyes (Lee et al. 1997a). Besides, the efficiency of the removal of various radionuclides (Cs, Sr, Eu, Th) has been proved (Dent et al. 1992). Furthermore, Sepiolite has been also used in similar studies; it was observed that it was efficient in the uptake of phenols and lignin compounds from paper mill industry wastewater samples (Ugurlu et al. 2005). It was as well investigated in the removal of cobalt from aqueous solution (Kara et al. 2003). What is more to mention is that clay has been coupled up to some natural polymers to elaborate clay/biopolymer bionanocomposite that has exhibited superior

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efficiency. For instance, Montmorillonite was used coupled to chitosan to study the adsorption of dyes such as Congo red (Wang and Wang 2007), and for the removal of selenium from well water with an adsorption capacity of 18.5 mg/g. It should be noted that this bionanocomposite removes selenium to levels below the WHO (World Health Organization) standards. Hence, it was suggested to use the granulation of this bionanocomposite in future fixed bed in real-life application (Bleiman and Mishael 2010). Another study reveals that Sepiolite can also be used as a reinforcement to synthesize the sepiolite polydiallydiamethylammonium nanocomposite (Sp-PDADMAC) for the purification of wastewater. In fact, it was observed that a winery effluent with an amount of 1610 mg/L of total suspended solid (TSS) can be completely clarified with this nanocomposite in 2 min, knowing that PDADMAC alone has not improved such clarification (Rytwo 2012). In a succeeding study, they used sepiolite–PDADMAC, sepiolite–chitosan, and volclay–PDADMAC nanocomposites to treat winery wastewater in an efficient manner between 15 and 60 min. It was deduced that the main item of these nanocomposites is their aptitude of coagu-flocculation in the order of few minutes. As a result, the use of these materials would be an approach that can improve and speed up the purification of large volumes of wastewater by diminishing the long periods required for sedimentation process, and overcoming the need for extent sedimentation tanks used in conventional flocculation process (Rytwo et al. 2013). In addition to chitosan, another biopolymer has been widely used as a matrix, which is the alginate, to prepare the Montmorillonite-Alginate bionanocomposite. This latter has been shown effective for the uptake of heavy metals ions like Pb (II) from aqueous solutions, with an adsorption capacity of 238.1 mg/g. Whereas, in the case of raw wastewater, this bionanocomposite has completely purified it with a removal rate of 100% of Al(III), Cr(III), Fe(III), Mn(II), Ni(II), Pb(II), and Zn (II) (Shawky 2010). The abilities of bionanocomposite go beyond what was mentioned. In fact, it can also destroy a cluster of infectious bacteria such as E. coli, Pseudomonas aeruginosa, Staphylococcus aureus, and this is what was implemented by a study using a bionanocomposite based on a Clay-polydimethyloxane-chitosan-silver (Zhou et al. 2007). Similar finding was reported by another approach of elaboration of tuned chitosan/polyaniline/metal hybrid bionanocomposite for treatment of meat industry wastewater. It showed that this material possessed 100% antibacterial activity against Bacillus subtilis and Escherichia coli due to the metal-well exposable attitude beyond polyaniline-grafted modified chitosan interfaces (Mostafa and Darwish 2014). Another bionanocomposite based on the N-(2-hydroxyl) propyl-3trimethyl ammonium chitosan chloride (HTCC) in solution and suspension of cetyltrimethyl ammonium bromide modified montmorillonite was found strongly adept in inhibiting the proliferation of a large variety of microorganisms, as well as fungi, Gram positive bacteria, and Gram negative bacteria in whichever medium (weak acid, wake base, or water) (Wang et al. 2015).

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16.4.1 Photocatalysis Heterogeneous photocatalysis is defined as a discipline which incorporates a mixture of reactions: organic synthesis, mild or total oxidations, water dividing, photo reduction, transfer of hydrogen, metal deposition, water detoxification, disinfection, anticancer therapy, and removal of gaseous pollutants (Herrmann 1999). In fact, it appertains to the advanced oxidation processes (AOPs), that can be broadly defined as oxidation methods that have the peculiarity of exploiting the lofty reactivity of hydroxyl and other radicals to oxidize obstinate, non-biodegradable and toxic compounds and contaminants to water, carbon dioxide and inorganics or, at best, to innocuous products (Tarr 2003). The environmental functions of heterogeneous photocatalysis are diverse; passing by air depollution and soil remediation (Higarashi and Jardim 2002), to water and wastewater treatment (elimination of organic and inorganic hazardous waste) (Ahmed et al. 2011). In this light, the coming section assesses the potency of heterogeneous photocatalysis in the purification of wastewater and aqueous solutions. Nevertheless, before dwelling on the details of the wastewater treatment by photocatalysis, a brief glimpse on the general mechanism of this technique and the most employed semiconductors is in order.

16.4.2 Photocatalysis: Mode of Action and General Mechanism The term photocatalytic marks reactions boosted by irradiation and done through semiconductors (SC) that can take many forms (from powders nanoparticles, colloids to dissolved molecules or ions). As mentioned before, photocatalysis cannot occur without a photocatalyst, which is a metal semiconductor characterized by an electronic band structure in which the highest occupied energy band, named valence band (VB), and the lowest-occupied energy band, termed conduction band (CB), are disconnected by a band gap, a region where no electron states can be present. The band gap energy (Eg) is the energy difference between the CB bottom and VB top (Hoffmann et al. 1995). When a photon with an energy of hν equivalent or larger than the bandgap energy Eg of the semiconductor, the initiation of the photocatalytic reaction will take place and get executed by a promotion of a photo excited electron from the valance band (VB) to the conduction band (CB) (The mechanism is summed up in the figure below) (Fig. 16.5). Thusly, a hole and an electron are generated on the (VB) and the (CB), respectively, forming a pair (h+/e
), this recombination leads to the dissipation of the input energy as heat, and the charge will carry over these electron–hole pairs and adsorbed species (reactants) on the semiconductor surface, then photo-oxidation take effect (Ravelli et al. 2009).

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Fig. 16.5 General mechanism of the photocatalysis (Robert and Malato 2002)

In this process, after the photo generation of the electron–hole pairs on the surface of the catalyst, the hole is captured by the water molecules leading to the formation of HO˙ radicals and H+, while the electrons grant the formation of H2O2 which break down in more OH˙ radicals as a result of its reaction with the oxygen provided by the medium. Finally, the radicals thus formed oxidize the pollutants by producing intermediates and end products. The following chain reactions demonstrate the overall process: SC þ hv ! hþ þ e


ð16:1Þ




ð16:2Þ

O2ads þ e ! O2

˙


Hþ þ H2 Oads ! HO˙ ads þ Hþ ads

ð16:3Þ

O2 ˙
þ H ! HO2 ˙

ð16:4Þ

˙

HO 2 ads ! H2 O2 ads þ O2

ð16:5Þ

H2 O2ads ! 2HO˙ ads

ð16:6Þ

HO˙ þ Pollutants ! Intermediates ! CO2 þ H2 O

ð16:7Þ

They are numerous semiconductors that have been tested for their efficiencies over against pollutants for several decades, and this is due to their stability in various conditions, biocompatibility, and ability to create charge carriers when activated with appropriate supply of light energy (Djurišić et al. 2014). We cite V2O5, ZnO, WO3, CdS, ZrO2, SnO2, CeO2, and TiO2. In this light, the following section will spotlight and assess the potency of this latter (TiO2), its structure, synthesis techniques, and applications.

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Fig. 16.6 Representation of the TiO2 Anatase, Rutile, and Brookite forms

16.4.3 Dioxide of Titanium TiO2 Based on the researches on the literature, the heterogeneous photocatalytic by dioxide of titanium has earned more attention. It is among the most auspicious and bright materials in photocatalytic applications because it exhibits several appealing peculiarities, being a strong oxidizing, a high photo stable, a redox selective agent, and a low cost. Not to mention its ability of light absorption and its arrangement of an electronic structure (Ohtani et al. 1997).

16.4.3.1

Structure

Titanium dioxide owns several distinct allotropic forms with different structures. The most well-known ones that have been most applied in photocatalysis studies are Anatase and Rutile structures which are conform to a tetragonal crystalline system, and Brookite structure that corresponds to an orthorhombic crystalline system (Fig. 16.6) (Dambournet et al. 2010). The Anatase, Rutile, and Brookite are naturally occurring phases. The difference between them lies on the inordinately difficult of the brookite’s synthesis, in the time that Anatase and rutile may be synthesized simply in the laboratory without any struggle (Dambournet et al. 2010). According to the several researches on photocatalysis efficiency, it has been clearly indicated and made unquestionable that Anatase is more active in photocatalysis than the other phases (Sclafani and Herrmann 1996). This emphasis in photo activity is apparently attributable to the efficiency of Anatase to create electron–hole pairs. Additionally, the difference in activities may refer to the difference in the structure and the degree of hydroxylation of the adsorbing surface, knowing that the increase of density of superficial OH
groups is on the same wavelength with the increase of photocatalytic activity of TiO2 (Kobayakawa et al. 1990).

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A remarkable item is that Anatase is thermodynamically unstable; it is only detectable relatively at temperature of 700  C phase, and above this temperature, there will be a transformation and a conversion to the Rutile phase, the more stable phase (Bickley et al. 1991). Regarding the TiO2 morphologies, several methods have been quoted on the literature, chiefly, mesoporous structures (Yi et al. 2001), Nano rods (Wu and Qi 2007), nanowires (Wen et al. 2005), and nanotubes (Varghese et al. 2003).

16.4.3.2

Methods of Synthesis

There are multiple methods available for the preparation of TiO2. In the following paragraph, these methods shall be briefly mentioned, giving emphasis to the sol gel method. As matter of fact, the numerous benefits emanate from the sol-gel preparation of TiO2 that have driven so many researchers to use it. It actually provides high pure nanosized crystallized powder at relatively low temperature, the possibility of preparing composite materials, the opportunity of controlling the process stoichiometrically and it contributes to the production of homogeneous materials (Akpan and Hameed 2010). The TiO2 can be prepared using continuous reaction (Kim and Kim 2002), dip coating (Patil et al. 2003), chemical solvent and chemical vapor decomposition (CSD, CVD) (Babelon et al. 1998; Kim et al. 2004), supercritical carbon dioxide (Wu et al. 2008), precipitation (Li and Demopoulos 2008), linking plasma treatment to inverse micelle (Arimitsu et al. 2007), hydrothermal and solvothermal reaction (Zhu et al. 2006; Li and Demopoulos 2008), ultrasonic irradiation (Peng et al. 2008), and multi-gelation method (Inoue et al. 2004).

Sol-gel Method Sol-gel method is one of the well-renowned approaches of synthesis of metal oxide, nanoparticles and mixed oxide composites. For a better understanding of the main principle of this process, a clear distinguishing between sol and gel is needed: Sol: is defined as a stable suspension of colloidal particles or polymers in a solvent. Gel: is a three-dimensional continuous network holding a liquid phase that can be constructed by an agglomeration of colloidal particles, as it can be built from linking polymers chains. In most cases, the gel process is irreversible when the system is governed by covalent interactions, whereas the gelation action can be reversible if other interactions are involved (Miao et al. 2004). A typical sol-gel process is done in few steps: hydrolysis, condensation, and drying process. Initially, a precursor that comes usually in a form of an inorganic metal salts or a metal alkoxide is subjected to a rapid hydrolysis to generate a metal hydroxide solution pursued by a condensation, which promotes the formation of three-dimensional networks: gels.

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Fig. 16.7 The process of sol-gel method

The switch from the sol to the gel is accomplished by a loss of the solvent and a complete polymerization (Rao et al. 2017) (Fig. 16.7). Afterward, the humid gel becomes solid through a heat treatment and a drying process. Based on this latter, different structures of the resulting solid product are obtained; if the gel is dehydrated by evaporation, then it will be collapsing and shrinking, then a xerogel is formed. Whereas the gel structure may be maintained with large pores and low density if the mode of drying is performed under supercritical conditions, and in this case an aerogel is formed (see equations below) (Miao et al. 2004). The following chain reactions display the process:   Hydrolysis : MðORÞx n þ H2 O ! R
OH þ M
O
H Condensation : 2M
OH ! M
O
M þ H2 O

ð16:8Þ ð16:9Þ

The reason behind commonly using metal alkoxides as precursors metal in the sol-gel process is the high reaction affinity of alkoxides toward water (Bradley 2001). The most commonly used precursors metal for the preparation of TiO2 nanoparticles are titanium tetraisopropoxide (TTIP) (Devi et al. 2014), titaniumbutoxide (TBT) (Devi et al. 2014), and titaniumtetrachloride (TiCl4) (Zhu et al. 2000).

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Doping TiO2

It is admitted that TiO2 exhibits under ultraviolet light (λ < 387 nm) high reactivity and chemical stability. However, it will be more suitable to enlarge the absorption spectrum of TiO2 into the visible field (λ > 400 nm) to enhance its surface and photocatalytic properties and make it convenient in solar energy applications. In this light, several approaches have been suggested for the development of dopedTiO2 by implanting different materials such as transition metals (Ag, Pt, Cu, Cr, Co, Ni, Mn, Mo, Nb, V, Fe, Au) (Anpo 2000), non-metal elements (F,N, S, C, B, P, I) (Ohno et al. 2003), semiconductor having lower band gap energy (Hirai et al. 2001).

16.5

Nanomaterials Based on TiO2 for the Water Treatment

Titanium dioxide has been used extensively as a photocatalytic material in the purification of water. It has been found unduly efficient for the mineralization of nearly all organic compounds and for the inactivation of pathogenic microorganisms present in contaminated water (Lee et al. 1997b). Yet, using bare TiO2 in photocatalytic presents some restrictions. For instance, the tendency of TiO2 particles to agglomerate in a suspension causes a loss in its effective surface area along with its catalytic efficiency (Bhattacharyya et al. 2004). Furthermore, it displays unimportant adsorption ability because of its nonporous property (Torimoto et al. 1997). Hence, recently, many efforts have been devoted to immobilize fine TiO2 on porous adsorbent materials such as zeolites (Sampath et al. 1994), activated carbon (El-Sheikh et al. 2004), silica (Lepore et al. 1996), and clays (Bineesh et al. 2011). Another reason behind supporting TiO2 nanoparticles on adsorbent is the higher specific surface area created along with the effective adsorption sites generated which are meaningful compared to those in the case of bare TiO2 (Anderson and Bard 1995). Among the TiO2-based materials used in water purification, we cite, integrated water treatment materials known as a photocatalytic membranes where TiO2 has been incorporated in various membrane matrixes to provide photocatalytic activities in purifying water (Mills and Le Hunte 1997). What is more, a composite based on TiO2 and graphene oxide has been used for the removal of Zn2+, Cd2+, and Pb2+ ions from water (Lee and Yang 2012). Another composite based on ZnO and TiO2 was prepared for photocatalytic inactivation of Escherichia coli (Sethi and Sakthivel 2017). Similar finding was reported by another study that confirms the degradation of various organic molecules under UV irradiation by a material based on TiO2 supplied with different loads of MgO (Bandara et al. 2004), and by a composite based on Cu2O/TiO2 that have also been demonstrated for photocatalytic degradation of organic compounds (Senevirathna et al. 2005). Furthermore, a doped TiO2 material with boron and nickel oxide Ni2O3/TiO2-xBx has shown an important

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degradation reactivity of these compounds, about 80% COD value and 70% of total organic carbon (TOC) were achieved, and about 80% of the total chloride content is converted into Cl
ions after illumination for 4 h, indicating that this TiO2-based material had been not only degraded but also mineralized efficiently under visible light irradiation (Zhao et al. 2004). For dye’s degradation, many TiO2-based composites have been studied, for instance, a core shell composite based on TiO2-SrO core shell nanowires have recently been arranged (Wang et al. 2013). Another composite based on zirconia incorporated titanium nanoparticles TiO2/ZrO2 composites have been tested for Selective photo-degradation of dyes like Rhodamine B.

16.6

Conclusion

Nanotechnology for water and wastewater treatment is recognized as one of the most compelling technologies of the twenty-first century. It enfolds a broad range of tools, techniques, and applications because of the unique physicochemical and surface properties of nanomaterials and their convergence with current treatment technologies. Even that many nanotechnologies cited in this chapter are still in the research and development phases and in the laboratory stage, they will have their way to pilot tasting or even commercializing in the coming years because of their considerable and undeniable potential in purifying water. Yet, nanomaterials exhibit some restrictions toward a few pollutants. Hence, an outstanding idea that would enhance all the processes of purifying crops up and is displayed in coupling the potential of nanotechnology to another magnificent process which is photocatalysis. This latter has an ability to completely mineralize wide spectrum of pollutants.

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Rytwo G, Lavi R, Rytwo Y et al (2013) Clarification of olive mill and winery wastewater by means of clay-polymer nanocomposites. Sci Total Environ 442:134–142 Sampath S, Uchida H, Yoneyama H (1994) Photocatalytic degradation of gaseous pyridine over zeolite-supported titanium dioxide. J Catal 149:189–194 Schadler LS, Brinson LC, Sawyer WG (2007) Polymer nanocomposites: a small part of the story. Kluwer Academic Publishers/Plenum Publishers, Boston, MA Sclafani A, Herrmann JM (1996) Comparison of the photoelectronic and photocatalytic activities of various anatase and rutile forms of titania in pure liquid organic phases and in aqueous solutions. J Phys Chem 100:13655–13661 Senevirathna MKI, Pitigala PKDDP, Tennakone K (2005) Water photoreduction with Cu2O quantum dots on TiO2 nano-particles. J Photochem Photobiol A 171:257–259 Sethi D, Sakthivel R (2017) ZnO/TiO2 composites for photocatalytic inactivation of Escherichia coli. J Photochem Photobiol B 168:117–123 Shawky HA (2010) Improvement of water quality using alginate/montmorillonite composite beads. Polym Polym Compos 21:449–456 Shchipunov Y (2012) Bionanocomposites: green sustainable materials for the near future. Pure and Applied Chemistry 84(12):2579–2607. https://doi.org/10.1351/PAC-CON-12-05-04 Tarr MA (2003) Chemical degradation methods for wastes and pollutants: environmental and industrial applications. M. Dekker, New York Thomas S, Stephen R (2010) Rubber nanocomposites: preparation, properties, and applications. John Wiley & Sons, Singapore Torimoto T, Okawa Y, Takeda N, Yoneyama H (1997) Effect of activated carbon content in TiO2loaded activated carbon on photodegradation behaviors of dichloromethane. J Photochem Photobiol A 103:153–157 Ugurlu M, Gurses A, Yalcin M, Dogar C (2005) Removal of phenolic and lignin compounds from bleached Kraft mill effluent by Fly ash and Sepiolite. Adsorption 11:87–97 Varghese OK, Gong D, Paulose M et al (2003) Extreme changes in the electrical resistance of Titania nanotubes with hydrogen exposure. Adv Mater 15:624–627 Wang X, Du Y, Jianhong Yang YTJL (2015) Preparation, characterization, and antimicrobial activity of quaternized chitosan/organic montmorillonite nanocomposites. Clin Exp Rheumatol 33:97–103 Wang L, Wang A (2007) Adsorption characteristics of Congo red onto the chitosan/montmorillonite nanocomposite. J Hazard Mater 147:979–985 Wang W, Yang J, Gong Y, Hong H (2013) Tunable synthesis of TiO2/SrO core/shell nanowire arrays with enhanced photocatalytic activity. Mater Res Bull 48:21–24 Wen B, Liu C, Liu Y (2005) Depositional characteristics of metal coating on single-crystal TiO2 nanowires. J Phys Chem B 109:12372–12375 Wu CI, Huang JW, Wen YL et al (2008) Preparation of TiO2 nanoparticles by supercritical carbon dioxide. Mater Lett 62:1923–1926 Wu JM, Qi B (2007) Low-temperature growth of a nitrogen-doped titania nanoflower film and its ability to assist photodegradation of rhodamine B in water. J Phys Chem C 111:666–673 Yi GR, Moon JH, Yang SM (2001) Ordered macroporous particles by colloidal templating. Chem Mater 13:2613–2618 Zhao W, Ma W, Chen C et al (2004) Efficient degradation of toxic organic pollutants with Ni2O3/ TiO2-xBx under visible irradiation. J Am Chem Soc 126:4782–4783 Zhou N, Liu Y, Li L et al (2007) A new nanocomposite biomedical material of polymer/Clay-CtsAg nanocomposites. Curr Appl Phys 7:58–62 Zhu J, Deng Z, Chen F et al (2006) Hydrothermal doping method for preparation of Cr3+-TiO2 photocatalysts with concentration gradient distribution of Cr3+. Appl Catal B 62:329–335 Zhu Y, Zhang L, Gao C, Cao L (2000) The synthesis of nanosized TiO2 powder using a sol-gel method with TiCl4 as a precursor. J Mater Sci 35:4049–4054

Chapter 17

Nanotechnology: The Technology for Efficient, Economic, and Ecological Treatment of Contaminated Water S. Vijayakumar and M. Priya

Contents 17.1 17.2 17.3 17.4

Background . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . History of Water Filtration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Filter Media for Water Filtration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Major Contaminants in Drinking Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.4.1 Pesticides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.4.2 Chlorination . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.4.3 Fluoride . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.4.4 Arsenic and Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.4.5 Lead . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.5 Nanotechnology for Removing Drinking Water Contaminants . . . . . . . . . . . . . . . . . . . . . . . . 17.6 Antimicrobial Action of Nanomaterial and Its Control Over Drinking Water . . . . . . . . . 17.7 Water and Society: Past, Present, and Future . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.8 Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

17.1

381 383 386 387 387 387 387 387 388 388 397 399 399 400

Background

Utilization of safe drinking water has improved a lot for the last decades in every part of the world, but around one billion people still lack to have drinking water. Water has a healthy influence on all aspects of human life which includes food, health, economy, and energy. The need of fresh water is necessary for the safety of children and almost all people (Theron and Cloete 2002; Eshelby 2007). The statistical report

S. Vijayakumar (*) · M. Priya Department of Science and Humanities, Sri Ramakrishna Institute of Technology, Coimbatore, Tamil Nadu, India © Springer Nature Switzerland AG 2019 R. Prasad, K. Thirugnanasambandham (eds.), Advanced Research in Nanosciences for Water Technology, Nanotechnology in the Life Sciences, https://doi.org/10.1007/978-3-030-02381-2_17

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says that an estimated number of 200 million people die every year owing to water scarcity and drinking contaminated water (Leonard et al. 2003). About 5000–6000 children die due to the water-related diseases like diarrhoea (Ashbolt 2004; Hutton et al. 2007). More than one billion people around the world are not able to access safe water (WHO 2013). Within another couple of decades the existing water bodies will shrink to one-third of its quantity ascribable to growing population. The stable fresh water flow was around 2500–15,000 km3 per year (Lvovich 1979; Postel et al. 1996), in which 4000 km3 per year is fresh water and used for industry, irrigation, and domestic purposes (Gleick 1993). This utility of water will increase to 4300–5000 km3 in 2025 (Shiklomanov 2000; Rockstrom 2003; Seckler et al. 2003). There is no possibility to increase the quantity of water due the increasing population and climatic changes (Orosmarty et al. 2000). According to UN projection, the worldwide population will increases by about 2.9 billion before 2050. One of the key roles for the demand of water supply is due to the result of exploitation of water resources for domestic and industrial purposes (Shannon et al. 2008). Another cause of the reduction of fresh water supply is due to the elevated emission of pollutant in the surface and ground water (Kemper 2004; Foley et al. 2005; Coetser et al. 2007). The discharge of endocrine disrupting compounds (EDCs) into the surface water affects the living organisms. The existing and conventional water treatment processes are not able to remove the wide range of toxins, chemicals, and pathogenic microorganisms in wastewater as expected. This is the time to have the latest technology like nanotechnology for the purification of water. Applying good purification technologies only gives good water for the long time with low operating cost. Nanotechnology has been considered as a relevant and effective technique to solve water-related problems (Bottero et al. 2006). Water is absolutely necessary for life on earth. All living organisms require water for their creation, survival, and evolution. Increased disposal of industrial, domestic, and agricultural waste into water bodies and consequent pollution of water has become a major issue in twenty-first century. Water, being a universal solvent, dissolves all kinds of contaminants like metal ions, inorganic anions, organic dyes, pesticides and insecticides, and radioactive elements. Therefore, great challenges lie in the removal of contaminants from drinking water at ultra-trace levels at affordable cost. Different techniques such as adsorption, membrane filtration, ion exchange, and capacitive deionization have evolved in the recent past. Among these technologies, adsorption has proved to be economical and efficient in removing targeted pollution from water in large scale. Nanomaterial due to large surface area, easy surface modification, and selective catalytic activity turned out to be an effective adsorbent for water purification. Nanomaterial such as carbon nanotubes, titanium dioxide, zinc oxide, gold, and silver are contributing to the development of more efficient treatment processes in the advanced water treatment systems (Obare and Meyer 2004).

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17.2

383

History of Water Filtration

Dating back to 2000 B.C. people had the practice of gravel filtration, heating and straining water might purify it. The aim was to have better tasting drinking water because they could not be able to distinguish clean and foul water. In earlier treatment, turbidity was a driving force, they don’t know about microorganisms or chemical contaminants. During 1500 B.C. Egyptians (Fig. 17.1) first discovered the principle of coagulation and this details was found on the tomb of Amenophis II and Ramses II, The practice of sieving water was invented by Hippocrates and he prepared the first bag filter called as Hippocratic Sleeve during 500 B.C. (Table 17.1). Rome built its first aqueducts during 250 B.C. (Fig. 17.2). Most of the aqueducts were in the underground to prevent pollution. Many of the Roman techniques are seen in modern day sewers and water transport systems. Archimedes invented a mechanical system called as water screw that pump water from lower water bodies to

Fig. 17.1 (a) Planned water bodies of Harappa town of Lothal. About 4000 years ago, this city was a thriving port and was counted among the principal centers of the Indus Valley Civilization (Courtesy: India Perspectives). (b) On the walls of the tombs of Egyptian rulers Amenophis II and Rameses II, which date back to the fifteenth and thirteenth century B.C., respectively, there are pictures of a water clarifying apparatus. (c) Mohenjo-Daro has been called the “city of wells.” The number of wells found in the excavated areas of the city may have had over 700 wells (Kenoyer 1948)

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Table 17.1 Important milestones in the history of water purification (2000 B.C.–2014) Year 2000 B.C.

1500 B.C. 500 B.C. 250 B.C. Ninth century

Seventeenth century 1627 1804 1810 1852 1879 1902 1906 1908 1914 1916 1935 1948 1959 1962 1965 1974 1975 1994 1997 1998 2000 2003 2004 2007

Milestone Boiling of water over fire, heating of water under the sun, dipping of heated iron into water, filtration through gravel and sand, as well as the use of the Strychnos potatorum seed (Sus’ruta Samhita) Principle of coagulation Hippocratic sleeve Aqueducts in Rome Moses and the Israelites found that the water in Marah was bitter. Following instructions from God, he cast a tree into the water, and the water was immediately sweetened (Bible) Italian physician Lucas Antonius Portius provided details of a multiple sand filtration method (Soldier’s Vade Mecum) Seawater desalination was experimented by Sir Francis bacon Setup of world’s first city-wide municipal water treatment plant (Scotland, sand filter technology) Discovery of chlorine as a disinfectant (Humphrey Davy) Formulation of Metropolis Water Act (England) Formulation of Germ Theory (Louis Pasteur) Use of Chlorine as disinfectant in drinking water supply (calcium hypo chlorite, Belgium) Use of ozone as disinfectant (France) Use of chlorine as disinfectant in municipal supply, New Jersey Federal regulation of drinking water quality (USPHS) Use of UV treatment in municipal supplies Discovery of synthetic ion exchange resin (Adams, Holmes) Nobel Prize to Paul Hermann Müller (insecticidal properties of DDT) Discovery of synthetic reverse osmosis membrane (Yuster, Loeb, Sourirajan) Publishing of Silent Spring, first report on harmful effects of DDT (Rachel Carson) World’s first commercial RO plant launched Reports on carcinogenic by-products of disinfection with chlorine Formulation of Safe Drinking Water Act (USEPA) Development of carbon block for drinking water purification Report on use Zerovalent Iron for degradation of halogenated organics (Gillham, Hannesin) Report on use Zerovalent Iron nanoparticles for degradation of halogenated organics (Wang, Zhang) Drinking Water Directive applied in EU Adoption of Millennium Declaration during the UN Millennium Summit (UN Millennium Development Goals) Report on use Noble metal nanoparticles for degradation of pesticides (Nair, Tom, Pradeep) Stockholm Convention, banning the use of persistent organic pollutants Launch of world’s first nanotechnology based domestic water purifier (Pradeep, Eureka Forbes Limited) (continued)

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385

Table 17.1 (continued) Year 2009 2014 2014

Milestone Water purifier originally targeting low income families in India, this water purifier uses rice husk ash and nano silver (Tata Swach) Multi-functional water filtration membrane. Uses titanium dioxide nanotechnology Photocatalytic Water Purification Technology photocatalysts and the UV rays from sunlight to detoxify polluted water at high speeds.

Credits to several governments, organizations, individuals and Dr. T. Pradeep, Indian Institute of Technology Madras

Fig. 17.2 (a) Rome’s first aqueduct was built in 312 B.C. and supplied a water fountain at the city’s cattle market (Gargarin and Fantham 2010). (b) An Archimedes’ screw, also known by the name the Archimedean screw, is a machine historically used for transferring water from a low-lying body of water into irrigation ditches (Kantert 2008)

higher land. This mechanism of water screw was a basis for modern day industrial pumps. In 1627, seawater desalination was experimented by Sir Francis bacon who passed the way for further development for many scientists. Regulations are focused on industrial water contamination and how water treatment plants are constructed. Techniques such as flocculation, active and aeration carbon adsorption were followed. Also, membrane developments for reverse osmosis have been developed. In wastewater treatment, nanotechnology plays a vital role; nanomembranes, nanoparticles, nanowires, carbon nanotubes, and grapheme are used with the

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purpose of softening the water and removal of contaminants. Nanotechnology is preferred for better water purification or treatment. Nanotechnology holds great promise in remediation, desalination, filtration, purification, and adsorption in water purification process.

17.3

Filter Media for Water Filtration

The components for mechanical media are inert in the sense that they do not interact with water or the materials or objects to be removed. Mechanical filtration removes sludge, waste material, or dust. Replacement of filtering media is only required when the media can no longer release all the dirt that it holds. Many different size and form of parasites are created when the media is arranged for filtration. Coarse media is easier to clean and reuse than filter media. The pore size is about 1/90 of the diameter of human hair in the finest mechanical media but it has the disadvantage that it easily clogs and may not be reusable. Biological filtering media (Fig. 17.3) able to break down dissolved solids to a very less toxic form and provide housing for beneficial bacteria. It is advisable to place the mechanical filtering media before the biological media so that the bacteria get its food and oxygen. Chemical media can remove one or more impurity by means of absorbents. These chemical and biological media helps to maintain water quality as unwanted matters adhere to it. The most widely used chemical media are activated carbon and resins.

Fig. 17.3 Different kinds of water pollutant such as microbes (Gilbert and Neufeld 2014), Lead (Courtesy: Boliden, Sweden), and pesticides (Courtesy: nutrilabs.gr). Water filtering media such as natural sand, fine gravel, activated carbon, and ion exchange resin (Gilbert and Neufeld 2014)

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17.4

387

Major Contaminants in Drinking Water

17.4.1 Pesticides Pesticides are chemical composites that are used to destroy and control pest in agriculture. This includes insecticides, nematocides, herbicides, fungicides, and rodenticides. Further, it seriously threatens the long term survival of ecosystem by disruption of predator prey relationships and loss of biodiversity. It is difficult to separate the ecological and human health effects of pesticides from those of industrial compounds that are intentionally or accidently released into the environment.

17.4.2 Chlorination The disinfection of drinking water by chlorination has in recent years come under closer scrutiny due to its potential hazards accomplished with the production of stable chlorinated organic chemicals. Environmental protection Agency is authorized to collect the data of public health effects resulting from the consumption of contaminated drinking water.

17.4.3 Fluoride Fluoride is an essential element for animals and humans. For a healthy life a minimum of at least 1 mg per kg of body weight is required. The primary studies establish that the fluoride via drinking water affects bones and teeth.

17.4.4 Arsenic and Mercury Arsenic is a natural component of the earth’s crust and it is highly toxic in its inorganic form (Table 17.2). The inorganic arsenic in drinking water is more responsible for carcinogen. In an extreme case of high intake of arsenic primarily leads to skin cancer, and also causes cancer in bladder and lungs. Neurological and renal disturbances are caused by mercury poisoning, and the ingestion of acute toxic dose causes cardiovascular collapse, shock, severe renal failure, and severe gastrointestinal damage.

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Table 17.2 Evaluation of various nano adsorbents for arsenic removal

Nanoadsorbent CeO2-CNT Magnetic nanoparticle PEG-MWCNTs Zirconium oxide nanoparticle Hydrous titanium dioxide α- Fe2O3 nanoparticle Iron doped TiCl2

Particle size – 20.0 17.4

pH 7.0 6.5 4.0

Temperature (0  C) – 25 25

Sorption capacity (mg/g) As As (III) (V) – 81.9 8.0 8.8 – 13.0

10.8

7.0

–

5.2

6.0

4.8

7.0

25

83

–

5.0 108

7.0 7.0

25 –

95 –

47 20.4

References Li et al. (2003a, b) Roy et al. (2013) Bhakat et al. (2006) Nabi and Aslam (2009) Liu et al. (2012) Lu and Liu (2006) Nabi and Aslam (2009)

17.4.5 Lead Lead is mostly considered as a number one health threat to children. The minimum dose of lead damage the nervous system, stunt a child’s growth, and cause learning disabilities. High concentration of lead in the body can cause brain and kidney damage, death or permanent damage to the central nervous system.

17.5

Nanotechnology for Removing Drinking Water Contaminants

It was estimated that over 1.1 billion people in the world lack adequate supply of water (WHO 2015). This is the reason why there is a rising cost of potable water, environment and climatic concerns and growing populations (Adeleye et al. 2016). A variety of organic and inorganic pollutants contaminate the fresh water resources (Schwarzenbach et al. 2006). To meet the current quality standards, there is a requirement of more efficient technologies to treat polluted water (Burkhard et al. 2000; Ferroudj et al. 2013). Even though we are having different technologies, the nanotechnology is a suitable remediation for water treatment (Sadegh et al. 2014). Quality drinking water is a need of today’s life. However, extensive use of chemicals as insecticides, herbicides, fungicides, and fertilizers as well as a growing industrialization has a profound influence in the quantity of pollutants in the drinking water via the aquatic system (Fig. 17.4). Day by day, the amount of potential hazardous chemicals are entering into our ecosystem. These pollutants are extremely carcinogenic, toxic, and persistent. Nanotechnology based treatment of water are efficient, economic, and ecologically friendly. The commercial suspension of

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Fig. 17.4 Leonardo da Vinci described water as “the vehicle of nature,” believing water to be to the world what blood is to our bodies. (a) Old Man with Water Studies. (b) Machine for raising water. (c) Machine for excavating canals (Courtesy: Sweet Briar College, USA)

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nanofer 25nZVI particles showed very efficient removal of selected organo chlorinated pesticides. The use of nZVI for the removal of synthetic organic substances is a promising activity due to its early application, efficiency, and high receptivity (Simkovic et al. 2015). Nanomaterials functionalized with various chemical groups show excellent removal of desired target compounds (Sundaram et al. 2008; Daifullah et al. 2007). It was reported that graphene has great adsorption capacities for pesticides. Further, dehalogenation and removal of persistent halocarbon pesticide from water was performed with graphene (Zarur and Ying 2000; Bakoyannakis et al. 2003). Graphene coated silica is a highly efficient sorbent used for the removal of residual organophosphorus pesticides from water (Li and Bai 2005). Nanocrystalline metal oxides not only adsorb the pollutant of pesticides but also destroy many chemical hazards by converting them into safer by-products (Paarimal et al. 2010). It is possible to remove pesticides by means of magnetic nanoparticles (Darmadi et al. 2008; Radushkevich 1949; Temkin and Pyzhev 1940; Hadi et al. 2010; Viswanathan et al. 2009). The multiple parameters of nanomaterial are considered for nanomaterial reactivity enhancement and its use in aqueous solution for removal of pesticides (Firozjaee et al. 2018). Oxide-based nanoparticles are extensively used for the hazardous pollutants removal from polluted water. These include titanium oxides (Gao et al. 2008), Titanium oxide/dendrimers composites (Barakat et al. 2013), and Zinc oxides (Tuzen and Soylak 2007). Iron-based nanoparticles are ecofriendly material which can be directly used on contaminated water bodies with less chance of secondary contamination (Li et al. 2003a, b). Surface modification of iron oxide nanoparticles are performed for the increase of adsorption (Ozmen et al. 2010; Palimi et al. 2014). The new functional group on the surface of carbon nanotubes increases the adsorption efficiency (Gupta et al. 2015). Gold nanoparticles over alumina has a capacity to eliminate Hg2+ from drinking water (Lisha et al. 2009) and the adsorption efficiency was about 10 times more than the previously reported adsorbents. The adsorption of toxic metals like mercury, lead, and cadmium from drinking water using monolayer protected silver nanoparticles was already reported (Bootharaju and Pradeep 2010). The immobilized graphene showed high Adsorption capacity for Rhodamine 6G (R6G), a rhodamine dye and chlorpyrites which was experimentally proved that the graphene adsorption of R6G in the commercial soft drinks leads to the decoloration of coco cola (Bootharaju and Pradeep 2012). Zerovalent silver nanoparticles were prepared in a greener way using Ficus tree, Ficus Benjamina leaf extract to remove the poisonous cadmium (II) element from the drinking water (Al-Qahtani 2017). Recent research investigated that graphite oxide successfully removed nearly 90% of heavy metal ions from aqueous solutions under optimum experimental conditions. Further silica/graphite oxide (2:3) composite was the most effective adsorbent for the heavy metal removal (Sheeta et al. 2014). Endemic fluorosis disease has become a major geo-environmental healthcare issue caused by fluoride ion. High-efficiency and low-cost materials to uptake fluoride from water have been a challenge for scientists and engineers. Bermuda grass as a starting source material converted into nanocomposite carbon fibers upon

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Table 17.3 Oxidants used to oxidize arsenite to arsenate, their operating condition and efficiencies

Oxidants Hypo chloride

Operating pH 7

Initial concentration mg/L 500

Monochloramine

8–12

50

Oxygen and ozone

7.6–8.5

46–62

Ground water

Chlorine dioxide

8–12

50

Ground water

Hydrogen per oxide

7.5–10.3

50

Chlorine

8.3

300

Fresh water and sea water Deionized water

Potassium permanganate

8–12

50

Type of water Ground water Ground water

Ground water

Remarks Complete oxidize After 18 h only 60% was oxidized Oxidation with ozone is very faster, complete oxidation was obtained. After 1 h 86% oxidation was obtained Efficiency increases with pH from 7.5 to 10.3 As (III) completely oxidized to As (V) by active chlorine Oxidation completed at 1 min

References Viet et al. (2003) Sorlini and Gialdini (2010) Kim and Nriagu (2000)

Sorlini and Gialdini (2010) Pettine et al. (1999) Hu et al. (2012)

Sorlini and Gialdini (2010)

heat treatment at 800  C for 1 h in Nitrogen atmosphere in the presence of metal oxides. Iron oxide-based nanocomposite (IBNC) is performing high (97%) removal of fluoride ion at a contact time of 60 min (pH 4) followed by titaniabased nanocomposite (TBNC) (92%) and micro carbon fiber (88%) respectively. Arsenicosis, or arsenic poisoning, results from entry of large amounts of arsenic in the body. This cumulative build-up of toxins causes serious health problems over the long run. Eventually, Arsenicosis leads to cancers of the liver, skin, lymphatic system, lungs, and urinary tract. More than 137 million people in 70 countries are currently suffering from arsenicosis, and a report recently predicted that in the next decade in Bangladesh alone, “One out of every ten adult deaths” will be a result of arsenicosis (Table 17.3). Gangrene of the extremities, known as “blackfoot disease,” has been reported with drinking arsenic contaminated well water. Nanotechnology (NT) is portrayed as an enabling technology package with endless application of which pollution and water scarcity do not escape the sphere of NT (Olvera et al. 2017). ZnO use in water and wastewater disinfection shows promise. It has the benefits of addressing the limitations of conventional water treatment methods (Asuncion et al. 2018). Water can be contaminated by natural or anthropogenic activities, such contaminant is fluoride (Pietrelli 2005). The source of high fluoride concentrations in ground water is primarily associated with addition

392

S. Vijayakumar and M. Priya

Table 17.4 Evaluation of different sorptive media used for arsenic removal

Adsorbent Activated rod mud Coconut shell carbon Chitosan resin Coconut pretreated with Fe (III) Coal based carbon Fly ash Sorghum biomass

Type of water Distilled water Distilled water Deionized water Distilled water Distilled water Distilled water Distilled water

Optimum pH 7.25

Sorption capacity (mg/L) As As (III) (V) 0.084 0.941

5.0

–

2.4

References Altundogan et al. (2000) Lorenzen et al. (1995)

6.0

4.45

–

Liu et al. (2012)

5.0

–

4.53

Lorenzen et al. (1995)

5.0

–

4.09

Lorenzen et al. (1995)

4

–

30

5.0

3.6

–

Diamadopoulos et al. (1993) Haque et al. (2007)

of fluoride by volcanic activities, high water-rock interaction, and low calcium concentration. Volcanoes represent the main natural persistent source of fluoride (Symonds et al. 1988). Fluoride has a narrow range between intakes that cause beneficial and detrimental health effects (WHO 2006). At low concentrations, it has beneficial effects on teeth; however excessive exposure to fluoride causes adverse health effects (CRC 2008). An intake of more than 6 mg fluoride per day will result in fluorosis (Jamode et al. 2004). Higher level of fluoride in ground water is a worldwide problem. Ethiopia is one of the countries in the world where a significant percentage of the population suffers from the consumption of fluoride rich drinking water. It has been reported that more than 11 million people in Ethiopian Rift Valley are potentially at risk of fluorosis (Haimanot et al. 1987; Nemade et al. 2002). The concentrations of fluoride in the Rift Valley communities, which are supplied from boreholes, are reported between 1 and 33 mg L 1 (Kloos and Haimanot 1999). Nano-AlOOH has considerable potential for the removal of excess fluoride from aqueous solution. High fluoride removal occurred between pH 6 to 8. The equilibrium data were tested to fit Freundlich, Langmuir, D-R, and Temkin isotherm models to reveal the understanding of the mechanism of fluoride adsorption at the surface of AlOOH (Table 17.4). The adsorption capacity and contact time required for maximum adsorption of fluoride (equilibrium time) with nano-AlOOH has shown better adsorption potential and performance (Adeno et al. 2014). Maximum (90%) fluoride removal was obtained with 0.6 g/L dosage of nanoMgO. Fluoride adsorption by nano MgO was found to be less sensitive to pH variations. Fluoride sorption was mainly influenced by the presence of OH ion. The presence of other ions studied did not affect the fluoride adsorption capacity of

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Nanotechnology: The Technology for Efficient, Economic, and Ecological. . .

393

Table 17.5 Efficiency of various nanomembrane for treatment of wastewater Technology Sodium- titanate Nanobelt membrane Non structured polymer based membrane ZrO2 microfiltration membrane

Nanofiltration Carbon Nanofiber membrane Nanomembrane prepared from casting r-alumina and titania nanocrystallites Nanofiltration with forward osmosis

Contaminant Removal of oil and radioactive Cs+ ions and Sr2+ Oil removal Pretreatment of dimethyl formamide (DMF) wastewater Remazol fiber roots dye Metal and metal oxide nanoparticles Microorganism and ions rejection from wastewater COD

Efficiency Cs+ 57.7% Sr2+ - 97.5%

References Wen et al. (2016)

99.75%

Ahmed et al. (2015)

Turbidity removal - 99.6% Suspended solids - 99.9%