Nanostructured Materials for Environmental Applications 3030720756, 9783030720759

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Subramanian Balakumar Valérie Keller M. V. Shankar  Editors

Nanostructured Materials for Environmental Applications

Nanostructured Materials for Environmental Applications

Subramanian Balakumar  •  Valérie Keller M. V. Shankar Editors

Nanostructured Materials for Environmental Applications

Editors Subramanian Balakumar NCNSNT University of Madras Chennai, India

Valérie Keller ICPEES University of Strasbourg Strasbourg Cedex 2, France

M. V. Shankar Department of Materials Science and Nanotech Yogi Vemana University Vemanapuram, Andhra Pradesh, India

ISBN 978-3-030-72075-9    ISBN 978-3-030-72076-6 (eBook) https://doi.org/10.1007/978-3-030-72076-6 © Springer Nature Switzerland AG 2021 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, expressed or implied, with respect to the material contained herein or for any errors or omissions that may have been made. The publisher remains neutral with regard to jurisdictional claims in published maps and institutional affiliations. This Springer imprint is published by the registered company Springer Nature Switzerland AG The registered company address is: Gewerbestrasse 11, 6330 Cham, Switzerland

Dedicated to Environmental Scientists

Foreword

Accelerated industrialization and urbanization in the last two centuries have caused colossal devastation of the pristine environment which we were blessed with. Attempt to re-establish and recover our habitat is a primal obligation of every individual who is a part of this biosphere. Above all, the responsibility is more with the research fraternity who can transform and revive the environment with the enormous scientific advances available. The field of nanomaterials has proven to exhibit refreshingly novel opportunities that can maneuver the catastrophic challenges that the environment is being subjected to. The unique scale which nanomaterials offer has rendered it the most sought-after approach in creating environmental sustainability. Pathbreaking advances in nanomaterial synthesis and nanostructure formulation have opened up newer and promising possibilities to re-create the lost glory. This book entitled Emerging Nanostructured Materials for Environmental Application edited by Prof. S. Balakumar and team offers an in-depth diverse range of emerging nanostructured materials that have been validated for environmental applications. Key attributes of this book are that the content is exquisite and comprehensible. It surpasses similar books of this kind as it is replete with exhaustive coverage on electrocatalysts, visible light/UV-based photocatalysts, magneto-­ photocatalysis, plasmonic nanostructured catalysts, electrocatalysts, thermo-­ catalysts, organo-catalysts, adsorption of nanostructured materials, chemical and surface kinetics, degradation of oil and water pollutants, and electromagnetic pollutants and shielding aspects. This book will serve as an indispensable source of reference to graduate/postgraduate students and researchers. In addition to placing emphasis on challenges related to latest advancements, it also elaborates on the trend for further research and future perspective that is needed for a multidisciplinary approach in environmental science with nanoscale multifunctional materials. Professor S. Gowri, B.E (TCE,MKU )., M.Tech (IITM )., Ph.D (IITM ) Vice-Chancellor University of Madras Chepauk, Chennai-600 005, India vii

Acknowledgments

First of all, we place our heartfelt thanks to Almighty, who has been kind enough to bless us with good health whilst working on this book and for making this task a great success. We would like to express our sincere appreciation to Prathik Roy, Group Product Manager, Nanoscience and Technology, Database Group, Springer Nature, Jersey City, NJ 07302, who encouraged us to initiate this project and also introduced us to the right person, Dr. Anita Lekhwani, to evaluate our proposal. We are also immensely grateful to Dr. Anita, who supported our proposal from the beginning and took all the initiative to get the approval of this project. We owe an enormous debt of gratitude to all the authors who contributed to research articles for the success of the physical creation and completion of this volume. Dr. S. Balakumar gratefully acknowledges supports received from the University of Madras. He would also like to extend his warmest thanks to research team members of the National Centre for Nanoscience and Nanotechnology for their constant support and enthusiasm which helped him to complete the task on time. Dr. M.V. Shankar would like to express his special gratitude and thanks to Prof. Munagala Surya Kalavathi, Vice Chancellor of Yogi Vemana University, Andhra Pradesh, India, for her valuable suggestions and constant inspiration. He is also thankful to alumni and present members of Nanocatalysis and Solar Fuels Research group for their dedicated research work and contributions to this book. Dr. Keller acknowledges financial support from ICPEES, Institut de Chimie et des Procédés pour l’Energie, l ‘Environnement et la Santé, UMR 7515, CNRS Université de Strasbourg, for this book project.

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Contents

1 Nanostructures in Photocatalysis: Opportunities and Challenges for Environmental Applications��������������������������������������������������������������    1 Y. V. Divyasri, Y. N. Teja, V. Nava Koteswara Rao, N. C. Gangi Reddy, Sakar Mohan, M. Mamatha Kumari, and M. V. Shankar 2 Nanostructured Heterojunction (1D-0D and 2D-0D) Photocatalysts for Environmental Remediation������������������������������������   35 Lakshmana Reddy Nagappagari, Kiyoung Lee, Ajay Rakesh, Subramanian Balakumar, and M. V. Shankar 3 Hierarchical Nanostructures for Photocatalytic Applications ������������   67 R. Ajay Rakkesh, Durgalakshmi Dhinasekaran, M. V. Shankar, and S. Balakumar 4 Nanocomposite Photocatalysts for the Degradation of Contaminants of Emerging Concerns����������������������������������������������������   89 Rokesh Karuppannan, Sakar Mohan, and Trong-On Do 5 Sunlight-Mediated Plasmonic Photocatalysis: Mechanism and Material Prospects����������������������������������������������������������������������������  119 Durgalakshmi Dhinasekaran, M. R. Ashwin Kishore, and Mohanraj Jagannathan 6 Photocatalytic Efficiency of Bi-Based Aurivillius Compounds: Critical Review and Discernment of the Factors Involved������������������  143 Manjunath Shetty, Murthy Muniyappa, M. Navya Rani, Vinay Gangaraju, Prasanna D. Shivaramu, and Dinesh Rangappa 7 Intrinsically Conducting Polymer Nanocomposites in Shielding of Electromagnetic Pollution������������������������������������������������������������������  173 Suneel Kumar Srivastava

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8 Nanostructuring of Hybrid Materials Using Wrapping Approach to Enhance the Efficiency of Visible Light-Responsive Semiconductor Photocatalyst������������������������������������������������������������������  223 V. Vinesh, A. R. Mahammed Shaheer, and B. Neppolian 9 Metal–Organic Frameworks (MOFs) with Hierarchical Structures for Visible Light Photocatalysis ������������������������������������������  239 P. Karthik and B. Neppolian 10 Soil Remediation by Zero-Valent Iron Nanoparticles for Organic Pollutant Elimination����������������������������������������������������������  253 Marco Stoller, Luca Di Palma, and Giorgio Vilardi 11 Black TiO2: An Emerging Photocatalyst and Its Applications������������  273 P. Anil Kumar Reddy, P. Venkata Laxma Reddy, and S. V. Prabhakar Vattikuti 12 Nanomaterials for Photocatalytic Decomposition of Endocrine Disruptors in Water ��������������������������������������������������������������������������������  305 Ajay Kumar, Vishal Sharma, Ashish Kumar, and Venkata Krishnan 13 Carbonaceous Nanomaterials for Environmental Remediation����������  327 Natarajan Sasirekha and Yu-Wen Chen 14 Magnetically Recyclable Photocatalysts for Degradation of Organic Pollutants in Aquatic Environment������������������������������������  371 Ashutosh Kumar and Sushil Kumar Kansal 15 Titanate Nanostructures as Potential Adsorbents for Defluoridation of Water��������������������������������������������������������������������������  389 C. Prathibha, Anjana Biswas, and M. V. Shankar 16 Photocatalytic Water Pollutant Treatment: Fundamental, Analysis and Benchmarking ������������������������������������������������������������������  407 Katherine Rebecca Davies, Ben Jones, Chiaki Terashima, Akira Fujishima, and Sudhagar Pitchaimuthu 17 Graphene-Based Photocatalytic Materials: An Overview ������������������  439 Alex T. Kuvarega, Rengaraj Selvaraj, and Bhekie B. Mamba 18 Recent Advances in Nanostructured Materials for Detoxification of Cr(VI) to Cr(III) for Environmental Remediation��������������������������  461 Udayabhanu, S. B. Patil, and G. Nagaraju 19 Metal Nitrides and Graphitic Carbon Nitrides as Novel Photocatalysts for Hydrogen Production and Environmental Remediation����������������������������������������������������������������������������������������������  491 Sudesh Kumar, Kakarla Raghava Reddy, Ch. Venkata Reddy, Nagaraj P. Shetti, Veera Sadhu, M. V. Shankar, Vasu Govardhana Reddy, A. V. Raghu, and Tejraj M. Aminabhavi

Contents

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20 Highly Functionalized Nanostructured Titanium Oxide-Based Photocatalysts for Direct Photocatalytic Decomposition of NOx/VOCs�����������������������������������������������������������������  527 Katchala Nanaji, Manavalan Vijayakumar, Ammaiyappan Bharathi Sankar, and Mani Karthik 21 Bandgap Engineering as a Potential Tool for Quantum Efficiency Enhancement��������������������������������������������������������������������������  551 Reddy Kunda Siri Kiran Janardhana, Raju Kumar, Tata Narsinga Rao, and Srinivasan Anandan 22 Nanostructure Material-Based Sensors for Environmental Applications����������������������������������������������������������������������������������������������  571 Vinutha Srikanth, Mahesh Shastri, M. Sindhu Sree, M. Navya Rani, Prasanna D. Shivaramu, and Dinesh Rangappa 23 Nanostructured MoS2 as Non-noble Metal-­Based Cocatalyst for Photocatalytic Applications��������������������������������������������������������������  597 Murthy Muniyappa, Manjunath Shetty, Mahesh Shastri, S. Jagadeesh Babu, M. Navya Rani, Prasanna D. Shivaramu, and Dinesh Rangappa Index������������������������������������������������������������������������������������������������������������������  611

Contributors

Tejraj M. Aminabhavi  Sonia College of Pharmacy, Dharwad, Karnataka, India Srinivasan Anandan  Centre for Nano Materials, International Advanced Research Centre for Powder Metallurgy and New Materials, Hyderabad, Telangana, India Subramanian Balakumar  National Centre for Nanoscience and Nanotechnology (NCNN), University of Madras, Chennai, Tamil Nadu, India Anjana Biswas  Department of Physics, Sri Sathya Sai Institute of Higher Learning, Anantapur Campus, Anantapur, Andhra Pradesh, India Yu-Wen  Chen  Department of Chemical and Materials Engineering, National Central University, Chung-Li, Taiwan Katherine  Rebecca  Davies  Multi-functional Photocatalyst and Coatings Group, SPECIFIC, College of Engineering, Swansea University (Bay Campus), Swansea, Wales, UK Luca  Di Palma  Department of Chemical Engineering Materials Environment, Sapienza University of Rome, Rome, Italy Y. V. Divyasri  Department of Chemistry, Yogi Vemana University, Kadapa, Andhra Pradesh, India Trong-On  Do  Department of Chemical Engineering, Laval University, Quebec, QC, Canada Durgalakshmi Dhinasekaran  Department of Medical Physics, Anna University, CEG Campus, Chennai, Tamil Nadu, India Akira Fujishima  Photocatalysis International Research Center, Tokyo University of Science, Nodashi, Chiba ken, Japan Vinay Gangaraju  Department of Applied Sciences, Visvesvaraya Technological University, Center for Postgraduate Studies, Muddenahalli, Chikkaballapur, Karnataka, India S. Jagadeesh Babu  Department of Applied Sciences, Visvesvaraya Technological University, Center for Postgraduate Studies, Muddenahalli, Chikkaballapur, Karnataka, India

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Contributors

Mohanraj Jagannathan  Department of Medical Physics, Anna University, CEG Campus, Chennai, Tamil Nadu, India Reddy Kunda Siri Kiran Janardhana  Centre for Nano Materials, International Advanced Research Centre for Powder Metallurgy and New Materials, Hyderabad, Telangana, India Ben  Jones  Multi-functional Photocatalyst and Coatings Group, SPECIFIC, College of Engineering, Swansea University (Bay Campus), Swansea, Wales, UK Sushil  Kumar  Kansal  Dr. S.  S. Bhatnagar University Institute of Chemical Engineering and Technology, Panjab University, Chandigarh, India Mani Karthik  Centre for Nanomaterials, International Advanced Research Centre for Powder Metallurgy and New Materials (ARCI), Balapur, Hyderabad, Telangana, India P.  Karthik  SRM Research Institute, SRM Institute of Science and Technology, Kattankulathur, Chennai, Tamil Nadu, India Rokesh  Karuppannan  Department of Chemical Engineering, Laval University, Quebec, QC, Canada M. R. Ashwin Kishore  Department of Chemical Engineering, University of Seoul, Seoul, Republic of Korea Venkata  Krishnan  School of Basic Sciences and Advanced Materials Research Center, Indian Institute of Technology Mandi, Kamand, Himachal Pradesh, India Ajay Kumar  School of Basic Sciences and Advanced Materials Research Center, Indian Institute of Technology Mandi, Kamand, Himachal Pradesh, India Ashish  Kumar  School of Basic Sciences and Advanced Materials Research Center, Indian Institute of Technology Mandi, Kamand, Himachal Pradesh, India Ashutosh  Kumar  School of Energy and Environment, Thapar Institute of Engineering and Technology, Patiala, Punjab, India M.  Mamatha  Kumari  Nanocatalysis and Solar Fuels Research Laboratory, Department of Materials Science & Nanotechnology, Yogi Vemana University, Kadapa, Andhra Pradesh, India Raju Kumar  Centre for Nano Materials, International Advanced Research Centre for Powder Metallurgy and New Materials, Hyderabad, Telangana, India Sudesh  Kumar  Department of Chemistry, Banasthali Vidyapeeth, Vanasthali, Rajasthan, India Alex  T.  Kuvarega  Nanotechnology and Water Sustainability Research Unit, College of Science, Engineering and Technology, University of South Africa, Florida Science Campus, Florida, Johannesburg, South Africa

Contributors

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Kiyoung  Lee  Department of Energy Chemical Engineering, School of Nano & Materials Science and Engineering, Kyungpook National University, Sangju, Republic of Korea Bhekie  B.  Mamba  Nanotechnology and Water Sustainability Research Unit, College of Science, Engineering and Technology, University of South Africa, Florida Science Campus, Florida, Johannesburg, South Africa Murthy Muniyappa  Department of Applied Sciences, Visvesvaraya Technological University, Center for Postgraduate Studies, Muddenahalli, Chikkaballapur, Karnataka, India Lakshmana Reddy Nagappagari  Department of Energy Chemical Engineering, School of Nano & Materials Science and Engineering, Kyungpook National University, Sangju, Republic of Korea G.  Nagaraju  Energy Materials Research Laboratory, Department of Chemistry, Siddaganga Institute of Technology, Tumakuru, Karnataka, India Katchala  Nanaji  Centre for Nanomaterials, International Advanced Research Centre for Powder Metallurgy and New Materials (ARCI), Balapur, Hyderabad, Telangana,, India B.  Neppolian  Energy and Environmental Remediation Lab, SRM Research Institute, SRM Institute of Science and Technology, Kattankulathur, Chennai, Tamil Nadu, India S.  B.  Patil  Energy Materials Research Laboratory, Department of Chemistry, Siddaganga Institute of Technology, Tumakuru, Karnataka, India Department of Chemistry, The Oxford College of Science, Bengaluru, Karnataka, India Sudhagar  Pitchaimuthu  Multi-Functional Photocatalyst and Coatings Group, SPECIFIC, College of Engineering, Swansea University (Bay Campus), Swansea, Wales, UK C. Prathibha  Department of Physics, Sri Sathya Sai Institute of Higher Learning, Anantapur Campus, Anantapur, Andhra Pradesh, India A. V. Raghu  Department of Chemistry, Faculty of Engineering and Technology, Jain (Deemed-to-be University), Bangalore, Karnataka, India Ajay  Rakesh  National Centre for Nanoscience and Nanotechnology (NCNN), University of Madras, Chennai, Tamil Nadu, India R.  Ajay  Rakkesh  National Centre for Nanoscience and Nanotechnology, University of Madras, Chennai, Tamil Nadu, India Department of Physics and Nanotechnology, SRM Institute of Science and Technology, Kattankulathur, Tamil Nadu, India

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Contributors

Dinesh Rangappa  Department of Applied Sciences, Visvesvaraya Technological University, Center for Postgraduate Studies, Muddenahalli, Chikkaballapur, Karnataka, India M. Navya Rani  School of Basic and Applied Sciences, Dayanand Sagar University, Bengaluru, Karnataka, India Tata Narsinga Rao  Centre for Nano Materials, International Advanced Research Centre for Powder Metallurgy and New Materials, Hyderabad, Telangana, India V.  Nava  Koteswara  Rao  Nanocatalysis and Solar Fuels Research Laboratory, Department of Materials Science & Nanotechnology, Yogi Vemana University, Kadapa, Andhra Pradesh, India Ch.  Venkata  Reddy  School of Mechanical Engineering, Yeungnam University, Gyeongsan, South Korea Kakarla  Raghava  Reddy  School of Chemical and Biomolecular Engineering, The University of Sydney, Sydney, NSW, Australia N. C. Gangi Reddy  Department of Chemistry, Yogi Vemana University, Kadapa, Andhra Pradesh, India P.  Anil  Kumar  Reddy  School of Mechanical and Nuclear Engineering, Ulsan National Institute of Science and Technology (UNIST), Ulsan, South Korea P. Venkata Laxma Reddy  Program in Environmental Science and Engineering, University of Texas El Paso, El Paso, TX, USA Vasu  Govardhana  Reddy  Department of Chemistry, Yogi Vemana University, Kadapa, Andhra Pradesh, India Veera Sadhu  School of Physical Sciences, Kakatiya Institute of Technology and Science (KITS), Warangal, Telangana, India Sakar Mohan  Centre for Nano and Material Sciences, Jain University, Bangalore, Karnataka, India Department of Chemical Engineering, Laval University, Quebec, QC, Canada Ammaiyappan  Bharathi  Sankar  School of Electronics Engineering, Vellore Institute of Technology (VIT), Chennai Campus, Chennai, Tamil Nadu, India Natarajan  Sasirekha  CAS in Crystallography and Biophysics, University of Madras, Guindy Campus, Chennai, Tamil Nadu, India Rengaraj  Selvaraj  Chemistry Department, Sultan Qaboos University, Muscat, Sultanate of Oman A. R. Mahammed Shaheer  Energy and Environmental Remediation Lab, SRM Research Institute, SRM Institute of Science and Technology, Kattankulathur, Chennai, Tamil Nadu, India

Contributors

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M. V. Shankar  Nanocatalysis and Solar Fuels Research Laboratory, Department of Materials Science & Nanotechnology, Yogi Vemana University, Kadapa, Andhra Pradesh, India Vishal  Sharma  School of Basic Sciences and Advanced Materials Research Center, Indian Institute of Technology Mandi, Kamand, Himachal Pradesh, India Mahesh  Shastri  Department of Applied Sciences and Visvesvaraya Center for Nanoscience and Technology, Centre Post Graduation Studies, Visvesvaraya Technological University, Muddenahalli Campus, Chikkaballapura, India Nagaraj  P.  Shetti  Department of Chemistry, K.  L. E.  Institute of Technology, Hubli, Karnataka, India Visvesvaraya Technological University, Belgaum, Karnataka, India Manjunath Shetty  Department of Applied Sciences, Visvesvaraya Technological University, Center for Postgraduate Studies, Muddenahalli, Chikkaballapur, Karnataka, India Prasanna  D.  Shivaramu  Department of Applied Sciences, Visvesvaraya Technological University, Center for Postgraduate Studies, Muddenahalli, Chikkaballapura, India M. Sindhu  Sree  Department of Applied Sciences and Visvesvaraya Center for Nanoscience and Technology, Centre Post Graduation Studies, Visvesvaraya Technological University, Muddenahalli Campus, Chikkaballapura, India Vinutha Srikanth  Department of Electrical and Electronics Engineering, KSSEM, Bengaluru, Karnataka, India Suneel  Kumar  Srivastava  Department of Chemistry, Indian Institute of Technology, Kharagpur, West Bengal, India Marco  Stoller  Department of Chemical Engineering Materials Environment, Sapienza University of Rome, Rome, Italy Y.  N.  Teja  Centre for Nano and Material Sciences, Jain University, Bangalore, Karnataka, India Chiaki Terashima  Photocatalysis International Research Center, Tokyo University of Science, Nodashi, Chiba ken, Japan Udayabhanu  Energy Materials Research Laboratory, Department of Chemistry, Siddaganga Institute of Technology, Tumakuru, Karnataka, India Center for Research and Innovations, BGSIT, Adichunchanagiri University, B. G. Nagara, Mandya, Karnataka, India S.  V.  Prabhakar  Vattikuti  School of Mechanical Engineering, Yeungnam University, Gyeongsan, South Korea

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Contributors

Manavalan  Vijayakumar  Centre for Nanomaterials, International Advanced Research Centre for Powder Metallurgy and New Materials (ARCI), Balapur, Hyderabad, India Global Innovative Centre for Advanced Nanomaterials (GICAN), Collage of Science, Engineering and Environment, The University of Newcastle, NSW, Callaghan, Australia Giorgio  Vilardi  Department of Chemical Engineering Materials Environment, Sapienza University of Rome, Rome, Italy V. Vinesh  Energy and Environmental Remediation Lab, SRM Research Institute, SRM Institute of Science and Technology, Kattankulathur, Chennai, Tamil Nadu, India

About the Editors

Subramanian Balakumar, FRSC, FASCh.  is a Professor and the Director of the National Centre for Nanoscience and Nanotechnology, University of Madras. He is a distinguished academician and a researcher par excellence in the field of nanoscience. He received his Ph.D. in the Faculty of Sciences, from the prestigious Anna University, India, in 1996 and further escalated his research competence as a postdoctoral fellow with Prof. J.B.  Xu, Department of Electronic Engineering, the Chinese University of Hong Kong, Hong Kong, from 1995 to 1997, and subsequently worked as NSTB postdoctoral fellow with Prof. Zeng Hua Chen, Department of Chemical Engineering, National University of Singapore, from 1997 to 1999. His remarkable stint in industry began as a Senior Research Engineer in Chartered Semiconductor Manufacturing Ltd., Singapore (now GLOBALFOUNDRIES, Singapore), till 2002 and consecutively worked as Senior Scientist at A-Star Institute of Microelectronics till 2008. In addition to his presence in industry, he remained deeply connected with academia as an Adjunct Associate Professor, Department of Mechanical Engineering, National University of Singapore, from 2004 to 2008. With intense urge in academics, he joined the highly renowned NCNSNT, University of Madras, during November 2008. Predictably, he established an unsurpassable academic footprint with more than 250 papers in peer-reviewed journals and also has more than 8 US patents to his credit. He has published several book chapters and edited many books. His versatile administrative capabilities were emblazoned by organizing 40 international and national conferences, workshops, seminars, shortterm courses, and seminars during the past 12 years. His formidable research conduct fetched him the role of Associate Editor of Chemical Papers Journal (Springer), and he serves as an editorial board member for various journals. He has received more than 20 funded projects from various funding agencies. As an epitome of his credits, he was recently bestowed with Tamil State Best Scientist Award 2018 (TANSA 2018) and received the Senior Scientist Award from TNHE, Science City, during 2019. He was also conferred fellow of Royal Society of Chemistry. His research work includes nanostructured materials for energy, environment, and biomedical applications.

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About the Editors

Valérie  Keller  is a senior scientist (Research Director) at ICPEES, Institute of Chemistry and Processes for Energy, Environment and Health in Strasbourg (France). She received her Ph.D. degree in Chemistry and Catalysis from the Louis Pasteur University of Strasbourg in 1993. In 1996, she returned to Strasbourg and was appointed as researcher in CNRS, where she was promoted as Research Director in 2012. She is now responsible for the team “Photocatalysis, Photoconversion and Green Chemical Processes.” Her main research activities concern photocatalysis for environmental, energy, and health applications and the synthesis and characterization of nanomaterials for photoconversion purposes. She is the author of over 130 publications in peer-reviewed journals and more than 100 oral communications in international conferences and symposium. She is also the author of 15 patents. In 2013, she was awarded the First Prize of the Strategic Reflection (awarded by the French Prime Minister, Manuel Valls). Since 2016, she has been co-director of the French Solar Fuels (GDR) Network. M. V. Shankar  is a Professor of Materials Science and Nanotechnology in Yogi Vemana University, Kadapa, Andhra Pradesh, India. He is leading Nanocatalysis and Solar Fuels research group in YV University. Renowned for his outstanding research work in the field of photocatalysis for hydrogen fuel production and multifunctional application of nanomaterials, he has authored/coauthored 122 publications and 6 patents with an h-index of 36, i-10 index of 65, and average impact factor of 4.69 with 6124+ citations. Highlight of his research work is the use of earth-abundant materials, simple concepts, and scalable experimental methods in the preparation of photocatalysts. He has proved that crude glycerol- and sulfide-­ containing wastewater can be used as a source for large-scale hydrogen generation under sunlight. He has chaired many national and international conferences, seminars, symposia, workshops, and DST-INSPIRE programs. He has delivered several keynotes and invited lectures. Dr. Shankar has very good R&D collaborators in the area of energy, environment, and healthcare applications. He has rich research experience in countries, viz., France, Japan, and Taiwan, and is a recipient of several prizes, awards, and fellowships. He is a recipient of several awards, including Fellow of Indian Chemical Society, Charted Chemist, and Fellow of Royal Society of Chemistry (FRSC), London. Dr. Shankar’s research work has been highlighted in national newspapers and TV channels, earning the mark of exceptional work.

Chapter 1

Nanostructures in Photocatalysis: Opportunities and Challenges for Environmental Applications Y. V. Divyasri, Y. N. Teja, V. Nava Koteswara Rao, N. C. Gangi Reddy, Sakar Mohan, M. Mamatha Kumari, and M. V. Shankar

1.1  Introduction The environment, which is essentially the air, water, and soil, is largely polluted due to the increased population and industrialization. These pollutants are mostly anthropogenic, and they generally include (i) the toxic-organic materials such as dyes, aromatic, and aliphatic molecules; (ii) agricultural wastages such as the pesticides, insecticides, and herbicides; (iii) plastics; (iv) pharmaceutical products and byproducts; (v) inorganic materials such as heavy metals; (vi) toxic gases such as CO, SOx, and NOx; and (vii) microorganisms such as bacteria, viruses, and fungi [1–3]. Release of these pollutants into the environment from various sources causes much adverse effects to the ecology, and it will make permanent damages and even more worse adverse effects if these pollutants are accumulated into the environment. Therefore, it is an urgent requirement to address such issues toward destructing and converting these pollutants into nontoxic. Considering current scenario of energy consumption, the world also requires energy- and cost-effective techniques to address the issues in the environmental remediation. In this aspect, photocatalysis is one of the reliable energy- and cost-effective and versatile techniques, which can almost degrade/convert into nontoxic/kill all of the abovementioned various categories of pollutants in the environment [4, 5]. Y. V. Divyasri · N. C. Gangi Reddy Department of Chemistry, Yogi Vemana University, Kadapa, Andhra Pradesh, India e-mail: [email protected] Y. N. Teja · S. Mohan Centre for Nano and Material Sciences, Jain University, Bangalore, Karnataka, India e-mail: [email protected] V. Nava Koteswara Rao · M. Mamatha Kumari · M. V. Shankar (*) Nanocatalysis and Solar Fuels Research Laboratory, Department of Materials Science & Nanotechnology, Yogi Vemana University, Kadapa, Andhra Pradesh, India e-mail: [email protected] © Springer Nature Switzerland AG 2021 S. Balakumar et al. (eds.), Nanostructured Materials for Environmental Applications, https://doi.org/10.1007/978-3-030-72076-6_1

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Photocatalysis is the process that leads to the oxidation and reduction reactions, where it is being done using a photoactive semiconductor (photocatalyst) by activating it using a suitable light source such as UV and/or visible light. The research in photocatalysis was started after the discovery of water splitting using the illuminated TiO2 by Fujishima and Honda in 1972 [6]. This profound observation revealed the concept of photo-induced reduction-oxidation (redox) reactions on the semiconductors. Eventually, it was realized that the redox reactions can be effectively applied to the environmental remediation applications when the photocatalytic oxidation of CN− and SO−3 was demonstrated on various semiconductors. Later, TiO2 was demonstrated for its photocatalytic ability to degrade the chlorinated compounds [7, 8] and kill microorganisms [9]. Research in photocatalysis has been advanced in many different ways such as the development of new processes such as advanced oxidation process (AOP) [10], Fenton’s process [11], photocatalytic memory effect [12], and new material systems such as heterojunction [13], plasmonic [14, 15], Z-Scheme [16], and ferroelectric-photocatalysts [17]. The requirement of different photocatalytic systems is essentially to produce redox species with suitable energy to destroy various pollutants. Therefore, the photocatalyst design, especially at nanoscale, has become important, and as a result there have been many synthesis methods emerged toward the synthesis of various nanostructures of photocatalytic systems. In this chapter, we have described the photocatalytic mechanism toward environmental remediation, various synthesis methods to produce nanostructured photocatalysts, and various photocatalytic applications toward environmental remediation that include the degradation of a range of pollutants such as dyes, pharmaceutical, pesticides, volatile compounds, plastics; detoxification of various heavy metals and gases; conversion of greenhouse gas CO2 into hydrocarbon fuels; and killing of pathogens such as bacteria, fungi, etc. Finally, it concludes with the summary and future perspective on how the nanostructuring of photocatalysts can improve the photocatalytic phenomenon and process.

1.2  Mechanics of Photocatalysis in Nanostructures The mechanism of photocatalysis generally involves the excitation of electron-hole pairs in a photocatalytic semiconductor under suitable light energy, where they further get promoted to the surrounding and produce the radical species to perform the reduction and oxidation reactions toward the desirable photocatalytic applications. In this process, the relative band edge position of the photocatalyst with respect to the redox species is very much important to have the desirable photocatalytic process. For instance, the photocatalytic water splitting requires the conduction band should be more positive and the valence band should be more negative. Similarly, the specific application needs specific type of radical species that essentially originates either electrons or holes. For instance, dye degradation process happens through oxidation reactions, where it requires the participation of holes, while the

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toxic heavy metals can be converted into nontoxic via reduction reactions that require the electrons. Considering the process of destructing the pollutants, the required photocatalytic reactions can be depicted as shown in Fig. 1.1, and it can be described in the following Eqs. (1.1)–(1.9).

PC + hv → e − + h +

(1.1)



h + + H 2 O → H + + OHÙ

(1.2)



h + + OH −



e − + O2

→ OHÙ → O2 −



2e − + O2 + 2H +



2e − + H 2 O2

(1.4)

→ H 2 O 2

(1.5)

→ OHÙ+OH −

(1.6)

R + OHÙ → Degradation products



(1.3)

(1.7)



R + h+

→ Oxidation degraded products

(1.8)



R + e−

→ Reduction degraded products

(1.9)

These abovementioned reactions and mechanisms are about the chemical reactions that occur in the process. However, these reactions can be effectively controlled or enhanced through physical structuring of photocatalysts, where the potential energy of generated electrons and holes thereby the radical species is often determined by the energy band edge (VB and CB) positions in the given photocatalyst. For instance, the band edge positions can be effectively tuned via reducing the particle size and changing the morphology of the photocatalysts as shown in Fig. 1.2(a)–(c), where it shows that when the layer thickness of g-C3N4 is reduced from its bulk structure to the nanomesh-like structure (a), the relative band edge positions get shifted (b, c) in g-C3N4 [18]. On the other hand, the band structures can be modified via doping and forming heterojunction, etc., where the former introduces new energy levels between the VB and CB, while the latter creates a kind of cascading process/channels to transport the excited carriers for the required redox reactions [13, 19, 20]. However, the Fig. 1.1 Photocatalytic mechanism toward the environmental remediation applications

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Fig. 1.2 (a) Exfoliation of bulk g-C3N4 into nanomesh, (b) XPS valence band spectra, and (c) relative shift in the band edge position in nanomesh of g-C3N4 as compared to its bulk [18]

nanostructure-induced enhancement in the photocatalytic process is relatively exotic as it leads to geometrics-dependent optical and electronic properties. For instance, recombination of charge carriers in one-dimensional nanostructures is relatively enhanced as compared to the spherical nanostructures, which is essentially due to the enhanced delocalization of electrons in the CB of one-­dimensionally structured photocatalysts.

1.3  Synthesis of Nanostructured Photocatalysts Synthesis of nanostructures can be broadly divided into two categories such as (i) top-down and (ii) bottom-up approaches. Most of the photocatalytic nanostructures have been prepared via bottom-up approach due to the specific reasons such as doping, composite formation, etc. Moreover, the bottom-up approach-mediated syntheses of nanomaterials pave ways to effectively control the dimension and morphology of the nanostructures. This section provides a glimpse on some of the selective synthesis methods (from the recent reports) toward synthesizing various photocatalytic nanostructures such as quantum dots, nanospheres, nanotubes, nanocubes, nanoneedles, nanofibers, nanoporous structures, nanosheets, nanoflowers, and other anisotropic nanostructures that are successfully used for a variety of photocatalytic applications.

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1.3.1  Sol–Gel Synthesis Sol-gel process generally involves the transformation of precursor solution into gel. In the first step of process, the metal precursors are hydrolyzed thereby forming the metal hydroxide matrix solution. Typically in the second step, the solution is condensed where the formation of gel takes place. Thereafter the gel is subjected to the process of drying or calcination where desired nanostructures are obtained. Extensive use of this method is done for synthesis of different nanostructures as it gives the advantage of controlling the morphology such as size and shape at low temperatures. Flake-like BiVO4 nanostructures were prepared using bismuth nitrate (Bi (NO3)3·5H2O) and ammonium vanadate (NH4VO3) as initial precursors and were used under visible light irradiation for the degradation of methylene blue (MB) [21]. Homogeneous uniformly dispersed spherical−/elliptical-shaped cadmium (Cd)-doped cerium oxide (CeO2) nanostructures with ferromagnetic nature were synthesized for the purpose of dye degradation using sol–gel method with cerium chloride heptahydrate (CeCl3∙7H2O) and cadmium chloride (CdCl2) as starting precursors with varying doping concentration of 1–3% [22]. The nanostructures showed an efficiency of 86.42% and 92.53% for the dye degradation of rhodamine B (RhB) and Congo red (CR) dyes, respectively, within 6-h irradiation under UV– Vis irradiation. In another example, spindle-like ZnO nanostructures codoped with neodymium (Nd) and vanadium (V), synthesized by ultrasonic-assisted sol–gel method, showed excellent efficiency for the photocatalytic degradation of organic pollutants under visible light irradiation [23]. Pure ZnO and Nd mono-doped ZnO nanostructures showed perfect spindle-like morphology, whereas the ZnO particles codoped with 4% Nd and 1% V showed more like a needle-shaped morphology with reduced diameter size. Using sol–gel method, TiO2–CdO–Ag nanocomposites were synthesized with an average particle size of 20  nm but with irregular morphologies [24]. The nanocomposites showed almost 92% dye degradation of methylene blue within 75-min irradiation under visible light. Also for the photocatalytic degradation of organic dyes, Cd2V2O7 nanostructures with an average diameter size of 10–20 nm were prepared [25]. These results also evidence that the sol-gel process could be effective to produce nanostructured materials in large scale as well.

1.3.2  Hydro-/Solvo-thermal Synthesis Hydrothermal synthesis process involves the performance of chemical reactions among precursors dissolved in the aqueous solvent under autogenous pressure, where the temperature of the solvent is maintained above the critical point. Solvo-­ thermal process is also similar to hydrothermal except that the solvent used is nonaqueous. Most of the hydro−/solvo-thermal processes are carried out in a metal autoclave made up of Teflon or alloy linings. For example, flowerlike ZnO nanostructures were synthesized using hydrothermal process [26]. In this procedure, two

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10 mL aqueous solutions, one with zinc acetate and another with potassium hydroxide (KOH), were prepared separately. Both solutions were added to double-distilled water and stirred vigorously together, and the resulting mixed solution was transferred into a Teflon-lined stainless steel autoclave and was heated for 8 h at a temperature of 120 °C. The resulting precipitate was centrifuged and washed with water and alcohol for several times to obtain a balanced pH and was finally dried at 60 °C for 24 h in an oven. It is observed that the obtained nanostructures exhibited flowerlike uniform morphology with an average diameter of 2–4 μm and were used for photocatalytic dye degradation of methylene blue dye. Core-shell CeO2 nanospheres were successfully synthesized using template-free hydrothermal method for the methyl orange dye degradation [27]. The average size of nanospheres synthesized with 20 mmol of urea was found to be 50–145 nm and showed degradation efficiency as high as 93.49% under UV irradiation of 160  min. Other examples of hydrothermally synthesized nanostructures are SnO2@ZnO [28] and CuO@ZnO [29] heterojunctions for H2O evolution and dye degradation, respectively. Using solvo-thermal method, Rh-In2O3 3D urchin-like and 2D rodlike nanostructures with an average diameter size of 500–600 nm and 20 nm–1 μm, respectively, were synthesized [30]. Other examples of solvo-thermally synthesized nanostructures include flowerlike BiVO4 [31] and CdS nanostructures [32], ZnO nanospheres, and hexagonal disklike ZnO nanostructures [33].

1.3.3  Precipitation Process Precipitation method is a cost-effective and low-temperature approach used for the synthesis of photocatalytic nanostructures. Typically, precipitation synthesis method involves three steps. In the first step, two types of solutions, one with inorganic metal salt precursor and another one with a chosen surfactant, are prepared separately by dissolving in water. In second step, the two solutions are added and stirred vigorously, and a basic solution such as NaOH is added in a dropwise manner under stirring and a precipitate is formed. In the next step, the precipitate is collected by centrifuging and washed with ethanol and water several times to balance the pH value, and finally the precipitate is dried in an oven followed by calcinations in some cases. For example, CuO nanosheets were prepared by dissolving copper sulfate pentahydrate (CuSO4·5H2O) and cetyltrimethylammonium bromide (CTAB) in deionized water followed by vigorous stirring and addition of NH3·H2O and NaOH one after the other in a dropwise manner [34]. After stirring the solution for a certain amount of time, a blue-colored precipitate was formed which was kept at room temperature for 2 days to settle down completely, and finally a black-colored precipitate was obtained. It was washed with ethanol and water several times, centrifuged, and finally dried at 80 °C for 6 h. The nanosheets showed an efficiency of 82% in dye degradation under visible light irradiation. Other examples of photocatalytic nanostructures synthesized through precipitation approach include CeO2/ Au/Ho cube-like and pyramid-like nansostructures [35], spherical holmium oxide

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(HoO3) nanoparticles [36], neodymium oxide (Nd2O3) bundle-like and spherical nanostructures [37], and SnO nanosheets [38].

1.3.4  Anodization Process Anodization process involves development of an oxide layer on the surface of metal using electrolytic reactions. Even though anodization was employed for synthesis of various mesoporous nanostructures, TiO2 nanotubes synthesized through anodization are most exploited for photocatalysis. In 1999 Zwilling et al. first reported the synthesis of self-assembled TiO2 nanotubes by anodization [39]. Generally the anodization setup consists of two electrode system with Ti as anode and inert metal as cathode (mostly Pt is used), an acidic electrolyte (mostly fluoride ion electrolyte is used and a DC power supply as shown in Fig. 1.3). The process of nanotube formation mainly involves the simultaneous reactions of oxidation and dissipation. First electrochemical oxidation of Ti into TiO2 takes place, which forms a compact oxide layer on the surface of Ti foil, and then due to fluorine ions (F−), small pores or pits on the compact layer are formed which reduces the resistance toward the applied current, and finally an equilibrium between the nanoporous oxide layer and chemical dissipation due to electrical field induction as well as fluorine ion-induced dissipation leads to the formation of nanotubes. Morphological properties such as size, shape, and crystalline phase of nanotubes are influenced by various anodization parameters such as type of electrolyte and electrodes, pH value of electrolyte, applied voltage, temperature, and current density.

Fig. 1.3  Schematic representation of anodization setup for synthesis of TiO2 nanotubes [41]

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TiO2 nanotubes were synthesized for degradation of methylene blue under UV irradiation using Ti foil as anode and stainless steel foil as cathode, respectively, and an electrolyte mixture of ethylene glycol, deionized water, and NH4F [40]. Nanotubes were synthesized after various voltages of 20–60  V was supplied for 24 h at a temperature of 30 °C. It is observed that with increase in voltage there was an increase in the diameter size of the nanotubes: 20 V, 80 nm; 40 V, 110 nm; and 60 V, 130 nm. Iron oxide nanostructures for the purpose of water splitting were also synthesized by anodization using iron rods and Pt foil as electrodes and an electrolyte same as in case of TiO2, but here different temperatures have been employed in order to observe the morphological changes [41]. At the temperature of 25 °C, stable nanotube structures were observed, but when the temperature was increased slowly, the morphology shrunk and finally at a temperature of 60 °C granular layer-­ like nanostructures were observed.

1.3.5  Electrospinning Electrospinning is widely used for the synthesis of nanofibers. It is a simple, cost-­ effective, and quick method to produce nanofibers of diameter size from submicron-­ size to nano-size by applying a high electric DC field to a polymer solution or melting at room temperature. Electrospun nanofibers carry various advantages such as higher surface area, high porosity, and small pore size, which are highly favorable for photocatalysis. The electrospinning setup is shown in Fig. 1.4. An electrically charged polymer jet is created when high electric DC field is applied to the polymer solution or melt. When the electric field is further increased, the polymer jet experiences stability changes where it is stretched before it reaches

Pump Syringe

Composite Solution Needle Taylor Cone

Power Supply (High voltage)

Solution Jet (stability)

V

Solution Jet (instability) / whipping

Nanofibers

Solution Parameters Viscosity/Concentration Conductivity Molecular weight Surface Tension Solvent Selection

Electrospinning Parameters Voltage Supply Needle Diameter t Flow Rate Collector

Ambient parameters Temperature Humidity Collector

Fig. 1.4  Schematic representation of electrospinning setup with various parameters [42]

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the collector placed just below. Nanofibers are obtained on the collector where the polymer jet is stiffened into fibers by evaporation of solvent. The diameter size and the functionality of the nanofibers can be altered by controlling different parameters mentioned in Fig. 1.4. CuO/SnO2 hallow nanofibers were developed through electrospinning followed by annealing for the degradation of methylene blue under visible light irradiation [43]. The obtained fiber morphology consisted of rough surfaces with average diameter size of 400–600 nm. α-Fe2O3 nanowires with an average diameter of 100 nm and with pipelike hallow morphology were prepared for photocatalytic dye degradation of rhodamine B and for water splitting through electrospinning assisted by calcination [44]. The nanowires obtained by calcination at 550 °C showed excellent photocatalytic efficiency than the samples obtained at 700 °C. Other examples of photocatalytic nanostructures synthesized by electrospinning method include ZnFe2O4 nanotubes and nanobelts for photodegradation of rhodamine B [45], lanthanum (La)-doped ZnO ceramic nanostructures for degradation of Congo red dye [46], and Au/ZnO nanostructures for organic pollutant degradation [47].

1.3.6  Pechini Method Pechini method is considered as a special case of sol–gel method used to obtain high degree homogeneous end products. In a typical method of Pechini process, a chelate is formed between the cationic salts or desired precursors that are dissolved in water and hydroxycarboxylic acid (citric acid is used mostly). Then, addition of polyalcohol and heating of obtained solution result in the esterification which leads to the formation of a gel kind of mixture. The final step involves drying of the gel followed by calcinations after which the desired nanomaterials are obtained. Gd2CoMnO6 nanostructures were synthesized using Pechini method for the photocatalytic dye degradation applications [48]. In a typical method, gadolinium nitrate (GdNO3) and citric acid were dissolved in distilled water followed by the addition of manganese nitrate hexahydrate Mn(NO3)2·6H2O and cobalt nitrate (CoNO3)2 under constant stirring. Then the solution was heated followed by the addition of propylene glycol after which a high viscous gel was obtained. Finally, the gel was dried in an oven followed by calcination where agglomerated nanoparticles with average size of 25–100 nm were obtained. Nd2Sn2O7 nanostructures with various morphologies such as less uniform nano-bundles, flake-like nanostructures, and uniform spherical nanoparticles were synthesized using different stabilizing agents for photodegradation of methylene orange under UV irradiation [49]. Zirconium nanosheets [50], sphere-like Cu2O/Li3BO3, and CuO/Li3BO3 nanocomposites with an average diameter size of 20 nm [51] and urchin-like Dy2CoMnO6 double perovskite nanostructures [52] were also synthesized through Pechini method for the purpose of dye degradation.

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1.3.7  Other Methods In the recent years, microemulsion technique has emerged as an interesting way to synthesize nanostructured photocatalysts owing to its promising advantage of controlling the particle size and morphology. Typically, two immiscible liquids, most preferably oil and water, are dispersed and stabilized with the addition of a surfactant. A transparent, isotropic, stable, and low-viscous system with water droplets will be obtained in which the water droplets are surrounded by the surfactant and stabilized. These water droplets act as nanoreactors for further chemical reactions and help to control the size of nanoparticles. For instance, ZnO-ZnWO4 nanoparticles were synthesized by emulsion method in which sodium tungstate dehydrate (Na2WO4·2H2O) and zinc (II) nitrate (Zn (NO3)2·4H2O) were used as metal precursors which were mixed in distilled water to form one microemulsion (sample-A) [53]. On the other hand, cyclohexane was used as oil agent, and brij35 and 1-butanol as surfactant and cosurfactant were mixed along with the addition of ammonium hydroxide (NH4OH) which acts as a precipitating agent to form the second microemulsion (sample-B). A precipitate was obtained with the addition of sample-B to sample-A in a dropwise manner which was then washed with ethanol and dried in oven and was finally calcined. The obtained nanoparticles exhibited hexagonal-­ shaped morphology with average diameter of 25 nm. Other examples of microemulsion method of synthesis includes development of TiO2- and Zr-doped TiO2 nanoparticles for methylene blue dedradation [54], spinel-type ferrite nanoparticles for photocatalytic water splitting [55], and Bi2MoO6 nanoparticles for dye degradation [56]. Sonochemical method is a facile simple approach for the synthesis of photocatalytic nanostructures. Ultrasonic waves are used for the chemical reactions that are carried out during the synthesis. For example, Cu/ZnO/Al2O3 ternary nano-hybrids were prepared via sonochemical method [57]. Firstly, 1 mmol of Zn(NO3)2·4H2O, Al(NO3)3·9H2O, and N2H4·H2O were dissolved in water and stirred for a certain period. The resulting solution was subjected to the ultrasonication for 10 min followed by the addition of Cu(NO3)2·6H2O solution in a dropwise manner. The ultrasonication was carried for another 5 min, and finally the resultant precipitant was annealed under vacuum, and the observed morphology was polyhedron microstructures with diameter size between 100 and 400 nm. ZnO-rGO nanocomposites [58], FeVO4 nanostructures [59], Tl4CdI6 nanostructures [60], and MnWO4/TmVO4 hybrid nanostructures [61] were also prepared with sonochemical method for various photocatalytic activities.

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1.4  Applications 1.4.1  Dye Degradation Various kinds of dyes are being released into water by different industries such as paper, textile, food processing, cosmetics, and paints. These dyes contaminate the water and are proven to be hazardous to environment and health. Photocatalysis has been employed for dye degradation applications from the past few decades, and various dyes such as acid dyes, basic dyes, direct dyes, pigment dyes, reactive dyes, etc. have been degraded by using different photocatalytic nanostructures. One such example was the use of CdS/ZnO heterojunction with nonuniform spherical-like structure, for the photodegradation of methylene blue in an aqueous solution under solar irradiation [62]. The degradation mechanism is shown in Fig.  1.5(c). Upon light irradiation, electron-­hole pairs are generated, and the excited electrons from the conduction band (CB) of CdS migrate to the CB of ZnO, but the transfer of holes does not take place as the valence band (VB) of CdS is more cathodic than ZnO. The electrons of

Fig. 1.5 (a) Photocatalytic degradation of methylene blue using pure ZnO and CdS/ZnO. (b) Recycling photocatalytic tests of CdS/ZnO. (c) Schematic representation of CdS/ZnO photocatalytic dye degradation

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ZnO produce superoxide radicals which react with holes and water producing large  number of hydroxyl radicals which are responsible for the degradation of methylene blue. The degradation efficiencies of pure ZnO and CdS/ZnO are shown in Fig. 1.5(a, b). After an irradiation time of 240 min, pure ZnO thin film showed an efficiency of 54%, whereas the CdS/ZnO heterojunction showed 91% degradation efficiency. The degradation efficiencies of various photocatalysts are provided in Table. 1.1.

1.4.2  Pharmaceutical Pollutant Degradation Rise in the pharmaceutical pollutants in the water is another major concern for the environment, and photocatalysis is believed to be one of the major tools for degradation of these dangerous pollutants. Magnetic ZnO@g-C3N4 heterojunction is one among the many photocatalysts that were explored for pharmaceutical pollutant degradation [74]. ZnO@g-C3N4 is employed for the degradation of sulfamethoxazole (SMX) antibiotic under UV irradiation (Fig. 1.6(a)). The proposed mechanism is shown in Fig. 1.6(b). The CB edge potentials of g-C3N4 and ZnO was found to be −1.12  eV and −0.31 eV, respectively, from which it is clear that the CB of g-C3N4 is more negative than the ZnO which leads to the migration of excited electrons from CB of g-C3N4 to ZnO and from there to Fe3O4 upon the UV light irradiation. From there, the electrons participate directly in pollutant degradation. As it is clear from Fig. 1.6(b) that the oxidation potential of holes in the VB of g-C3N4 is quite low to produce hydroxyl radicals, the redox potential of ZnO helps for the production of Table 1.1  Photocatalytic degradation of various dyes Photocatalyst MoS2-FeZnO WO3/g-C3N4 g-C3N4/MnV2O6 TiO2-CdO-Ag Cd-doped CeO2 Nd2Zr2O7

Dye MB MB MB/indigo carmine MB Rhodamine B/CR Erythrosine/ Eriochrome/black T Methyl orange Fe2V4O13/ZnO TiO2 Reactive black 5 CeO2 Reactive violet 1 SrTiO3/BiOI Crystal violet Ag3PO4@MWCNTs@ Malachite green Cr:SrTiO3 Fe@MoPO Malachite green TiO2/rGO Acid Red 14

Degradation time (min) 140 90 210/60 75 360 50

Efficiency of degradation (%) 95.2 95 95/94 92 86.42/92.53 88/84

Ref. [63] [64] [65] [24] [22] [66]

60 180 180 30 10

99 70 99.9 92.5 100

[67] [68] [69] [70] [71]

60 120

77 96.38

[72] [73]

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Fig. 1.6 (a) Degradation efficiency of sulfamethoxazole with various photocatalysts. (b) Pollutant degradation schematic representation

o2 e-

e-

Excitation

h+

e-

e-

o2- •

(b)

MTP Degradation products

Recombination

h+

h+

H2O or OH–

h

+

MTP •

MTP

OH

Degradation products

Degradation products

Concentration (C/Co)

(a)

26nm

1

53nm 0.8

80nm

0.6

106nm

0.4 0.2 0 0

30

60

90

120

Irradiation Time (min)

Fig. 1.7 (a) Schematic representation of degradation of metoprolol by photocatalysis. (b) Degradation efficiency by different sizes of TiO2 nanotubes

hydroxyl radicals, and holes can also participate directly in the degradation of pollutant or the intermediate products formed. Accordingly, 90.4% SMX pollutant degradation efficiency was found after 60-min irradiation of UV light which met the predicted value of 92.51% obtained by response surface methodology (RSM). Another example of photocatalysis application in pollutant degradation is TiO2 nanotube arrays which are used for the degradation of metoprolol under UV light irradiation [75]. Degradation mechanism is schematically represented in Fig. 1.7(a). The degradation of metoprolol was carried out by four different diameter sizes of TiO2 nanotubes. The degradation efficiency curves are represented in Fig. 1.7(b). Degradation efficiency was increased from 63.66% to 82.88% with increase in the nanotubes size from 26 to 53 nm, but further increase of nanotubes to 80–106 nm does not show any particular increase in the degradation efficiency. The reason for this is that as the diameter size of the nanotubes increases, the surface area decreases which affects the absorption ability of TiO2. Further different pharmaceutical pollutant degradation by various photocatalysts is showed in Table. 1.2.

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Table 1.2  various pollutant degradation efficiencies of different photocatalysts with respective degradation time Photocatalyst TiO2 Mg-doped ZnO-Al2O3 Pt-TiO2-Nb2O5

Pollutant Ibuprofen Caffeine

Diclofenac Ketoprofen Bi2WO6/Fe3O4/GSC Ampicillin Oxytetracycline Amoxicillin WO3 N-TiO2/rGO Tetracycline hydrochloride Tetracycline β-Bi2O3@g-C3N4 WO3–TiO2 @ Aspirin gC3N4 Caffeine Ibuprofen g-C3N4/Ag/AgCl/ BiVO4 Ag@BiPO4/BiOBr/ Norfloxacin BiFeO3 Tetracycline CuBi2O4/Bi2WO6 GO–Ag–ZnFe2O4 17 α-Ethinyl estradiol Ciprofloxacin RGO-Ce/WO3

Degradation time Efficiency of (min) degradation (%) 240 100 70 98.9 20 30 60

100

Reference [76] [77] [78]

95 94 99.9 98

[79]

[82] [83]

60

80.2 98 97 94.7

90

98.1

[85]

60 240

93 80

[86] [87]

60

100

[88]

180 60 50 90

[80] [81]

[84]

1.4.3  Plastic Degradation Plastic has been used widely in day-to-day life because of its various advantages such as durability, light weight, and cost-effectiveness. With enormous increase in the usage caused an enormous increase in disposal of this material which is becoming an environmental threat as the plastic is not biodegradable and also possesses hydrophobic nature which leads to the formation of fungi and microorganism on its surface. Even though strategies for plastic degradation such as recycling and burying are suggested, they also possess certain drawbacks that limited their implementations. Hence, the need for degradable plastic is a hot topic among the research fraternity which led the way to the exploration of photocatalysts for the development of degradable plastic. The incorporation of photocatalysts in the plastic materials will help for the degradation when exposed to irradiation as the radicals produced during the photocatalytic process will react with the organic materials and lead to oxidation and decomposition. In this strategy of using photocatalysts for  degradable plastic, TiO2 has been explored extensively due to its various properties  such as high stability, nontoxicity, high photocatalytic ability, and ­ inexpensiveness.

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In this direction, TiO2-based nanocomposites were used for photodegradation study of polypropylene (PP) under solar irradiation. Reduced graphene oxide (rGO)–TiO2 composites were used to form uniform PP/TiO2–rGO film whose synthesis is shown in Fig. 1.8(a, b) [89]. The mechanism of photocatalytic degradation is schematically represented in Fig. 1.8(c). As rGO is a good electron acceptor, the photogenerated electrons are transferred from the CB of TiO2 to CB of rGO which helps for the suppression of undesired recombination of electron-hole pairs. Also the surface area of the film has been increased with the incorporation of rGO which provides more active sites which in turn improves the photocatalytic activity. The FESEM characterization helped toward determining of the photocatalyst before and after photodegradation process, which is presented in Fig. 1.9. The upper section of the image represents the morphology of PP/TiO2 film, while the downward section is of PP/TiO2–rGO. Figure 1.9(a1) and (b1) is before irradiance, which shows a smooth surface, while in Fig. 1.9(a3), (a4), (b3), and (b4), it can be seen that there are deep cavities formed on the surface which show that the irradiation has a significant influence on the films and thus the degradation of polypropylene has been confirmed. It has been determined that the degradation efficiency of PP/TiO2–rGO is better than PP/TiO2. Another example is the use of ZnO for the degradation of low-density polyethylene (LDPE) under UV irradiation [90]. Firstly, ZnO nanoparticles were grafted with polystyrene to prepare ZnO composites, and then LDPE film with ZnO nanocomposites were prepared. The LDPE-ZnO film was exposed to UV irradiation for 200 h. The weight loss assessment and tensile strength of pure LDPE and grafted ZnO-LDPE for various irradiation times of 50, 100, and 200  min with different concentration of ZnO were investigated, which helped for the confirmation of photodegradation. With the increasing ZnO concentration, the weight loss also increased, but after a certain amount, there is a decrease which states that the right amount of ZnO incorporation should be done carefully.

Fig. 1.8 (a) Synthesis of TiO2–rGO nanocomposite. (b) Synthesis of uniform PP/TiO2–rGO film. (c) Schematic representation of photocatalytic degradation

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Fig. 1.9  FESEM images of PP/TiO2 and PP/TiO2–rGO before and after irradiance of light

Fig. 1.10  The appearance of film samples before and after underwent photo- and biodegradation studies [91]

Other examples include TiO2 nanoparticles in combination with biodegradable polymer polylactic acid for degradation of LDPE as shown in Fig. 1.10 [91], photodegradation of polypropylene-­ascorbic acid- TiO2 composite films [92], photodegradation of polyvinyl borate (PVB)-TiO2 nanocomposites prepared by condensation of polyvinyl alcohol and boric acid in the presence of TiO2 nanoparticles [93], and ZnO nanorods used for degradation of LDPE under visible light [94].

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1.4.4  CO2 Reduction Rapid consumption of fossil fuels leads to the increment of CO2 in the atmosphere which is one of the main contributors to the greenhouse effect. Photocatalytic CO2 conversion is a greener approach to convert CO2 into carbon-containing fuels to tackle this environmental issue. In this process, various photocatalysts have been exploited for conversion of CO2 into CO, HCOOH, CH3OH, and CH4. Table  1.3 represents photocatalytic conversion of CO2 into various fuels using various photocatalytic systems. A direct Z-scheme TiO2/CuInS2 nanostructure was used for the reduction of CO2 into CH4 and CH3OH [95]. Firstly, TiO2 nanofibers were synthesized by electrospun method followed by the development of TiO2/CuInS2 heterostructure through hydrothermal process. The CO2 reduction efficiency and mechanism are shown in Fig. 1.11(a, b). With light irradiation, electron-hole pairs are generated in the CB and VB of both TiO2 and CuInS2. TiO2 possess more negative CB edge potential than that of CuInS2. This results in the recombination of electrons from the CB of TiO2 with the holes in the VB of CuInS2. Electrons in the CB of CuInS2 with high reducibility are used for the reduction of CO2–CH4 and CH3OH, while holes in the VB of TiO2 with high oxidizibility react with water molecules. Type-II heterojunction-­ based ZnO/ZnSe composite is also used for the reduction of CO2 to methanol whose reduction mechanism and rate of reduction are shown in Fig. 1.12(a, b) [96]. The electrons from the CB of ZnSe transfer to CB of ZnO which reduced the CO2 molecules absorbed on the surface of photocatalyst, while the holes in the VB of ZnSe are used for oxidation of C3H8O to form C3H6O.

Table 1.3  Photocatalytic CO2 reduction by various photocatalysts Photocatalyst g-C3N4/FeWO4 α-Fe2O3/g-C3N4 ZnO/Ag1−xCux/CdS BVO/C/Cu2O Ag/Pd/TiO2 TiO2/Ti3C2 In2O3 coated with carbon In2O3/CeO2/HATP MoS2/TiO2 MXene/Bi2WO6 TiO2–MnOx–Pt

Final product CO CO CO CO CH4 CH4 CO/CH4 CO/CH4 CO/CH4 CH4/CH3OH CH4/CH3OH

Yield in μmol h−1 g−1 6 27.2 327.4 3.01 79 0.22 126.6/27.9 32.03/16.94 269.97/49.93 1.78/0.44 104/91

Reference [97] [98] [99] [100] [101] [102] [103] [104] [105] [106] [107]

Y. V. Divyasri et al.

2.5 (a)

CH3OH

2.0 1.5 1.0 0.5 0.0

T

(b)

CH4

TC1 TC2.5 TC5 TC10 Samples

-1 V vs NHE pH = 0 (eV)

Reduction of CO2 (µmol h-1 g-1)

18

TiO2 e-

Build-in electric field

e- e- e-

0

-

+

1

e-

+

2 3 H O 2 • OH

h+ h +

-

h+ + CulnS2

-

+

CO2 CH4 and CH3OH

Fig. 1.11 (a) CO2 reduction of TiO2/CuInS2 with various concentrations. (b) Z-scheme photocatalytic reduction mechanism

Fig. 1.12 (a) Photocatalytic CO2 reduction mechanism of ZnO/ZnSe. (b) CO2 conversion efficiency into methanol by various photocatalysts

1.4.5  N2 Fixation Nitrogen (N2) fixation through photocatalysis involves the reduction of N2 to ammonia (NH3), which plays a crucial role in the biological synthesis of agricultural fertilizers. Even though NH3 is synthesized through traditional Haber-Bosch process, disadvantages such as harsh reaction conditions and high energy consumption make it necessary to find a cost-effective and environmentally friendly approach for NH3 synthesis. Photocatalysis promises both these advantages and hence has been explored widely in the recent times. The reduction reaction for N2 fixation is similar to that of CO2 reduction, but the main limitation in the process of photocatalytic N2 fixation is the adsorption of N2 molecules onto the surface of the photocatalyst. Bismuth subcarbonate (BOC) with controllable defect density (BOC-X) was synthesized using sodium bismuthate (NaBiO3) and graphitic carbon nitride (g-C3N4) as precursors via hydrothermal process in order to overcome this drawback [108]. The photocatalytic N2 fixation is carried under three different irradiation

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Fig. 1.13 N2 fixation efficiencies of various BOC-X samples under (a) UV irradiation, (b) visible light irradiation, and (c) sunlight irradiation. (d) Schematic representation of photocatalytic N2 fixation mechanism under two different light sources

sources such as UV, visible light, and solar irradiation. The N2 conversion efficiencies with various concentrations of g-C3N4 deposited onto the defect-engineered BOC-X are shown in Fig. 1.13(a–c). It can be seen from Fig. 1.13(a) the samples showed highest NH3 yield under UV light irradiation. The photocatalytic mechanism under two different light sources is shown in Fig. 1.13(d). The defects created on the surface of photocatalysts provide oxygen vacancies which play a major role in the absorption of N2 molecules onto the surface, thereby increasing the photocatalytic efficiency. The excited electrons transfer from VB of BOC-X to the CB when the photon energy is higher than that of the bandgap of photocatalyst and participate in the conversion of N2 to NH3 directly. But when the photon energy is less than the bandgap, the excited electrons migrate to the defect levels where they participate in the N2 reduction reactions. There was a decrease in the NH3 yield after BOC-2 which states that the proper defect engineering is highly important. Too many of defects in the system will require higher photon energy for excitation of electrons, while less number of defects will result in the decrease in reduction ability of electrons, both of which are not favorable for better N2 fixation (Table 1.4).

1.4.6  Heavy Metal Reduction One of the main causes of water contamination is the continuous ingestion of heavy toxic metals like chromium (Cr), copper (Cu), manganese (Mn), nickel (Ni), lead (Pb), silver (Ag), and cadmium (Cd) into the water in the form of industrial wastage.

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Table 1.4 NH3 yield obtained by photocatalytic N2 fixation using various photocatalysts Photocatalyst ZnIn2S4/BiOCl

Yield of NH3 (μmol g−1 h−1) 14.6

A-CeOx

109

TiO2@C/g-C3N4

250.6

KOH-treated g-C3N4

3632

BiO quantum dots TiO2 nanotubes TiO2 – x g-C3N4 codoped with sulfur g-C3N4/Ag2CO3

1226 106.6 2.5 6.2 mg L−1 h−1 g−1

Au–Ag2O Ru/TiO2(P25) CdS nanorods-NiS

28.2 mg m−2 h−1 419 μmol L−1 g−1 1.7 mg L−1

11 mg L−1 h−1 g−1

Light source Visible light Visible light Visible light Visible light Sunlight Sunlight UV Visible light Visible light Sunlight UV–Vis Visible light

Irradiation time (min) 360

Ref. [109]

60

[110]

120

[111]

240

[112]

24 h 60 48 h 240

[113] [114] [115] [116]

240

[117]

60 480 60

[118] [119] [120]

Reduction of these heavy toxic metals into nontoxic or less toxic elements through photocatalysis is considered as a greener and cost-effective approach. Various photocatalysts are employed for the reduction of hexavalent chromium Cr (VI), one of the heavy toxic metals, to less toxic Cr (III) through photocatalysis. One such investigation involved the usage of CC@SnS2/SnO2 heterojunction for the reduction of Cr (VI) under visible light irradiation as shown in Fig. 1.14(a–d) [121]. Upon light irradiation, electrons and holes are generated in the CB and VB of SnS2, but no generation of charge carriers occurs in SnO2 as it is incapable of absorbing visible light. Excited electrons migrate to the CB of SnO2, while the holes remain in the VB of SnS2 leading to the suppression of recombination of charge carriers. Finally the reduction of Cr (VI) to Cr (III) was carried out by the reaction of electrons in the CB of SnO2 with the Cr (VI) absorbed on the surface, and the holes in the VB of SnS2 participate in the oxidation of H2O producing O2. Reduction activity was carried out for different CC@SnS2/SnO2 composites prepared at different calcination time. Highest photocatalytic activity was shown by C ­ C@SnS2/ SnO2-2 at pH-2, which clearly shows that the photocatalyst size and pH value play an important role in the reduction performance. As the size of the nanosheets increased, the absorption ability of the photocatalyst decreased leading to the poor reduction performance. Some of the photocatalysts developed for the heavy metal reduction are listed in Table 1.5.

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Fig. 1.14 (a) Cr (VI) removal efficiency of CC@SnS2/SnO2-2 at different pH values. (b) Removal efficiencies of different CC@SnS2/SnO2-x composites. (c) Cyclic performance of CC@SnS2/ SnO2-2. (d) schematic representation of photocatalytic reduction mechanism

Table 1.5  Heavy metal reduction efficiencies using various photocatalysts Photocatalyst ZnO

Fe3O4/RGO/ PAM

TiO2/graphene TiO2 nanofibers Nb2O5-CF Sulfur-doped g-C3N4 BiOBr/Ti3C2

Reduction efficiency% >85 40 dB

Absorption —

8.2–12.4 GHz

SE: 38 dB (59 ± 4 μm), 60 dB (133 ± 4 μm) RLmax: −33 dB at 13.6 GHz (Thickness: 2 mm) SET  =  77.87 dB EMI SE: 33.95–46.22 dB

—

Co/PPy (30 wt% in paraffin matrix (thickness: 2 mm) [107]

1–18 GHz

Ni/PPy (50 wt%) [108] [102] PANI/Co–Ni coating on fabrics [109] PANI‐PTSA/FeNi/epoxy resin [110]

8.2–12.4 GHz 8.2–12.4 GHz 8–18 GHz

RL:−22 dB (9.52 GHz) for 9.7 mm, RL:−20.7 dB (14.7 GHz) for 6.5 mm

—

— — —

PANI/RGO composite. PANI/graphene/amine functionalized MWCNT hybrid thin film exhibited highest conductivity of 0.11 S/cm and maximum EMI SE of 21 dB in the frequency range 12–18 GHz [118]. In another report, Au-MWCNT/ PANI nanocomposites prepared by in situ polymerization exhibited total shielding effectiveness of −16 dB [averaged over the X-band (GHz)] and a minimum reflection loss of −56.5  dB [119]. Such significantly high value of total SE could be explained considering its higher dielectric properties and the high electrical conductivity. Gupta et al. [120] modified PANI by natural graphite flakes (NGF) to improve the electrical conductivity followed by incorporating different weight fractions of MWCNTs to it. This modified PANI was ball milled in presence of MWCNTs for several hours to generate in situ graphene to form multiphasic composite. Figure  7.16 shows variation of frequency dependence of shielding effectiveness due to absorption (SEA), reflection (SER), total (SET) of composites on frequency and comparison of theoretical and experimental shielding effectiveness (SER and SEA) for the composites with respect to frequency. It is noted that total EMI-SE of this composite increases with increasing the MWCNTs content due to the

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Fig. 7.16  Dependence of shielding effectiveness due to (a) absorption (SEA), (b) reflection (SER), (c) total (SET) of composites on frequency and (d) comparison of theoretical and experimental shielding effectiveness (SER and SEA) for the composites with respect to frequency [120]

synergetic effect of functionalized MWCNTs, graphene and PANI. The multiphase composite attained maximum value of −98 dB at 10 wt% loading of MWCNTs with SET dominated by the absorption phenomena compared to the reflection. Further, theoretical and experimental values of SEA and SER are found to be in good agreement with one other. The highly conducting PANI coated MWCNTs prepared by in situ polymerization method exhibited EMI SE of 47.03 dB due to synergistic combination of the conductive components following absorption as dominant shielding mechanism [122]. Cao et  al. [123] compared microwave absorption and effective band width of wax composites comprising Fe3O4/MWCNT and PANI/Fe3O4/MWCNT. Figure 7.17a shows variation of reflection loss of Fe3O4/ MWCNT and PANI/Fe3O4/MWCNT in wax composites of different thicknesses in the frequency in the range of 2–18  GHz. It is clearly inevitable that maximum reflection loss in Fe3O4/MWCNT composites attained∼75 dB (3 mm in thickness). Figure 7.18b suggests that PANI/Fe3O4/MWCNT/wax to be less effective in microwave absorption compared to Fe3O4/MWCNT/wax composites. Such excellent microwave absorption performance of Fe3O4/MWCNT/wax could arise from the dielectric and magnetic loss contributions. Polystyrene microsphere/PANI/

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Fig. 7.17  Reflection loss plots of (a) Fe3O4/ MWCNT and (b) PANI/ Fe3O4/MWCNT wax composites [123]. Reproduced with permission from Amer Chem Soc

MWCNT [124], PANI/MWCNT/polystyrene [125] PANI/SWCNT [126], PANI/ graphene/MWCNT [120], para-toluene sulfonic acid-doped PANI-graphene nanoplatelets [127] and PANI/graphene [128] have also been investigated for their performance EMI shielding. PPY-MWCNT-Ag composites produced via UV-reduction showed higher electrical conductivity and SE in comparison to that formed by chemical reduction [101]. The effect of oxygen plasma treatment of MWCNT has been investigated on electromagnetic interference shielding of PPY-coated carbon nanotubes [129]. It was observed that average EMI SE of PANI deposited on MWCNT (plasma treated) increased to 28.3 dB compared to without oxygen plasma treatment (21.5 dB) and followed absorption as the main mechanism of EMI shielding. Kim and others [130] observed remarkable increases EMI SE (~28.6  dB) in PPy coated on oxyfluorination-­treated MWCNT.

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18

12

(A) PFF11 (A) PFF12 (A) PFF13 (A) PPy

10

SEA (dB)

15 (R) PFF11 (R) PFF12 (R) PFF13 (R) PPy

12 9

3

microwave conductivity (σs) (S/m)

(b)

24

13

14

15

16

Frequency (GHz)

17

18

21 18

(σ ) PPy s

2 3.5

(σ ) PFF11 s

(σs) PFF12 (σ ) PFF13 s

3.0 2.5

15 12

(δ) PFF11 (δ) PFF12 (δ) PFF13 (δ) PPy

9 6

2.0 1.5

3 0 12

6 4

6

12

8 SER (dB)

21

skin depth (δ) (mm)

(a)

1.0 13

14

15

16

Frequency (GHz)

17

18

Fig. 7.18 (a) Dependence of shielding effectiveness (SEA and SER) of PPy and PFF composites as a function of frequency, showing the effect of ferrofluid concentration on the SE value of the nanocomposites for sample thickness d ∼2.0 mm. (b) Variation of microwave conductivity and skin depth of PPy and PFF composites as a function of frequency [143]

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7.5.2  P  ANI/Graphene and PPy/Graphene Nanocomposites in EMI Shielding Applications PANI/graphene (free standing) hybrid films prepared by a solution intercalation showed reflection dominated SET achieving up to 42 dB (SER:32 dB) and 32 dB (SER: 24.2 dB) in the frequency range of 4–8 GHz and 8–12 GHz, respectively [131]. The contributions from reflection of 32 and 24.2 dB to this total EMI SE was found to be very high compared to that of absorption. It was suggested that high electrical conductivity and larger surface area of the hybrid film could account for the observed reflection dominant shielding mechanism. Modak et al. [132] also studied effect of functionalized graphene loading (1, 3 and 5 wt%) in polyaniline on its EMI SE. It was noted that EMI SE of nanocomposites increased in the frequency range of 2–12 GHz with increasing graphene loading in polyaniline due to high conductivity of graphene. The maximum value of EMI SE obtained in 5% (w/w) graphene loaded PANI composite is found to be between 51 and 52 dB in the above frequency range. Fe3O4/GNPs-NH-PANI) composites prepared by in situ polymerization/hydrothermal reaction possessed maximum RL value of absorption of −40.31 dB (thickness: 2.6 mm) [133]. The minimum reflection loss up to −60.3 dB (11.1 GHz) with sample thickness of 2.86 mm and the effective absorption bandwidth less than −10  dB (90% microwave absorption) was displayed by reduced graphene oxide covalently grafted polyaniline (30 wt%) in paraffin matrix [134]. It is anticipated that covalent bonds between reduced graphene oxide nanosheets and PANI nanorods facilitate the formation of the highly conductive networks and account for the absorption of electromagnetic wave. Hu et  al. [135] measured EMI SE of the samples in the frequency range of 300 KHz–3 GHz for 5, 10, 20 and 30 wt% of the PANI/CuS/RGO (20 wt% CuS/RGO) composites in paraffin wax (thickness:3 mm). It was observed that EMI SE of the samples improved to −18.0 dB for 20 wt% of the filler in wax. PPY-coated carbon fiber@graphene exhibited RLMin, corresponding to −45.12 dB (7.9 GHz) compared to bare carbon fiber@graphene of −30.53 dB (RLmin.:14.6 GHz) [136]. In another study, shielding effectiveness of sodium lauryl sulphate-doped polypyrrole (SLSDPPy), SLSDPPy-MWCNT, SLSDPPy-graphene and SLSDPPy-hybrid carbon composites Ku-band correspond to ~ –28.8, –40.6, –53.3 and –79.9 dB respectively [137]. The highest enhancement of SET in SLSDPPy-hybrid carbon composites could be attributed to the huge surface area offered by microporous hybrid carbon structures and presence of highly electrically conductive networks. The sponge-like PPy/RGO aerogel-­based composite at 10  wt% filler loading achieved effective electromagnetic absorption bandwidth (below −10  dB) of 6.76 GHz, and the highest reflection coefficient of −54.4 dB at 12.76 GHz [138]. Table 7.3 display EMI shielding performance data on nanocarbon/ICP nanocomposites

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Table 7.3  EMI shielding performance data on PANI/Nanocarbon, PPy/Nanocarbon, PANI/ Graphene, and PPy/Graphene Material PANI/MWCNTs (7 wt%) [114] β-naphthalene sulfonic acid (b-NSA)-doped PANI/Carbon fillers [115] PANI/MWCNT [117]

Range 8.2–12.4 GHz 8.2–12.4 GHz

12.4–18.0 GHz

PANI/Graphene/functionalized 12-18 GHz MWCNT [118] Au-MWCNT/PANI [119] 8–12 GHz MWCNT (10 wt%)–Graphene– PANI [120] Polyaniline-coated MWCNT/ maghemite [121]

PANI-coated MWCNT (oxyfluorinated) [122] Polystyrene microsphere/ PANI/MWCNT[124] PANI/SWCNT (25 wt%) [126] PANI/Graphene (33 wt%) [126] Para-Toluene Sulfonic Acid (p-TSA)-doped PANI– Graphene nanoplatelets (10 wt%) [127] PANI/Graphene (33 wt%) [128] COOH-MWCNT [101] MWCNT-PPy [101] PPy-MWCNT-Ag (Chem. red.) [101] PPy-MWCNT-Ag (UV-red.) [101] PPy-coated carbon nanotubes [129]

12.4–18 GHz

800 MHz–2.5 GHz

Electromagnetic interference Dominant Shielding performance mechanism SE: 60 dB Absorption SE: PANI/MWCNT: 37 dB), — SE: PANI/CF : 31 dB) SE: PANI/rGO : 39 dB) Shielding effectiveness: Absorption −27.5 to −39.2 dB EMI SE: 21 dB Absorption SET: −16 dB, Min. RL:−56.5 dB EMI SE: −98 dB SE: 34.1 dB

800 MHz–3 GHz EMI SE: 47.03 dB

SEA ≫ SER Absorption

Synergetic effect of reflection and absorption Adsorption

8.2–12.4 GHz

EMI SE; ∼22.7–23.2 dB

—

2–18 GHz 2–18 GHz

EMI SE: 31.5dB EMI SE: 34.2 dB

Absorption Absorption

8–12 GHz

−14.5 dB

Absorption

2–18 GHz

RLmax = −45.1 dB (Thickness: 2.5 mm) 21 dB 23 dB 27 dB 30 dB

—

4.7–7.7 GHz

800 MHz– t3 GHz

Average EMI shielding efficiency of PPy-coated MWCNTs increased from 21.5 to 28.3 dB

—

Absorption

(continued)

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Table 7.3 (continued) Material PPy-coated MWCNT(oxyfluorinated) [130] Polyaniline/Graphene [131]

Electromagnetic interference Dominant Range Shielding performance mechanism 800 MHz–3 GHz EMI SE: 28.6 dB Absorption

4–12 GHz

PANI/Functionalized graphene 2–12 GHZ (5%) [132] Covalently bonded GNPs-NH-­ 2–18 GHZ PANI nanorod modified by Fe3O4 [133]

rGO-g-PANI (30 wt%) in paraffin matrix [134]

Reflection Absorption

Absorption Max. RL: −40.31 dB (Thickness: 2.6 mm) and corresponding bandwidth with effective attenuation (RL  FC (Fe3O4@C: ∼41 − 20 dB > Fe3O4 : ∼15 dB. Such high value of shielding efficiency could be ascribed to the presence of dual interfaces and dielectric-magnetic integration in Fe3O4@C@Polyaniline. In other words, higher dielectric loss through interface polarization and relaxation effects in Fe3O4@C@Polyaniline could also contribute toward its superior microwave absorption ability. Alternatively, possibility of more and more electromagnetic energy converted to microcurrent, trilaminar Fe3O4@C@Polyaniline core-shell structure due to the increase in impedance matching also cannot be overruled. They also extended their work on Fe3O4@SiO2@PPy and observed highest total shielding efficiency (∼32 dB) for Fe3O4@SiO2/Pyrrole wt/wt = 1:9 and followed reflection as the dominant shielding mechanism [162]. Such performance was attributed to poor impedance matching between the PPy (conducting)/SiO2 (insulating), high electrical conductivity of Fe3O4@SiO2@PPy and presence of interface could account such enhancing the total shielding efficiency. Thus EM interference shielding in Fe3O4@SiO2@PPy and Fe3O4@C@PANI trilaminar core@shell nanocomposites is controlled by tuning of the shells through switching of the mechanism as shown in Scheme 7.4.

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Fig. 7.19 (a) Reflection loss of PPy/silane/MNPs prior to microwave heat treatment with a concentration of 70 wt% in paraffin wax [159]. Reproduced with permission from Springer. (b) Reflection loss of microwave-treated PPy/silane/MNPs with a concentration of 70 wt% in paraffin wax [159]. Reproduced with permission from Springer

Movassagh-Alanagh and others [163] preparedepoxy-based hybrid composites (thickness:1.5  mm) filled with 1  wt% of polyaniline@nano-Fe3O4@CFs (carbon fiber). It showed a maximum reflection loss (RL) value corresponding to −11.11 dB with an effective absorption bandwidth of about 6  GHz (frequency range:

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Fig. 7.20  Schematic Presentation of fabrication of Fe3O4@C@PANI ternary composite [161]. Reproduced with permission from ACS

Fig. 7.21 (a) Plots of frequency vs SEA, (b) frequency vs SER, and (c) frequency vs EMI SE of PFC composites [161]. Reproduced with permission from ACS

8.2–18  GHz). The reflection loss and maximum absorption of the N-doped graphene@ PANI nanorods@Fe3O4 hierarchical structures correspond to −10  dB (10.4–15.5 GHz) and −40.8 dB (14.8 GHz), respectively [164]. Wang et al. [165] observed superior microwave absorption N-doped graphen@polyaniline @ Fe3O4nanocluster compared to graphene@Fe3O4in the range of 2 and 18 GHz in all probability due to impedance matching and interfacial polarization.

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Scheme 7.4 Tuning of Shells in Trilaminar  Core@Shell  Nanocomposites in Controlling Electromagnetic Interference through Switching of the Shielding Mechanism [162]. Reproduced with permission from ACS

Xu et al. [166] reported that minimum reflection losses of Ni/PPy composites of Ni/Py ratio of 4:1 and 2:1correspond to −15.2  dB (13.0  GHz) and  −14.8  dB (14.4  GHz), respectively. They also noted electromagnetic absorption less than −10 dB in the 11–15.4 GHz and 12–17.5 GHz respectively due to the synergetic consequence of the Ni cores and PPy shells in Ni/PPY composite. Dong et al. [167] reported enhanced microwave absorption of Ni/PANI core-shell nanocomposites in the 2–18 GHzby dual dielectric relaxation and observed absorption properties less than −10 dB in the frequency range of 4.2–18GHz. Such performance is mainly ascribed to the EM matching (ferromagnetic Ni core/dielectric PANi shells) and the enhanced core/shell interfacial relaxation. Core/shell/shell γ-Fe2O3/microporous SiO2/polypyrrole microspheres (14.2 wt%) loaded in paraffin wax matrix (Thickness: 4.0  mm) showed maximum reflection loss of −51.24 dB (7.44 GHz) and effective absorption bandwidth (RL  GO [84]. The rGO has more preference to form π–π

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bonds with the aromatic VOCs. The adsorption capacity of these materials can be improved by increasing the specific surface area using activation. Molecular simulation studies revealed the suitability of modified CNTs for the adsorption of gas molecules. Fe-doped vacancy-defected CNTs were used for the adsorption of CO [85] and SO2 [86] molecules, whereas functionalized CNTs (hydroxyl, carboxyl, cyclodextrin) were found to be suitable for the adsorption of H2S molecules. In the case of H2S adsorption on cyclodextrin-functionalized CNTs, dispersion and electrostatic interactions were dominant [87]. Photocatalytic degradation is an effective method for the removal of gaseous pollutants even at trace levels. Carbonaceous nanomaterials can be used as a support for the dispersion of active semiconductors. They can provide adsorption sites for the gaseous pollutants to adsorb for subsequent degradation and can also modify the photogenerated charge-transfer process. Most of the degradation kinetics followed Langmuir–Hinshelwood equation. Based on the physico-chemical properties of the carbon nanomaterials, the photocatalytic efficiency varies. In the case of CNTs, the helicity and the diameter of the CNTs decide their conducting property (metallic, semiconductor). On the other hand, graphene is a conductor with zero bandgap that needs special treatment to convert the hydrophobic graphene into hydrophilic to deposit metal oxides on its surface to execute adsorption of gaseous molecules and subsequent photocatalytic degradation. Modification of graphene by doping and functionalization improves the adsorption-reactive sites on the surface. The degradation efficiency was quite prominent for rGO due to the strong interactions with the photocatalyst and less mass transfer limitations. In order to achieve high photocatalytic performance, the physico-chemical properties of the carbon nanomaterials such as surface area, porosity, active sites, hydrophobicity, hydrophilicity, pore volume, light absorption, charge-transfer, etc. should also be considered. The reaction parameters such as temperature, light irradiation, moisture, and concentration of pollutants have to be optimized. The VOCs tend to adsorb through van der Waals interactions, hydrogen bonding, hydrophobic effect, and π–π bonding [88]. Zhang et al. attempted to fabricate a novel monolithic protonated g-C3N4/GO aerogel for the photocatalytic removal of nitric oxide to nitrate with a removal ratio of 46.1%. The positive protonated g-C3N4 combined with the negative GO and enhanced the photocatalytic activity of the aerogel to oxidize even the secondary pollutant nitrogen dioxide to nitrate [89]. Both the adsorption and photocatalytic degradation of gaseous pollutants have their own limitations. In order to enhance the degradation efficiency, both adsorption and photocatalytic degradation can be integrated. The integrated adsorptive and photocatalytic method uses the adsorption process to adsorb the pollutants from the gas phase to the solid phase effectively followed by photocatalytic degradation of the adsorbed pollutants to complete mineralization.

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13.5  Carbonaceous Nanomaterials for Soil Remediation One of the important soil contaminants that need immediate attention is metalloids such as arsenic, mercury, chromium, lead, zinc, etc. The main sources are mining activities, industrial processes, application of contaminated domestic sludge, coal burning, and agrochemicals. The toxic level of metalloids depends upon its chemical form. For example, arsenite [As(III)] is more toxic than arsenate [As(V)], and organic mercury is more toxic than other forms of mercury. In soil, metalloids because of their mobile nature may be taken up by plants or transport to groundwater along with soil water. Some of the ways to immobilize metalloids are through sorption, precipitation, and complexation methods that depend on the soil properties and environmental factors. Organic contaminants such as pesticides, polycyclic aromatic hydrocarbons, polychlorinated biphenyls, organic solvents, and pharmaceuticals enter into the soil through anthropogenic activities. Most of them are toxic and persistent in soil and can undergo biomagnification to affect higher tropic levels. Carbon nanomaterials such as CNTs, graphene, and fullerene are some of the promising materials for soil remediation because of their high adsorption capacities, surface area, and hydrophobicity. Because of high hydrophobicity, CNTs prefer to adsorb hydrocarbons than alcohol. Based on the nature of the soil, the organic pollutants have variable affinity toward the soil. The source of antibiotics in the soil is mainly because of the incomplete metabolism of antibiotics in humans as well as animals, and their discharge to the waste stream. Eventually, the disposal of sewage sludge, poultry and livestock wastes, and discharge of municipal wastewater into the soil accounts for the high concentration of antibiotics in soil. The presence of antibiotics in soil disturbs the homeostasis of soil by inhibiting the growth of soil microorganisms, decreasing soil respiration, changing the turnover rate of soil nutrients, biogeochemical cycles, and specific enzyme activity. Some of the highly sensitive microorganisms may disappear, and the development of soil antibiotic-resistant bacteria and pathogenic antibiotic-­ resistant bacteria may emerge as a result. Even the sub-lethal level of antibiotics can increase the mutation rate and resistance genes in soil. Notable antibiotics that are found in soil are sulfonamides, tetracyclines, amoxicillin, oxytetracycline, ciprofloxacin, tylosin, cephalosporins, etc. High concentrations of antibiotics can affect the growth of plants and made their way to the edible part of the crop. Hence, it is mandatory to find a method to significantly reduce the concentration of antibiotics in soil. In soil, the degradation or transformation of antibiotics starts immediately upon application to soil, based on the nature and concentration of antibiotics, biotic, and abiotic conditions. The fate of antibiotics widely depends upon their physico-­ chemical properties and molecular structure. One of the important abiotic factors is hydrolysis, which plays a prominent role in the degradation of β-lactams, whereas it has an insignificant role in macrolides and sulfonamides. Another abiotic factor that has a significant role in the degradation of antibiotics is photodegradation. The

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physico-chemical characteristics of soil and climatic factors also determine the degree of antibiotics degradation. The sorption coefficient is an important parameter that decides the persistence of antibiotics and degradation rate in soil. Based on the presence of organic and inorganic soil components, the antibiotics tend to form complexes or bind strongly with soil components. When the sorption coefficient of antibiotics in the soil is less than 15  L/Kg, the antibiotics can easily move and degrade fast in soil, whereas the sorption coefficient greater than 4000 L/Kg leads to highly persistent and stable residues of antibiotics. It has been reported that due to high affinity toward soil components, antibiotics such as tetracyclines, macrolides, fluoroquinolones, and sulfonamides form stable residues. The pH of the soil also influences the adsorption and desorption characteristics of antibiotics. An increase in pH decreases adsorption of sulfonamides, due to the transformation of cationic form to neutral and anionic form. The main sorption mechanism in the case of the cationic form of antibiotics is electrostatic interactions through surface complexation, cation exchange, and cation bridging sorption. The half-life or DT50 indicates the persistent nature of antibiotics in different soil, which depends upon the organic carbon content in soil. Apart from abiotic degradation, microbial degradation of antibiotics by soil microbes such as Microbacterium, Ochrobactrum, Labrys etc. are also significant. Carbon-based materials are found to be biocompatible and act as a good sorbent for the removal of toxic substances from the environment. Graphene oxide has the capacity to bind with antibiotics, especially tetracycline and sulfamethoxazole. Zou et al. investigated the role of GO in inhibiting sulfamethoxazole uptake by bacteria and the transfer of antibiotic resistance genes (ARG) among microorganisms. GO readily forms a complex with sulfamethoxazole (GO-SMZ) and hinders the uptake of antibiotics in bacteria. The non-covalent combination of GO-ARG alters the properties of ARG [90]. Biochar is a solid carbonaceous material that originates from pyrolysis of biomass in the presence of insignificant amount of oxygen, with a pyrolytic temperature above 250 °C. It is considered to be a cheap and environmental-friendly material that has been used for carbon sequestration as well as for the improvement of soil quality because of its large specific surface area, surface functional groups, and porous structure. It is a good sorbent for the removal of pollutants from wastewater, soil and air. However, the physico-chemical properties of biochar vary with respect to the composition and chosen biomass. Carbonaceous nanomaterials have proved to be a good adsorbent, photocatalyst, and membrane filter for various environmental applications. In contrast, the fate of carbon nanomaterials is also a major concern as it is inevitably entering the environment. Due to the high stability of these materials, their degradation in nature is quite uncertain. The main degradation pathways are chemical degradation, photodegradation, and biodegradation. CNTs, graphene, and fullerene are reported to have cytotoxicity and found to influence oxidative stress response, mechanical damage, and biological enzymes. It is noted that at low concentrations, carbon nanomaterials have a positive influence on microorganisms. However, at high concentrations, they affect microbes by destroying the electrostatic equilibrium or cell membrane and

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organelles of microbial cells. Moreover, there will be changes in lipids and proteins of microbial cells due to the formation of oxidative stress, which often results in abnormal metabolism of microorganisms and further cell damage [91]. Similarly, low concentrations do enhance the vegetative growth and crop yield, by enhancing water uptake and transport, seed germination and growth, physiological activities, antioxidant activities, activating water channel proteins, and promoting nutrition absorption, whereas high concentrations reverse the effect [92]. The beneficial effects depend on the type, physical characteristics, and concentration of carbon nanomaterials along with the type of vegetation, application method, etc. In general, they have shown toxicity towards animals, human beings, plants, and microorganisms, as it is difficult to identify the thin line of demarcation between its beneficial and harmful concentrations. The studies on the toxicity of carbonaceous nanomaterials are in vivo and in its early stages, which needs a substantial amount of work to understand the degradation mechanism and toxicity of these materials [93].

13.6  Conclusion Carbonaceous nanomaterials have proven their ability to remove various pollutants from wastewater, air, and soil by means of adsorption, photocatalysis, and membrane technology. The physico-chemical properties of carbon-based nanomaterials can be engineered through functionalization or by combining different materials to form nanocomposites or nanohybrids in order to improve their performance based on the target pollutant. Carbonaceous nanomaterials like CNTs, graphene, GO, rGO, fullerene, etc. are ideal for adsorbing both organic pollutants and inorganic metals. The high specific surface area, adsorption capacity and the flexibility to tune the physico-chemical properties make these nanomaterials a competent adsorbent. The interactions with the adsorbent vary with the target pollutant. Organic pollutants (antibiotics, pesticides, dyes) in water interact with the carbon nanomaterials through van der Waals forces, π–π interactions, hydrophobic interactions, and electrostatic interactions, whereas inorganic metals use electrostatic interactions, ion exchange, and surface complexation, dissolution, co-precipitation, and surface precipitation to get adsorb. Adsorption isotherms are studied to understand the interaction between the adsorbent and adsorbate, which helps to design an effective treatment system based on the interpretations of the capacity of adsorbents, surface properties, adsorption mechanism, and adsorption phenomenon. Adsorption kinetics reveals the dynamics of the adsorption process by providing information about the rate of adsorption and time taken to attain equilibrium along with probable adsorption pathways and mechanisms. The pseudo-second-order kinetics holds good for most of the adsorption of organic pollutants. Gaseous pollutants are removed by adsorption, photocatalysis, and membrane filtration. Integrating both adsorption and photocatalysis techniques provides a promising technology to degrade gaseous pollutants efficiently. Soil reclamation by employing carbon-based

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nanomaterials needs in-depth study to understand the fate of the adsorbents in the soil. In order to design a novel adsorbent, it is essential to understand the interactions between the adsorbent and adsorbate. Some of the current research focused on investigating at a microcosmic level using computational tools in order to screen a wide range of pollutants. From a practical standpoint, the difficulty encountered in synthesizing these nanomaterials in large scale economically to meet out the demand in real industries limits their applications. There is still a wide gap between the laboratory scale to real commercialization and practical applications. The prime focus of research should be on fabricating a commercially viable carbon-based hybrid material with recycling ability to make it economically cheap and efficient.

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57. Fiyadh, S. S., AlSaadi, M. A., Jaafar, W. Z., AlOmar, M. K., Fayaed, S. S., Mohd, N. S., Hin, L. S., & El-Shafie, A. (2019). Review on heavy metal adsorption processes by carbon nanotubes. Journal of Cleaner Production, 230, 783–793. https://doi.org/10.1016/j.jclepro.2019.05.154. 58. Ouni, L., Ramazani, A., & Taghavi Fardood, S. (2019). An overview of carbon nanotubes role in heavy metals removal from wastewater. Frontiers of Chemical Science and Engineering, 13(2), 274–295. https://doi.org/10.1007/s11705-­018-­1765-­0. 59. Yang, X., Wan, Y., Zheng, Y., He, F., Yu, Z., Huang, J., Wang, H., Ok, Y. S., Jiang, Y., & Gao, B. (2019, February). Surface functional groups of carbon-based adsorbents and their roles in the removal of heavy metals from aqueous solutions: A critical review. Chemical Engineering Journal, 366, 608–621. https://doi.org/10.1016/j.cej.2019.02.119. 60. Singha Deb, A. K., Dhume, N., Dasgupta, K., Ali, S. M., Shenoy, K. T., & Mohan, S. (2019). Sulphur Ligand Functionalized Carbon Nanotubes for Removal of mercury from waste water– experimental and density functional theoretical study. Separation Science and Technology (Philadelphia), 54(10), 1573–1587. https://doi.org/10.1080/01496395.2018.1529044. 61. Ghorbani, M., Seyedin, O., & Aghamohammadhassan, M. (2020, October). Adsorptive removal of lead (II) ion from water and wastewater media using carbon-based nanomaterials as unique sorbents: A review. Journal of Environmental Management, 254, 109814. https://doi. org/10.1016/j.jenvman.2019.109814. 62. Verma, B., & Balomajumder, C. (2020). Surface modification of one-dimensional Carbon Nanotubes: A review for the management of heavy metals in wastewater. Environmental Technology and Innovation, 17, 100596. https://doi.org/10.1016/j.eti.2019.100596. 63. Hayati, B., Maleki, A., Najafi, F., Daraei, H., Gharibi, F., & McKay, G. (2017). Super high removal capacities of heavy metals (Pb2+ and Cu2+) using CNT dendrimer. Journal of Hazardous Materials, 336, 146–157. https://doi.org/10.1016/j.jhazmat.2017.02.059. 64. Liang, J., Liu, J., Yuan, X., Dong, H., Zeng, G., Wu, H., Wang, H., Liu, J., Hua, S., Zhang, S., Yu, Z., He, X., & He, Y. (2015). Facile synthesis of alumina-decorated multi-walled carbon nanotubes for simultaneous adsorption of cadmium ion and trichloroethylene. Chemical Engineering Journal, 273, 101–110. https://doi.org/10.1016/j.cej.2015.03.069. 65. Geng, J., Yin, Y., Liang, Q., Zhu, Z., & Luo, H. (2019). Polyethyleneimine cross-linked graphene oxide for removing hazardous hexavalent chromium: Adsorption performance and mechanism. Chemical Engineering Journal, 361, 1497–1510. https://doi.org/10.1016/j. cej.2018.10.141. 66. Elgengehi, S. M., El-Taher, S., Ibrahim, M. A. A., Desmarais, J. K., & El-Kelany, K. E. (2020). Graphene and graphene oxide as adsorbents for cadmium and lead heavy metals: A theoretical investigation. Applied Surface Science, 507, 145038. https://doi.org/10.1016/j. apsusc.2019.145038. 67. Yao, M., Wang, Z., Liu, Y., Yang, G., & Chen, J. (2019, February). Preparation of dialdehyde cellulose graftead graphene oxide composite and its adsorption behavior for heavy metals from aqueous solution. Carbohydrate Polymers, 212, 345–351. https://doi.org/10.1016/j. carbpol.2019.02.052. 68. Kong, Q., Preis, S., Li, L., Luo, P., Hu, Y., & Wei, C. (2020). Graphene oxide-terminated hyperbranched amino polymer-carboxymethyl cellulose ternary nanocomposite for efficient removal of heavy metals from aqueous solutions. International Journal of Biological Macromolecules, 149, 581–592. https://doi.org/10.1016/j.ijbiomac.2020.01.185. 69. Ahmad, S. Z. N., Wan Salleh, W. N., Ismail, A. F., Yusof, N., Mohd Yusop, M. Z., & Aziz, F. (2020). Adsorptive removal of heavy metal ions using graphene-based nanomaterials: Toxicity, roles of functional groups and mechanisms. Chemosphere, 248, 126008. https://doi. org/10.1016/j.chemosphere.2020.126008. 70. Abdelsalam, H., Teleb, N. H., Yahia, I. S., Zahran, H. Y., Elhaes, H., & Ibrahim, M. A. (2019). Graphene-based adsorbents for the removal of toxic organic pollutants: A review. Journal of Physics and Chemistry of Solids, 130, 32–40. https://doi.org/10.1016/j.jpcs.2019.02.014.

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Chapter 14

Magnetically Recyclable Photocatalysts for Degradation of Organic Pollutants in Aquatic Environment Ashutosh Kumar and Sushil Kumar Kansal

14.1  Photocatalysis Photocatalysis—a light-driven catalysis process—employed for degradation of organic pollutants through activation of a semiconductor material, i.e., photocatalyst upon light irradiation. Basically, the degradation of organic pollutants through photocatalysis process can be achieved in the following five steps (Fig. 14.1) [1, 2]: 1 . Adsorption of organic pollutant on to the photocatalyst’s surface. 2. Absorption of light energy by the photocatalyst, resulting in the photogeneration of electrons (e−) and holes (h+). The e−–h+ pair can be generated, if the light energy is more than or equal to the band gap (Eg) of a photocatalyst. 3. Separation of e−–h+ pair. 4. Generation of reactive oxygen species (ROS). 5. Degradation of organic pollutant adsorbed onto the photocatalyst’s surface through oxidation–reduction reaction carried out by the e−, h+, and ROS.

A. Kumar (*) School of Energy and Environment, Thapar Institute of Engineering and Technology, Patiala, Punjab, India e-mail: [email protected]; [email protected] S. K. Kansal (*) Dr. S. S. Bhatnagar University Institute of Chemical Engineering and Technology, Panjab University, Chandigarh, India e-mail: [email protected] © Springer Nature Switzerland AG 2021 S. Balakumar et al. (eds.), Nanostructured Materials for Environmental Applications, https://doi.org/10.1007/978-3-030-72076-6_14

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Fig. 14.1 Diagram showing degradation of organic pollutant in a typical photocatalysis process

Fig. 14.2  A typical diagram showing generation of ROS onto the photocatalyst’s surface

14.1.1  Generation of ROS In fact, upon light irradiation onto the photocatalyst’s surface, the e−–h+ pairs are generated. By virtue of their opposite charge, the e− and h+ have tendency to recombine each other which is known as recombination process. Furthermore, the efficiency of the photocatalysis process can be enhanced by lowering the recombination rate of e− and h+, which can provide more time for the generation of ROS. Accordingly, the low rate of recombination of e− and h+ provides relatively more time for the generation of ROS through the respective oxidation–reduction reactions with the reductive e− and oxidative h+, which ultimately results in relatively more number of ROS available for the pollutant degradation through the photocatalysis process. A typical diagram showing generation of ROS onto the photocatalyst’s surface is shown in Fig. 14.2. In a typical photocatalysis process, the oxidative photogenerated h+ present in the valence band (VB) reacts with the H2O present in the aquatic environment to form the hydroxyl radical (•OH). Similarly, the reductive photogenerated e− present in the conduction band (CB) reacts with the O2 present in the aquatic environment to form the superoxide radical (•O2). Due to further protonation of the •O2, it can also form the hydrogen peroxide (H2O2) which can subsequently generate the •OH. It is to be noted that the •OH is the most powerful and

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non–selective ROS in the photocatalytic degradation process which has ability to degrade almost all types of organic pollutants [2, 3].

14.1.2  Challenges in Photocatalysis Ideally, photocatalysis process leads to mineralization of organic pollutants into CO2 and H2O, which are relatively less harmful products than their reactants, i.e., organic pollutants. However, the basic challenges of the photocatalysis process limit its application which can be due to (a) the activity of the photocatalyst in relatively higher energy or lower wavelength of light present in solar spectrum (e.g., UV) only; (b) the high recombination rate of e−–h+ pair, providing relatively low residence time for the generation ROS in the solution which in turn produces relatively lower number of ROS; and (c) dissipation of the captured light energy though recombination of e−–h+ pair [2, 4]. In order to understand the limited activity of the photocatalyst in the lower wavelength of light, the solar spectrum is shown in Fig.  14.3. As shown in Fig.  14.3, when one moves from the UV region toward the visible light or infrared (IR) region, the light energy has relatively lower energy but higher wavelength. It depicts that activity in lower wavelength region requires higher energy and thus the higher cost of photocatalysis process from its application point of view. Moreover, the UV light shares only 2–3% of the solar radiation, while the visible light accounts to ~45% of the total solar radiation [6, 7]. Therefore, designing a photocatalyst which can be activated by (a) the major share of solar spectrum, i.e., visible light, or (b) by most of the part of solar radiation, i.e., by UV, visible light, and IR together, can overcome one of the major limitations of the photocatalysis process [8, 9]. The high rate of recombination of e−–h+ pair is another major limitation of the photocatalysis process. Briefly, the recombination can be overcome by adopting any of the Fig. 14.3  Solar spectral irradiance at air mass 1.5 (AM 1.5) collected from American Society for Testing and Materials (ASTM). Standard: ASTM G-173-03. Reproduced with permission from Elsevier [5]

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following strategies: (a) by developing a heterojunction [2, 10], (b) by regulating a defect or vacancy in the photocatalyst [11, 12], or (c) by doping a photocatalyst with metals or non-metals [7, 13, 14]. It is to be noted that the challenges of the photocatalysis process is not the major focus of this chapter, and thus only important limitations vis-à-vis challenges are discussed.

14.2  Organic Pollutants in Aquatic Environment Aquatic environment, i.e., water and wastewater, contains various types of organic pollutants. Any undesirable chemical substance of organic nature which can potentially affect the human health are categorized as an organic pollutant. From the literature, it is observed that the researchers have targeted various groups of organic pollutants for degradation, such as dyes [15, 16], endocrine disrupting chemicals (EDCs) [17, 18], pharmaceuticals and personal care products (PPCPs) [19–21], persistent organic pollutants (POPs) [22, 23], poly-and per-fluoroalkyl substances (PFAS) [24, 25], emerging chemicals [26, 27], etc. PFAS are a group of anthropogenic substances with amphiphilic properties which are used in our day to day life as fire suppressants, stain and water repellants. After their use, the PFAS find their ways into the aquatic environment and thereafter to the human being and animals, after consuming such water. The PFAS can cause adverse immunological effects, reproductive disorders, liver ailments in animals, and immuno-toxicological and neuro-developmental issues in children [24, 25]. PPCPs represent a group of chemicals containing pharmaceuticals, viz., analgesics, antibiotics, anti-inflammatory, anti-epileptic drugs, etc., and personal care products, viz., insect repellants, synthetic musks, UV filters, etc. The adverse effects of these chemicals on aquatic lives include birth defects, endocrine disruptions, oxidation stress, post-embryonic developmental issues, and reproductive disorder [7, 20, 21]. Dyes are another class of visibly recognized harmful pollutants which are discharged into water bodies by various industries, viz., food, hair dyeing, leather tanning, paper and pulp, photographic, printing, textile, etc. The dyes and their toxic metabolites are known to cause bladder cancer in human being [14, 28]. EDCs are a group of exogenous chemical substances which can alter the function of endocrine system and consequently causes adverse effects in an organism or its progeny. These chemicals are brominated flame retardants, dioxins, bisphenol A, organic solvents, parabens, phthalates, pesticides, polycyclic aromatic hydrocarbons (PAHs), heavy metals, some naturally occurring phytoestrogens, etc. The EDCs can potentially cause developmental abnormalities and reproductive issues in aquatic lives, birds, and animals [17, 29]. POPs are a group of chemicals which remain in the environment for longer period, and hence they can bioaccumulate and transported into the food chain. The POPs include chemicals, such as herbicides, insecticides, fungicides, pesticides, etc., which are potentially harmful to the environment and human health [3, 23]. Considering adverse effects of these organic pollutants on human

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being and aquatic environment, their effective treatment becomes a quintessential task. Wastewater treatment plants (WWTPs) are considered as the largest sink of these organic pollutants in aquatic environment. In fact, due to inefficient treatment of these organic pollutants by WWTPs, they find their ways into receiving water bodies and thereafter they become part of the aquatic environment [2]. Most of these pollutants are not efficiently treated by the conventional water and wastewater treatment processes, and thus they again become part of the aquatic environment. Thus, due to their inefficient treatment and regular discharge in the aquatic environment, their concentration is expected to increase continuously and so is their scale of related acute and chronic toxicity to the human being and aquatic lives. Alternatively, photocatalysis, being the effective treatment process for degradation of these organic pollutants, has invited immense attention from the researchers across the globe. The key features of photocatalytic degradation process, its mechanism of generation of ROS, and its basic challenges in degradation of organic pollutants are already discussed in Sect. 14.1 of this chapter.

14.3  Magnetically Recyclable Photocatalysts Recycling process consists of separation, regeneration, and reuse of a photocatalyst. The successful recycling of a photocatalyst can potentially reduce the total cost of the photocatalysis process by repetitive use of the same photocatalytic material. Generally, the repetitive use of photocatalytic material can be performed by adopting one of the following two strategies: (a) by immobilizing the photocatalyst on a stable support, such as clay, glass, silica, zeolite, etc. [30–32], or (b) by suspending the photocatalysts in a suspension solution and its recovery after the photocatalytic degradation process [33–35]. However, the use of former strategy can potentially reduce the overall surface area of the photocatalyst which can in turn affect the photocatalytic performance. Thus, the application of photocatalysts in a suspension solution and their recovery after the photocatalytic degradation process has become a relatively more popular strategy than immobilizing the photocatalyst on a stable support. In recycling process, the separation of photocatalysts from the suspension solution is usually conducted by adopting the ways, such as (a) centrifugation or (b) magnetic separation. In laboratory, the centrifugation is performed by utilizing a centrifuge wherein the photocatalysts and solvents are separated by applying the mechanical force (centrifugal force). However, the magnetic separation of photocatalysts in the laboratory is performed either by employing a handheld laboratory magnet or by employing an electromagnetic separation system [2, 36]. Due to the easy handling of magnetic separation process, it has attracted immense attention among the researchers. Accordingly, the magnetic photocatalysts were developed in various ways by adopting different strategies. On the basis of the role of magnetic material in a photocatalyst or composite photocatalyst, it can be divided into two

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broad categories: (a) magnetic material as a photocatalyst, (b) magnetic material as a non-photocatalyst [37, 38]. At large scale, the magnetic separation of magnetic photocatalysts is considered to be a challenging task. In industries, the magnetic separation of magnetic nanoparticles can be performed by applying two types of magnetic separators, such as high gradient magnetic separator (HGMS) and open gradient magnetic separator (OGMS). In HGMS, the separator column provides a high magnetic field gradient which enables the efficient separation of magnetic nanoparticles, and thus the outlet of the separator provides a clean solution. It should be noted that the gradient inside the separator is generated by the ferromagnetic matrix. However, clogging of the filter by separated magnetic material is the major issue faced by such separators [39, 40]. Alternatively, in OGMS, the magnetic system is arranged in a convenient way around the wall of the separator in switch mode. When the magnetic system is switched on, the magnetic photocatalysts can be retained around the wall, and the separated solution can be obtained at the outlet [40, 41]. The advantage of the OGMS over HGMS is that it has high separation efficacy while having lower gradient of the magnetic field. Therefore, at large scale, magnetic separators can be selected by considering the pros and cons of these two magnetic separators. After photocatalytic reaction, regeneration of photocatalysts is considered to be a major challenge. However, regeneration of magnetic photocatalysts after their magnetic separation can potentially reduce the total cost of the photocatalytic degradation process by reusing the same material, i.e., photocatalyst for multiple cycles. Recently, Kumar et al. [2] have reviewed the various regeneration methods applied to regenerate the photocatalysts. After reviewing the various regeneration methods, it was found that the regeneration was achieved either by washing with regeneration solution, such as de-ionized (DI) water, organic solvents, H2O2, or by treatment with gases, such as O3. Among these regeneration methods, it was found that washing with DI water as well as ethanol followed by drying overnight at 60°C is the most effective and easy to handle regeneration method. Therefore, in recycling process, separation and regeneration methods play a crucial role.

14.3.1  Magnetic Material as Photocatalyst The magnetic material used as a photocatalyst in a composite photocatalyst possesses two functions: (a) involves in photocatalytic reactions and (b) enables magnetic separation of the whole composite photocatalyst due to its magnetic property. From the literature, it is evident that the magnetic material utilized for the photocatalysis process can play the various roles, viz., as a (a) catalyst, [42, 43] (b) co-­ catalyst [12], or (c) dopant [44]. It is to be noted that the magnetic materials utilized in photocatalytic role are mostly iron-based materials, such as FeO, Fe2O3, Fe3O4, ferrites, etc. Among the various phases of iron oxides, FeO, Fe2O3, and Fe3O4 (magnetite) are the most frequently used magnetic materials in the magnetic photocatalysis process. The Fe (III) oxide exists in various forms, such as α–Fe2O3, β–Fe2O3,

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γ–Fe2O3, and ε–Fe2O3. Among these phases of Fe (III) oxide, α–Fe2O3 (hematite), and γ–Fe2O3 (maghemite) are the more stable forms, and thus they are frequently used in the photocatalysis process. Moreover, ferrite (MFe2O4) is a transition metal oxide-based spinal structure in which M represents Co, Fe, Mn, Ni, Zn, etc. [45]. The Fe2O3 possesses a band gap of ~2.3 eV which makes it a visible light active photocatalyst. It is to be noted that the α–Fe2O3 is found to be stable in the aqueous solution of pH > 3 which makes it a suitable candidate for the photocatalytic degradation of dyes in aqueous solution [42]. Morphology manipulation is another approach which can further enhance the photocatalytic efficiency of a photocatalyst. In order to improve the photocatalytic efficiency of the α–Fe2O3 photocatalyst, Zhang et al. [43] developed various morphology of α–Fe2O3, such as nanoparticles, mesoporous nanorods, and microplates. The scanning electron microscope (SEM) images of the various morphology of α–Fe2O3 are shown in Fig. 14.4 which are the as-synthesized nanoparticles (Fig. 14.4a), mesoporous nanorods (Fig. 14.4b), and microplates (Fig.  14.4c) of α–Fe2O3. Accordingly, the visible light-driven photocatalytic activity of the various morphology of α–Fe2O3 was tested for degradation of methylene blue (MB) dye (Fig. 14.5). As shown in Fig. 14.5, among the various morphology of α–Fe2O3 photocatalysts, mesoporous α–Fe2O3 nanorods have shown relatively better photocatalytic performance than the other morphologies for degradation of 100 mL MB of concentration 10−5 M. Moreover, in the presence of H2O2, the activity of the mesoporous α–Fe2O3 nanorods was found to be further improved than the mesoporous α–Fe2O3 nanorods alone. After photocatalytic degradation of MB, the α–Fe2O3 nanorods were separated, regenerated, and employed again for the degradation of MB under visible light. From the recycling experiments, it was found that over 82% of photocatalytic efficiency of the mesoporous α–Fe2O3 nanorods was maintained after 5 cycles. However, the photocatalytic activity of the α–Fe2O3 was still found to be restricted by the common challenges of photocatalysis, i.e., the high rate of e−–h+ recombination. Therefore, efforts were made to combine such photocatalyst with another photocatalyst which were expected to lower the rate of recombination of e−–h+ pair, resulting in improved photocatalytic performance. The magnetic photocatalysts, viz., Fe2O3 and Fe3O4, have visible light absorption ability. However, their main limitation as a photocatalyst is the high recombination rate of charge carriers. Accordingly, efforts were made to combine such magnetic photocatalysts with other semiconductor materials with superior photocatalytic ability, such as BiOBr, BiVO4, g–C3N4, graphene oxide, SnO2, TiO2, ZnO, etc. [2, 45, 46]. In such cases, the magnetic photocatalyst acts as a co–catalyst in a composite photocatalyst, resulting in the low recombination rate of charge carriers and generation of more number of ROS which are ultimately responsible for photocatalytic degradation of organic pollutants. For instance, γ–Fe2O3 nanosheets, a 2D graphene–like nanostructure, have excellent electrochemical and optical properties. However, in photocatalysis process, the γ–Fe2O3 nanosheets suffer from the limitations, such as short life time of photogenerated charge carriers (~10 ps) and their low diffusion length (2–4  nm) [47]. The TiO2 has excellent corrosion resistance, chemical inertness, low toxicity, and high redox ability. However, due to its high

372 Fig. 14.4  SEM image of α–Fe2O3 (a) nanoparticles, (b) mesoporous nanorods, and (c) microplates. Reproduced with permission from Elsevier [43]

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Fig. 14.5 Photodegradation efficiencies of MB solutions on various α–Fe2O3 photocatalysts. Reproduced with permission from Elsevier [43]

Fig. 14.6 (a) UV–vis spectra, and (b) Magnetization hysteresis loops of α–Fe2O3/b–TiO2 and γ– Fe2O3/b–TiO2 heterojunctions. Reproduced with permission from Elsevier [12]

band gap (~3.2  eV), it can be activated upon irradiation of UV light only [2]. Alternatively, the black TiO2 (b–TiO2) can show an extended light absorption of the solar spectrum. The b–TiO2 is known to absorb the light up to the near infrared (NIR) region, resulting in a higher photogeneration of charge carriers, and thus an improved photocatalytic performance [48]. Considering the limitations of the γ–Fe2O3 nanosheets and superiority of the b– TiO2 in the photocatalysis process, Ren et al. [12] developed a γ–Fe2O3/black TiO2 heterojunction photocatalyst for the solar light-driven photocatalytic degradation of a PPCP, i.e., tetracycline. Briefly, the magnetic ultrathin γ–Fe2O3 nanosheets were hybridized with the mesoporous black TiO2 (b–TiO2) hollow spheres to form a γ– Fe2O3/b–TiO2 heterojunction photocatalyst, wherein the γ–Fe2O3 acted as a co-­ catalyst. In order to show extended light absorption in the solar spectrum and improved magnetic recyclability performance of the heterojunction photocatalyst, Ren et al. [12] synthesized two types of heterojunction photocatalyst with the same method, i.e., α–Fe2O3/b–TiO2 and γ–Fe2O3/b–TiO2 heterojunction photocatalysts. The UV–vis absorption spectra and saturation magnetization (Ms) values obtained from the the vibrating sample magnetometry (VSM) analysis are shown in Fig.  14.6a, b, respectively. As shown in Fig.  14.6a, the γ–Fe2O3/b–TiO2

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heterojunction photocatalyst has shown the improved visible light absorption compared to the α–Fe2O3/b–TiO2 heterojunction photocatalyst. Moreover, the γ– Fe2O3/b–TiO2 heterojunction photocatalyst has also shown absorption in the NIR region of the solar spectrum. From Fig. 14.6b, the Ms value of α–Fe2O3/b–TiO2 and γ–Fe2O3/b–TiO2 heterojunction photocatalysts were found to be 0.68 and 8.17 emu/g, respectively, which depicts the improved magnetic behavior of the γ–Fe2O3/b–TiO2 heterojunction photocatalyst compared to the α–Fe2O3/b–TiO2 heterojunction photocatalyst. Thus, by employing the γ–Fe2O3 as a co–catalyst, a γ–Fe2O3/b–TiO2 heterojunction photocatalyst with extended light absorption and improved magnetization value was developed. The schematic diagram of the band structures and charge transfer mechanism in the γ–Fe2O3/b–TiO2 heterojunction photocatalyst are shown in Fig. 14.7. From the migration of the photogenerated charge carriers onto the heterojunction photocatalyst’s surface, it is depicted to be a type–II heterojunction photocatalyst, wherein the h+ generated in the VB of γ–Fe2O3 migrates toward the VB of b–TiO2 and the photogenerated e− in the CB of TiO2 migrates toward the CB of γ–Fe2O3 nanosheets. Such migration of the charge carriers, i.e., e− and h+ in this type–II heterojunction photocatalyst, results in relatively low rate of recombination of charge carriers which ultimately leads to an improved photocatalytic efficiency. Thus, by employing γ–Fe2O3 as a co-catalyst in a heterojunction, their photocatalytic efficiency can be ultimately improved along with their attractive feature of magnetic recyclability. Doping of photocatalyst is another strategy for improving the photocatalytic efficiency of the semiconductor photocatalyst that can be achieved either by extending the light absorption ability of the photocatalyst or by lowering the e−–h+ recombination rate. For instance, TiO2 is a stable ideal semiconductor which is activated only after irradiation of UV light. The doping in the TiO2 lattice can be achieved by adopting the three different strategies, which are shown in Fig. 14.8. As shown, the main rationale behind the various strategies of doping in TiO2 is to decrease its wide band gap which can be achieved either by a lower shift of conduction band minimum (CBM) (Fig.  14.8a), a higher shift of valence band maximum (VBM) (Fig. 14.8b), or creating the impurity states (Fig. 14.8c).

Fig. 14.7 Schematic illustration of the energy band structure for γ–Fe2O3/b–TiO2 heterojunctions and the proposed photogenerated charge transfer mechanism. Reproduced with permission from Elsevier [12]

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Fig. 14.8  Three schemes of the band gap modifications for visible light sensitization with a lower shift of CBM (a), a higher shift of VBM (b), and impurity states (c). Reprinted with permission from [44]. Copyright (2014) American Chemical Society Fig. 14.9  Comparison of atomic p levels among anions. The band gap of TiO2 is formed between O 2pπ and Ti 3d states. Reprinted with permission from Asahi et al. [44]. Copyright (2014) American Chemical Society

To understand the feasibility of various strategies of doping in the TiO2 lattice, an atomic level energy diagram of the TiO2, showing the O 2pπ and Ti d states in the band gap, is shown in Fig. 14.9. As shown, the CBM of TiO2 consists of the Ti d states, where doping can be easily achieved by substituting the Ti with 3d transition metals, viz., Cr, Fe, Mn, V, etc. Contrarily, the VBM of TiO2 contains the O 2pπ states, where doping can be easily achieved by replacing the anions, containing the atomic p levels, viz., B, C, F, O, P, N, S, etc. It is to be noted that the doping in the CBM can be performed mostly by the metals, while the doping in the VBM can be performed by the non-metals. However, the impurity levels can be created both by metals and non-metals [2, 44]. Overall, the magnetic material as a photocatalyst is found to have their wide application in the photocatalysis process, such as (a) catalyst, (b) co-catalyst, or (c) dopant, wherein their application as a co-catalyst is found to be more promising in the photocatalytic degradation of organic pollutants by virtue of the fact that it has ability to extend the light absorption in the wide range of solar spectrum, i.e., UV to NIR region.

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14.3.2  Magnetic Material as Non-photocatalyst Generally, the non-photocatalytic use of magnetic material in a composite photocatalyst is to facilitate the magnetic separation of the composite photocatalyst from the suspension after use. In such cases, the main photocatalytic material is insulated from the magnetic material by providing a non-conductive reaction barrier. For instance, Kumar et  al. [49] developed a terephthalic acid functionalized g–C3N4/ TiO2 heterojunction photocatalyst which was separated from the magnetic Fe3O4 (magnetite) nanoparticles using a non-conductive SiO2 layer through formation of a core–shell Fe3O4@SiO2 structure. The transmission electron microscope (TEM) image of an unique core–shell Fe3O4@SiO2 nanoparticles can be seen in Fig. 14.10a, wherein the SiO2 layer of ~5 nm thickness can be seen encapsulating the Fe3O4 core of thickness ~22 nm containing the Fe3O4 nanoparticles. Similarly, Álvarez et al. [50] developed a TiO2/SiO2/Fe3O4 photocatalyst for degradation of PPCPs under UV light irradiation. Figure 14.10b shows the TEM image of the TiO2/SiO2/Fe3O4 photocatalyst where the TiO2 shell can be clearly seen over the non-conductive SiO2 layer. Moreover, the Ms values of the visible light-driven terephthalic acid functionalized g–C3N4/TiO2/Fe3O4@SiO2 and UV–light–driven TiO2/SiO2/Fe3O4 composite photocatalysts were found to be 8 and 40  emu/g, respectively. It should also be noted that the Ms value of 8 emu/g is the sufficient Ms value which can help in the magnetic recycling of the composite photocatalysts. In such cases, the SiO2 shell prevents the flow of charge carriers from the main photocatalyst toward the Fe3O4 nanoparticles (also known as the photodissolution effect) which can ultimately result in the reduction of overall photocatalytic degradation efficiency. Furthermore, the SiO2 shell also protects the Fe3O4 core nanoparticles from the oxidation which can help in maintaining its superparamagnetic behavior. Therefore, by utilizing such strategies, the magnetically recyclable photocatalysts can be developed where the magnetic material is in non-photocatalytic role. Fig. 14.10 (a) TEM image of as-synthesized core– shell Fe3O4@SiO2 nanoparticle and (b) TEM image of TiO2/SiO2/Fe3O4 nanoparticles. Reproduced with permission from Elsevier [49, 50]

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14.4  Magnetic Recycling: Practical Challenges Ahead As mentioned in Sect. 14.3, at laboratory scale, the magnetic recycling of the photocatalysts can be performed either by using a handheld laboratory magnet or by employing an electromagnetic separation unit. Various researchers have developed the visible light-driven magnetically recyclable photocatalysts for degradation of organic pollutants, where magnetic separation of photocatalysts from the suspension solution was performed by using a handheld laboratory magnet [4, 36, 38]. However, very few researchers have also performed the degradation of organic pollutant in a visible light-driven prototype reactor rather than in a bench scale reactor. In this regard, after photocatalytic degradation of organic pollutants, i.e., PPCPs, Kumar et al. [7] and Khan et al. [4] were able to separate the magnetic photocatalysts from the suspension solution by employing a ~200 mT electromagnetic magnetic separation unit. In fact, the electromagnetic separation unit in these studies was part of a 5 L visible light-driven prototype reactor system. The schematic diagram (Fig. 14.11a) and the real photographs of the photocatalytic reactor system (Fig. 14.11b) are shown in Fig. 14.11. As shown in Fig. 14.11b, the main photocatalytic reactor unit was made of a 5 L capacity cylindrical acrylic glass unit, encompassed with the multiple visible light sources of integrated irradiance 317 W/m2. For the treatment of large amount of real PPCP-laden wastewater, high capacity of such photocatalytic reactor system is required. In such real scenario, the up-scaling cost of the photocatalytic reactor system would be very high due to the higher cost of the electromagnetic separation unit. Moreover, the installation of an electromagnetic separation unit is another limitation that is due to its large size which could require more space. Similarly, the installation of the real magnet in place of a magnetic separation unit is also not suggested due to the similar reason. Therefore, more research work is required to scale-up the photocatalytic reaction unit with an integrated magnetic separation unit for degradation of organic pollutants by a magnetically recyclable photocatalyst in real practical application. Generally, the photocatalytic degradation efficiency of the magnetically recyclable photocatalysts in the laboratory is tested for organic pollutants present in DI water. However, in real practical applications, other than the organic pollutants, the aqueous environment contains various other chemical substances, such as anions, cations, degradation by-products, humic acid, natural organic matter, etc., which are proven to be detrimental to the photocatalysis process [2]. In fact, the substances, such as humic acid and natural organic matter, either work as scavenger for ROS or reduce the light absorption on the photocatalyst’s surface by shielding the light irradiated in the aqueous solution. Moreover, the anions, cations, or degradation byproducts either deactivate the active sites of the photocatalysts or scavenge the ROS which results in decreased photocatalytic degradation efficiency. Therefore, it is required to test the environmental effectiveness of the magnetically recyclable photocatalysts for degradation of organic pollutants present in the real aqueous environment.

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Wastewater

P

Visible light source 5 L photocatalytic reactor Electromagnetic system Magnetic separation unit

P

Treated water

P

Recovered magnetic photocatalyst

Fig. 14.11 (a) Prototype photocatalytic reactor integrated with a magnetic separation unit and (b) photographs of the developed prototype photocatalytic reactor integrated with a magnetic separation unit. Reproduced with permission from Elsevier [4]

14.5  Conclusions and Future Prospects Photocatalysis, a light-driven catalysis process, has been inviting immense attention among the researchers worldwide due to its effectiveness for degradation of organic pollutants present in water and wastewater. The effectiveness of the process is

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known to be due to its ability to mineralize the organic pollutants into relatively less harmful CO2 and H2O. However, the basic challenges of the photocatalysis process limit its application due to (a) the activity of the photocatalyst in relatively higher energy light only (e.g., UV light of the solar spectrum), (b) the high recombination rate of e−–h+ pair, resulting in dissipation of the captured light energy, and (c) the recovery of the photocatalysts after its use. Recycling process consists of separation, regeneration, and reuse of a photocatalyst. The successful recycling of a photocatalyst can potentially reduce the total cost of the photocatalysis process by repetitive use of the same photocatalytic material. Due to importance of the recycling process, the strategies, viz., (a) immobilizing the photocatalyst on a stable support, such as clay, glass, silica, zeolite, etc. or (b) suspending the photocatalyst in a suspension solution and its recovery after the photocatalytic degradation process, have been adopted. The later strategy has become a more popular strategy than the former one, due to the fact that it can provide more surface area for the photocatalytic reaction. Thus, the development of a magnetically recyclable photocatalyst was found to be a way forward which can effectively facilitate the separation of photocatalysts from the suspension solution after its use. Based on the role of magnetic material in a photocatalyst or composite photocatalyst, such photocatalysts are divided into two broad categories: (a) magnetic material as a photocatalyst and (b) magnetic material as a non-photocatalyst. The magnetic material utilized as a photocatalyst can play roles, such as (a) catalyst, (b) co-catalyst, or (c) dopant. Among these roles, the application of magnetic material as a co-catalyst is found to be the more promising role due to its ability to extend the light absorption in the wide range of solar spectrum (i.e., UV to NIR region). The magnetic material in a non-photocatalytic role is usually insulated from the photocatalytic material by a non-conductive reaction barrier (e.g., SiO2). Moreover, the photocatalytic degradation of organic pollutants at large scale has certain limitations which can be overcome by (a) scaling–up the photocatalytic reaction unit with an integrated magnetic separation unit and (b) ensuring the environmental effectiveness of the photocatalysts in the real aqueous environment. Therefore, the application of magnetically recyclable photocatalysts for real practical purposes requires more research in these areas in future.

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36. Fung, C.  S., Khan, M., Kumar, A., & Lo, I.  M. C. (2019). Visible–light–driven photocatalytic removal of PPCPs using magnetically separable bismuth oxybromo-iodide solid solutions: Mechanisms, pathways, and reusability in real sewage. Separation and Purification Technology, 216, 102–114. 37. Kaur, N., Shahi, S. K., Shahi, J. S., Sandhu, S., Sharma, R., & Singh, V. (2020). Comprehensive review and future perspectives of efficient N–doped, Fe–doped and (N, Fe)–codoped titania as visible light active photocatalysts. Vacuum, 178, 109429. 38. Kumar, A., Khan, M., Zeng, X., & Lo, I. M. (2018). Development of g–C3N4/TiO2/Fe3O4@ SiO2 heterojunction via sol–gel route: A magnetically recyclable direct contact Z–scheme nanophotocatalyst for enhanced photocatalytic removal of ibuprofen from real sewage effluent under visible light. Chemical Engineering Journal, 353, 645–656. 39. Eskandarpour, A., Iwai, K., & Asai, S. (2009). Superconducting magnetic filter: Performance, recovery, and design. IEEE Transactions on Applied Superconductivity, 19(2), 84–95. 40. Gómez-Pastora, J., Dominguez, S., Bringas, E., Rivero, M.  J., Ortiz, I., & Dionysios, D. D. (2017). Review and perspectives on the use of magnetic nanophotocatalysts (MNPCs) in water treatment. Chemical Engineering Journal, 310, 407–427. 41. Ahoranta, M., Lehtonen, J., & Mikkonen, R. (2003). Magnet design for superconducting open gradient magnetic separator. Physica C: Superconductivity, 386, 398–402. 42. Mishra, M., & Chun, D. M. (2015). α–Fe2O3 as a photocatalytic material: A review. Applied Catalysis A: General, 498, 126–141. 43. Zhang, G. Y., Feng, Y., Xu, Y. Y., Gao, D. Z., & Sun, Y. Q. (2012). Controlled synthesis of mesoporous α–Fe2O3 nanorods and visible light photocatalytic property. Materials Research Bulletin, 47(3), 625–630. 44. Asahi, R., Morikawa, T., Irie, H., & Ohwaki, T. (2014). Nitrogen–doped titanium dioxide as visible–light–sensitive photocatalyst: Designs, developments, and prospects. Chemical Reviews, 114(19), 9824–9852. 45. Singh, P., Sharma, K., Hasija, V., Sharma, V., Sharma, S., Raizada, P., Singh, M., Saini, A. K., Hosseini-Bandegharaei, A., & Thakur, V.  K. (2019). Systematic review on applicability of magnetic iron oxides–integrated photocatalysts for degradation of organic pollutants in water. Materials Today Chemistry, 14, 100186. 46. Lamba, R., Umar, A., Mehta, S.  K., & Kansal, S.  K. (2017). Enhanced visible light driven photocatalytic application of Ag2O decorated ZnO nanorods heterostructures. Separation and Purification Technology, 183, 341–349. 47. Formal, F. L., Grätzel, M., & Sivula, K. (2010). Controlling photoactivity in ultrathin hematite films for solar water-splitting. Advanced Functional Materials, 20(7), 1099–1107. 48. Chen, X., Liu, L., Peter, Y. Y., & Mao, S. S. (2011). Increasing solar absorption for photocatalysis with black hydrogenated titanium dioxide nanocrystals. Science, 331(6018), 746–750. 49. Kumar, A., Khan, M., He, J., & Lo, I. M. C. (2020b). Visible–light–driven magnetically recyclable terephthalic acid functionalized g–C3N4/TiO2 heterojunction nanophotocatalyst for enhanced degradation of PPCPs. Applied Catalysis B: Environmental, 270, 118898. 50. Álvarez, P.  M., Jaramillo, J., Lopez-Pinero, F., & Plucinski, P.  K. (2010). Preparation and characterization of magnetic TiO2 nanoparticles and their utilization for the degradation of emerging pollutants in water. Applied Catalysis B: Environmental, 100(1–2), 338–345.

Chapter 15

Titanate Nanostructures as Potential Adsorbents for Defluoridation of Water C. Prathibha, Anjana Biswas, and M. V. Shankar

15.1  Introduction Water is not only an essential component for life but also a basic building block to maintain quality of life. Its purity and availability are inextricably linked to global health and economic development. The presence of several naturally occurring, anthropogenic, and industry-generated ions such as fluoride, arsenic, nitrate, sulfate, iron, manganese, chloride, selenium, heavy metals, and radioactive materials greatly affects the water quality, leading to health problems. The most significant inorganic pollutants in groundwater affecting human health at the global scale, according to the World Health Organization (WHO), are arsenic and fluoride [1]. Fluoride is the only chemical in potable water that can cause varied health effects depending upon its concentration in dissolved form. It is often described as a “double-­edged sword” as inadequate ingestion is associated with dental caries, whereas excessive intake leads to dental, skeletal, and soft tissue fluorosis which has no cure. A very small amount of fluoride (0.4–1.0 mg/L) is beneficial for bone and teeth development and dental health. Especially for young children, it promotes calcification of dental enamel and protects teeth against tooth decay. Therefore, it is considered as an essential mineral with a narrow margin of safety. Due to these clinical manifestations caused by drinking fluoride-contaminated water, the WHO has recommended 1.5 mg/L as the maximum contaminant level (MCL) in drinking water. Fluorosis due to excessive concentration of fluoride >1.5  mg/L has been reported in at least 28 countries from South Asia; Africa; the Middle East; North, C. Prathibha (*) · A. Biswas Department of Physics, Sri Sathya Sai Institute of Higher Learning, Anantapur Campus, Anantapur, Andhra Pradesh, India e-mail: [email protected] M. V. Shankar Nanocatalysis and Solar Fuels Research Laboratory, Department of Materials Science and Nanotechnology, Yogi Vemana University, Kadapa, Andhra Pradesh, India © Springer Nature Switzerland AG 2021 S. Balakumar et al. (eds.), Nanostructured Materials for Environmental Applications, https://doi.org/10.1007/978-3-030-72076-6_15

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Central, and South America; and Europe as shown in Fig. 15.1 [2]. China and India are the most affected countries. Defluoridation of drinking water is the only solution to prevent fluorosis. Various technologies like chemical precipitation, electrocoagulation, reverse osmosis, and electrodialysis have been reported to remove fluoride from drinking water [4, 5]. Despite their unique advantages, these technologies have gained limited social acceptance due to unaddressed problems such as the high costs, poor regeneration, interference of other ions, customary replacement of sacrificial electrodes, consumption of electric power, membrane fouling, requirement of experienced operators, and poor water recovery. Scientific evidence recommends adsorption as the most suitable method as it offers attractive merits such as ease of operation, simplicity of design, and economics. Adsorption is the adhesion of atoms, ions, or molecules from liquid to a solid surface as described in Fig. 15.2 [6]. This process creates a film of the adsorbate on the surface of the adsorbent. The solid material that provides the surface for adsorption is referred to as adsorbent; the species that gets adsorbed is named as adsorbate (fluoride). The success of the adsorption technique entirely depends on the efficiency of the adsorbents used. A wide variety of micron-sized adsorbents have been used for the removal of fluoride from water. These include biosorbents [8], clays [9], soils [10], carbons [11], zeolites [12], alumina-based materials [13], and synthetic resins [14]. They

Fig. 15.1  Predicted probability of fluoride concentration in the groundwater exceeding the WHO guideline for drinking water of 1.5  mg  L−1. Reprinted with permission from [3]. Copyright © American Chemical Society 2008

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Fig. 15.2  Basic terms of adsorption [7]

have attained limited success due to their poor efficiency and kinetics offered. Nanomaterials are the best alternatives to the traditional adsorbents. It has been found that the unique properties of various metal and nonmetal nanomaterials can be used to develop high capacity and selective sorbents for contaminant removal from drinking water. This chapter focuses on a conceptual overview of the recent trends, principles, and applications of advanced nanostructures for the removal of fluoride from aqueous solution. Most of the fluoride water filters available in the market for domestic use are very expensive. A US-based company Crystal Quest manufactures fluoride water filters for residential use. This system removes ions through a combination of adsorption and chemical reaction with the media. The price varies from 99 USD to 1566 USD. Aquagear water filtration, USA, has come up with a 1.3-kg water filter pitcher, which claims to remove fluoride from water. This product is sold in India on different online platforms with price per piece starting from 16,000 INR. It is possible to reduce the cost of these filters with the use of highly efficient cost-effective nanoadsorbents.

15.2  Nanomaterials as Potential Adsorbents Nanotechnology has revolutionized the entire scientific and technological fields. Environmental safety is no exception. One of the most promising and well-­developed environmental applications of nanotechnology has been in water remediation and treatment where different nanomaterials can be used to purify water through adsorption. They have proven themselves to be efficient adsorbents compared to traditionally used materials due to their unique properties at nanoscale. The most vital properties of these particles which are responsible for their high defluoridation potential are small size, large surface area, catalytic potential, large number of active sites, short diffusion route, and high reactivity. These properties fascinated the scientific community and triggered a great activity in the field of defluoridation. In the past 10 years, many researchers have devoted their attention to develop low-­ cost and highly efficient nanoadsorbents for the removal of fluoride from aqueous solution.

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Alumina is the most widely used oxide for defluoridation. Nano-Al2O3 in its gamma phase (γ-Al2O3) with a very high fluoride adsorption capacity was demonstrated as the best among all the alumina-based materials [15]. Iron-based oxides [16] is another well-known adsorbent used for the removal of fluoride from water. Mixed oxides of calcium, magnesium, and aluminum-based materials in their nanoform have also been proven to be the highly efficient adsorbents for defluoridation of water [17, 18]. Adsorption is a surface phenomenon; the surface area of the adsorbent plays a major role. Large surface area is the property of nanomaterials which can be a useful parameter in enhancing the fluoride adsorption capacity in drinking water treatment. In general, nanomaterials possess higher surface area due to their small size. It is known that surface properties are determined not only by size but also by shape [19–21]. For example, sphere and a cube having the same volume, the cube has a larger surface area than the sphere. Developing nanomaterials in varied morphologies has been the recent trend in nanoscience research. Moreover, the discovery of carbon nanotubes stimulated the focus of many researchers toward one-dimensional nanostructures such as nanorods, nanobelts, and nanotubes due to their exceptional electrical, mechanical, and chemical properties. Numerous reports [22–24] have highlighted the use of TiO2-based nanostructures and its composites for sustainable energy and environmental applications. This chapter describes in detail about TiO2-­ based nanostructures and their application in the field of water treatment for contaminant removal.

15.2.1  Titania-Based Nanostructures for Defluoridation In the past decade, design and fabrication of nanostructures based on metal oxides have attracted much attention because of their peculiar electronic and optical properties and their potential applications in the industry and technology [25]. One of the most intensively studied oxides is titania. Titanium dioxide is proven to be less toxic [26], and its multifunctional properties enabled them to use in toothpastes; in food packing; in cosmetic products such as sunscreens, lipsticks, body powders, soaps, and pearl essence pigments; and also in special pharmaceutics [27]. Various applications of TiO2 nanoparticles are summarized in Fig. 15.3. TiO2 exists in three basic phases, namely, anatase, rutile, and brookite [28]. In general, the anatase phase of TiO2 is preferred due to its higher photocatalytic activity, photochemically stable nature, nontoxic nature, and more importantly it is relatively inexpensive. The titanates have layered structure which has close structural resemblance to titanium dioxide, both composed of TiO6 octahedral units connected by sharing corners and edges as shown in Fig. 15.4. The three most general approaches to the synthesis of titanate nanostructures are chemical (template) synthesis, electrochemical approaches (e.g., anodizing of Ti), and the alkaline hydrothermal method. Alkaline hydrothermal method [30] using TiO2 (anatase) as a precursor material is one of the reliable ways of producing

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Fig. 15.3  Applications of TiO2 nanoparticles. Reprinted with permission from [27]. Copyright ©Springer Nature Switzerland AG 2019

Fig. 15.4  Crystal structure of the studied TiO2 polymorphs: (a) anatase, (b) rutile, (c) TiO2–B, and (d) hydrogen titanate (H2Ti3O7). Blue, red, and white spheres represent titanium, oxygen, and hydrogen atoms, respectively. Reproduced under Creative Commons license, [29]

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Fig. 15.5  Common synthesis routes of TiO2 nanostructures. Reproduced under Creative Commons license, [32]

Ti-O-based material in varied morphologies such as titanate nanorods (TNRs), titanate nanobelts (TNBs), and titanate nanotubes (TNTs) [28, 31]. Figure 15.5 portrays all possible synthesis routes of TiO2 nanostructures. In the nanostructure form, titania exhibits favorable electronic and optoelectrochemical properties with additional features such as high porosity and surface area. These features greatly enhance the efficiency of the aforementioned applications and make them suitable for many new applications [33]. The unique structural and functional properties of TiO2-based nanomaterials have led to breakthroughs in the field of photocatalysis, photovoltaics, fast-charging lithium-ion batteries, and smart coatings [28, 34]. Recently, they have also been explored in the field of water treatment for defluoridation of drinking water. Figure  15.6 shows the TEM image of titanate nanostructures.

15.3  Fluoride Adsorption Studies Fluoride adsorption studies are generally carried out by potentiometric method by using the ion-selective electrode. The fluoride adsorption capacity of an adsorbent is determined by the following expression.



 C − Ce qe =  O  m

 V 

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Fig. 15.6  TEM images of titanate nanostructures. Reprinted with permission from [30, 35, 36]. Copyright © Springer Nature 2016, © Elsevier 2016, © Taylor and Francis group 2020

where C0 and Ce are the initial and equilibrium fluoride concentration measured in mg/L. V (L) is the volume of fluoride-containing water treated, and m (g) is the mass of the adsorbent used. The optimization of the parameters, majorly influencing the fluoride adsorption process, is evaluated through batch adsorption studies.

15.3.1  E  ffect of Various Parameters on Fluoride Removal Efficiency of Titanate Nanostructures The adsorption of fluoride ions by an adsorbent depends upon different aspects, for example, surface area, dose, and the isoelectric point (IEP) of an adsorbent. The initial concentration of fluoride ion, contact time, and pH of the solution equally influence the adsorption efficiency of an adsorbent. The optimization of these parameters plays an important role to identify the best operating parameters that would give the maximum adsorption capacity of the adsorbent. This becomes specifically essential while estimating the applicability of the adsorbent and designing a product for real-time application such as a filter for remediating groundwater and

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wastewater. Further, knowing the best operating parameters, one could fine-tune the operating conditions in order to enhance the adsorption capacity. The optimization is usually done using the batch adsorption experiments. An ideal adsorbent would have a wide pH window for adsorption and high fluoride uptake in a short contact time with low adsorbent dosage. Surface Area Higher surface area offers a major portion of its particles to be present at the surface which in turn increase the number of active sites for fluoride adsorption [37]. As shown in Fig. 15.7, the particles forming the bulk of any material experiences equal forces from all sides and, hence they do not have any unbalanced force. On the other hand, the particles on the surface of the materials experience unbalanced forces, which form the driving force for attracting the adsorbate ions/molecules. As the surface area of the material increases, the driving force for adsorption also increases. Surface area of an adsorbent greatly influences its adsorption efficiency, since adsorption is a surface phenomenon. Among the three variants of titanate nanostructures, TNTs possess the highest surface area which could be attributed to their hollow tubular nature. Comparing the adsorption capacities of the three one-dimensional nanostructures, titanate nanotubes (TNTs), titanate nanobelts (TNBs), and titanate nanorods (TNRs), it can be observed that fluoride adsorption capacity follows the trend TNT > TNB > TNR. The higher adsorption capacity of TNT compared to the TNB and TNR is mainly due to higher surface area as a result of its tubular morphology. The surface area of TNRs, TNBs, and TNTs is reported to be 29, 38, and 282  m2/g, respectively, which is nearly 5, 6, and 60 times higher than the precursor material, TiO2 fine particles (5 m2/g) [30, 35]. TNT with higher surface area offers a greater number of adsorption sites on its surface and hence the higher adsorption capacity. Further, the TNR with the lowest surface area has the least number of adsorption sites available on the

Fig. 15.7  Driving force for adsorption: unbalanced forces of surface atoms

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surface and therefore the least adsorption capacity. However, all the three one-­ dimensional morphologies of titanates have superior fluoride adsorption compared to TiO2 with spherical morphology. It could be seen that there exists a perfect correlation between fluoride adsorption efficiency and the specific surface area of the adsorbent. Titanate nanostructures demonstrate the effect of morphology on fluoride adsorption capacity. Hence, for effective adsorption to take place, adsorbents with higher surface areas are required. Dose of an Adsorbent Used Dosage is the quantity of adsorbent in g/L that is added to the fluoride-containing water. An ideal adsorbent should exhibit maximum adsorption capacity with minimum dosage. In general, adsorption capacity can be increased with increase in adsorbent dose. This is due to the increased number of available adsorption sites which results in enhanced adsorption capacity. This trend in variation of adsorption capacity with adsorbent dosage was reported by various researchers [30, 35, 38]. However, to be a cost-effective adsorbent, the optimum dose for defluoridation ought to be as low as possible. This requirement was satisfied by the titanate nanostructures. TNT is reported to have reduced higher fluoride concentration to WHO limits with the use of as low as 1 g/L dosage. This was possible due to the high surface area of TNTs by virtue of its tubular morphology [35]. In general, nanomaterials with very low adsorbent dosage have proven to exhibit much higher fluoride adsorption capacity when compared to bulk materials. Figure 15.8 shows the comparison. Rajan et al. [39] reported the usage of 15 g/L of zirconium-impregnated walnut shell carbon to treat water with 3 mg/L of fluoride which is a higher dosage when compared to dosage of titanates used for defluoridation.

Fig. 15.8  Variation of adsorption capacity with adsorbent dosage using (a) natural clay (pH 5.8, initial fluoride concentration 5 mg/L, contact time 3 h) and (b) titanate nanotubes (pH 2, initial fluoride concentration 10  mg/L, contact time 2  h). Reprinted with permission from [9, 35]. Copyright © Académie des sciences 2018, © Elsevier 2016

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Contact Time Contact time is another vital factor in fluoride adsorption process by an adsorbent. Initially, the adsorption is rapid due to availability of vacant active sites on the surface of an adsorbent; as the equilibrium approaches, the number of free sites on the adsorbent surface decreases, and the adsorption capacity attains saturation. Equilibrium time is the minimum time for which adsorbent needs to be in contact with fluoride ions in order to obtain maximum fluoride removal. Numerous reports [35, 38, 40] indicate the quick fluoride adsorption process, and lesser equilibrium time  taken by one-dimensional nanostructures compared to conventionally used micron-sized adsorbents. This is illustrated in Fig. 15.9. This could be attributed to the fact that nanomaterials have much greater number of adsorbent sites readily available for the uptake of adsorbate ions compared to their bulk counterparts. This is due to their high surface area which in turn ensures quick adsorption. Hence, researchers are working toward developing adsorbents which exhibit instantaneous adsorption using nanomaterials, in which equilibrium adsorption occurs within a few minutes. Isoelectric Point and pH of the Solution Zeta potential is a measure of effective surface charge of nanoadsorbent which varies with the pH of adsorbent solution. The isoelectric point (IEP) is the pH at which the zeta potential of the adsorbent is zero. At IEP, both positive and negative charges are well compensated that leads to no excess surface charges on the surface of titanate nanostructures. At this pH the number of positive surface sites equals the number of negative surface sites. In addition, at all pH  IEP the surface charge of the adsorbent is negative. This is

Fig. 15.9  Comparison of equilibrium contact time using (a) nanomaterial (CeO2–ZrO2 nanocages) (adsorbent dosage 0.2 g/L, pH = 4) and (b) bulk material (zirconium I-impregnated coconut shell fiber) (adsorbent dosage 10 g/L, pH = 4). Reprinted with permission from [41, 42]. Copyright © Elsevier 2013, © Taylor and Francis Group LLC 2008

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Table 15.1  Fluoride adsorption efficiencies of Ti-O-based adsorbents and pH conditions at which they exhibit maximum efficiency

Adsorbent TiO2 TiO2-SiO2 nanocomposite Titanate nanotubes Anatase TiO2 TiO2 Microbially synthesized Aluminum titanate Magnesium titanate FZTNT: Fe(III) and Zr(IV) surface-functionalized TNT

pH at which adsorbent showed maximum adsorption capacity 1 1 2–3 3–4 2–4

Maximum adsorption capacity (mg/g) 100.7 152.2 58 32.15 0.85

References [47] [47] [35] [48] [49]

3–9 3–11 7

0.85 0.029 229

[50] [51] [45]

an important parameter which determines the pH sensitivity of an adsorbent in adsorption process. Fluoride adsorption is favorable when pH of the solution is less than IEP of an adsorbent as it favors electrostatic attraction between the positively charged surface and negative fluoride ion, whereas at pH greater than IEP the reverse process occurs, as the negative adsorbent surface now repels the negatively charged fluoride ions. Most of the titanates reported in literature exhibit favorable fluoride adsorption in acidic conditions due to their lower IEP values (Table 15.1).  luoride Adsorption Mechanism by One-Dimensional F Titanate Nanostructures The protonic trititanate nanostructures are made up of Ti and O, stacked up into multiple layers in its crystal structure with ion-exchangeable H+ and OH−. The mechanism of fluoride adsorption by titanate nanostructures is well understood based on the abundant availability of Ti-OH ions on the titanate nanostructures’ surface. These hydroxyl ions act as the active sites for effective adsorption of fluoride ions from water. The similar ionic radii of F− and OH− ions indicate the possibility of fluoride ions from the aqueous media undergoing ion exchange with the surface OH− ions [30]. “Ion exchange is a stoichiometric process where any counter ion leaving the ion exchanger surface is replaced by an equivalent number of moles of another counter ion to maintain electro-neutrality of the ion exchanger” [43]. Figure 15.10 depicts the ion-exchange mechanism by titanate nanostructures. The mechanism is usually supported by experimental evidences such as Fourier transform infrared spectroscopy (FTIR) and x-ray photoelectron spectroscopy (XPS) of the adsorbent before and after fluoride adsorption. Using both the characterization techniques, the presence of M-OH (M = metal) peak could be observed which confirms the presence of -OH on adsorbent surface. The decrease in intensity of the M-OH peak in FTIR spectra of post-fluoride adsorption could be ascribed to

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Fig. 15.10  Schematic representation of ion-exchange mechanism 1-D titanate nanostructures and spherical nanoparticle. Reprinted with permission from [35]. Copyright ©Elsevier 2016

the successful ion exchange between the hydroxyl group and the fluoride ions [35, 44–46]. In addition, the shift in the high-resolution peaks of Ti in the XPS spectra of the titanates toward higher binding energy in post-fluoride adsorption proves the successful uptake of fluoride ions by the adsorbent [44, 45]. The initial pH of fluoride-containing solution also greatly affects the surface charge of adsorbent, which is demonstrated using isoelectric point (IEP). The concentrations of protons or surface hydroxyl groups which are responsible for fluoride uptake process by an adsorbent depend on its IEP and pH of the fluoride-containing solution. Therefore, adsorption mechanism is well understood based on the following two possible reactions at the oxide water interface:

Ti − OH + H 2 O → Ti − OH 2+ + OH −

pH < IEP

(15.1)



Ti − OH + H 2 O → Ti − O − + H 3 O +

pH > IEP

(15.2)

Equilibrium (15.1) is favorable when pH of the solution is lesser than IEP, whereas equilibrium (15.2) is more favorable when pH of the solution is greater than IEP. However, in any condition whether acidic or basic, Ti − OH 2+, Ti − OH, and Ti − O− are present but in different concentrations which determine the adsorption capacity. In acidic media (pH  420 nm Visible light (λ > 420 nm Xe lamp (λ ≥ 420 nm) Xe lamp (λ ≥ 420 nm)

H2 production μmol h−1 g−1 Refs. 1232 [138] 104.7

[139]

867.6

[140]

1009

[141]

288

[142]

19.11  Conclusions In summary, we have reviewed the recent developments in different metal nitrides and graphitic carbon nitrides to achieve efficient H2 generation. Our discussions include different synthesize approaches of different metal nitrides, crystal structures, as well as efficient hydrogen evaluation. In general, absorption of light, separation of charge carrier’s pair, passage of hole, surface kinetics, suitable structure, and optimized composition of materials are the important features in research and design. Different metal nitrides and g-C3N4-based hybrid composite systems for catalytic H2 generation have been studied. Metal nitrides are the most appropriate photocatalysts that can be used in various photocatalysis methods. A combination of nano-sized metal nitrides with appropriate small bandgap materials produces hybrid catalysts which show development in H2 generation from the visible light region. Still, advanced technological developments like doping and ease growth mechanisms are being subtle. It is also intended in this book chapter that understanding of other material and compositional systems is necessary to join the whole solar spectrum. However, still research works are going on to integrate metal nitride nanowires with other materials which still need improvement toward attaining efficient H2 generation to meet the future needs. In conclusion, the following upcoming research aspects of metal nitrides for their widespread usage in practical applications include defect-free crystallite structures, desired thickness to prevent misfit disorders, controlled doping, enhanced surface area, less cost, and consistency. Moreover, the progress of new developments with a strong theoretical background is necessary for a better sympathetic mechanism for hydrogen generation and eco-friendly water-­ splitting procedure for H2 generation. Acknowledgments  This work was supported by the National Research Foundation of Korea grant funded by the Korea government (No. NRF-2017R1A4A1015581).

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Chapter 20

Highly Functionalized Nanostructured Titanium Oxide-Based Photocatalysts for Direct Photocatalytic Decomposition of NOx/VOCs Katchala Nanaji, Manavalan Vijayakumar, Ammaiyappan Bharathi Sankar, and Mani Karthik

20.1  Introduction In a new civilized world, air pollution is one of the major concerns, and it has a serious toxic impact on human health as well as the environment. The ground-level ozone (GLO), particulate matter (PM), sulphur oxides (SOx), carbon monoxide (CO), nitrogen oxides (NOx) and volatile organic compounds (VOCs) are considered as six major air pollutants according to the World Health Organization (WHO). Air pollutants can create several serious diseases like asthma, eye irritation, headaches, skin diseases, respiratory and cardiovascular diseases, foetal growth, ventricular hypertrophy, lung cancer, psychological complications and low birth weight [1–4]. Combination of hydrocarbons and NOx in the atmosphere produces a photochemical smog via photochemical oxidation under the sunlight, and the photochemical smog is environmentally hazardous. Generally, air pollutants can be created by natural sources or manmade sources, and it is classified as primary air pollutants or secondary air pollutants. Primary air pollutants are directly emitted from natural volcanic eruption, vehicle exhaust and industrial processes. Secondary air pollutants are not directly emitted from the direct resources, but primary pollutants could K. Nanaji · M. Karthik (*) Centre for Nanomaterials, International Advanced Research Centre for Powder Metallurgy and New Materials (ARCI), Balapur, Hyderabad, Telangana, India e-mail: [email protected] M. Vijayakumar Centre for Nanomaterials, International Advanced Research Centre for Powder Metallurgy and New Materials (ARCI), Balapur, Hyderabad, India Global Innovative Centre for Advanced Nanomaterials (GICAN), Collage of Science, Engineering and Environment, The University of Newcastle, NSW, Callaghan, Australia A. B. Sankar School of Electronics Engineering, Vellore Institute of Technology (VIT), Chennai Campus, Chennai, Tamil Nadu, India © Springer Nature Switzerland AG 2021 S. Balakumar et al. (eds.), Nanostructured Materials for Environmental Applications, https://doi.org/10.1007/978-3-030-72076-6_20

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react or interact to form secondary pollutants. For example, ground-level ozone is one of the secondary air pollutants which formed from the photochemical smog.

20.1.1  Sources of Air Pollutants and Its Impact PM pollutants can be created by manmade sources and/or natural sources. For example, forest and grassland fires, natural volcanoes, living vegetation, sea spray, etc. are the major resources of PM pollutants. On the other hand, PM pollutants can also be created by manmade activities such as burning of fossil fuels in the vehicles, power plants and various industrial activities. The creation of fine particles in the air can cause several health hazards to the living things. On the other hand, the emissions of NOx and VOCs are also creating severe impacts on human health, environment and biological ecosystem [4–7]. The emissions of these pollutants cause severe environmental threats including acid rain, PM  releases, smog and greenhouse gasses. VOCs can react with NOx and other airborne chemicals in the air to produce the ozone through photochemical oxidation under the sunlight, and this ozone is a primary component of smog as environmentally hazardous. Figure 20.1 illustrates the various potential resources of major air pollutants. Due to air pollution, around 8.3 million people are killed in the world, and about 87% of death rate occurs in the low- and middle-income countries as per the WHO reports (WHO, 2016) [1, 2]. For example, India and China are among the most polluted countries owing to the rapid urbanization [1, 2]. The estimated number of premature pollution-related deaths per year in the world is represented in Fig. 20.2. Photocatalytic oxidation (PCO) using titanium dioxide (TiO2) semiconductors has been recognized as one of the efficient and economic techniques for the complete decomposition of harmful pollutants into nontoxic final products [8, 9]. The novel photocatalytic processes are being developed that are totally a safer and cleaner process and efficient and environmental-friendly nature. Generally, the photocatalytic processes can be utilized not only in ultraviolet but also in solar or visible light region. The functionalized photocatalytic systems will be essential for the future prosperity of mankind. In this context, several studies have been conducted by using TiO2 photocatalyst for the photocatalytic oxidation of various pollutants [10–18]. However, most of the reports and reviews have mainly described on single pollutant treatment system using either TiO2 or TiO2-supported catalysts. For example, Anpo and co-workers have extensively studied and also reviewed [13–15] the photocatalytic degradation of NOx by using TiO2 as well as modified TiO2 as photocatalysts. However, there are very limited studies on the photocatalytic removals of multiple air pollutants over TiO2-based photocatalysts. Generally, indoor environment in most cases will often contain more than one pollutant. Hence, the photocatalytic material capable of destroying multi-air pollutants in the single system is showed more attraction for practical photocatalytic implementation. On the other hand, the dual functional adsorbent/photocatalyst system has been suggested as one of the effective and

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Fig. 20.1  Various potential resources of air pollutants such as particulate matter (PM), nitrogen oxides (NOx) and volatile organic compounds (VOCs)

economic techniques for the removal of pollutants [19–21]. From the obtained results, it was found that the removal efficiency of pollutant over combination of adsorption/desorption with photocatalytic oxidation system was three times higher as compared with the normal photocatalytic oxidation process. Thus, these adsorbent/photocatalyst composite materials deserve research attention due to their combination properties of adsorption and photocatalysts which enhance the activity and selectivity towards specific air pollutants. In the present chapter, the photocatalytic destructions of major air pollutants like VOCs and NOx over TiO2 and functionalized nanostructured TiO2 photocatalysts are discussed. The fundamental photocatalytic concepts and strategies of photocatalysts on environmental remediation are summarized. This chapter also provides some highlights and examples regarding recent progressive research on dual functional composites of TiO2/porous catalytic adsorbents for photocatalytic destruction of air pollutants. Finally, the current and future research directions on photocatalytic decomposition of NOx and VOCs for environmental remediation are also highlighted.

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Fig. 20.2  The estimated number of premature pollution-related deaths per year in the world [1, 2]

20.2  P  hotocatalytic Destruction of Air Pollutants by Using TiO2 Photocatalysts Photocatalytic destruction of air pollutants using semiconductor photocatalyst is one of the most promising techniques owing to its nontoxicity, long-term durability, high photosensitivity, environmental-friendly nature, easy availability, economics and high reactivity for the complete decomposition of toxic compounds into harmless products. In general, semiconductors like TiO2, ZnO, WO3, CdS, Fe2O3, ZnS and SnO2 are commonly utilized as photocatalysts. Among various semiconductor photocatalysts investigated, TiO2 is one of the most effective and widely used commercial photocatalysts. And TiO2 photocatalyst has an outstanding advantage and strong capacity in the mineralization of environmental pollutants even with a low concentration. Thus the photocatalytic oxidation of air contaminants using TiO2 photocatalysts has been extensively studied by several research groups over the past several decades [4, 16–18]. Although TiO2 photocatalysts have been extensively investigated for environmental remediation, the large band gap of anatase TiO2 (3.2 eV) limits its visible light absorption (λ  380 nm). The basic principle of photocatalytic destruction of VOCs is depicted in Fig. 20.10. The reaction steps involved in the photocatalytic destruction of VOC pollutants over semiconductor TiO2 photocatalyst are as follows [43]:

Photoexcitation : TiO2 + hϑ → h + + e − (20.12) OH − + h + → OH• (20.13)



Oxidation reaction :



Reduction reaction : O2(ads) + e − → O2−(ads)

(20.14)

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Ionization of water : H 2 O → OH − + H + (20.15) Protonation of superoxide :

O2 − + H + → HOO• (20.16) •



Electron scavenger : HOO + e − → HOO − (20.17)



Formation of H 2 O2 : HOO − + H + → H 2 O2 (20.18)



OH· + pollutant + O2− → products ( CO2 , H 2 O, etc. ) (20.19)

20.3.4  M  esoporous Materials as Supports for Photocatalytic Destruction Mesoporous materials have attracted research interest owing to their high surface area, high pore volume and high pore diameters, and hence it has been used as a supporting material for several applications [44–47]. In the recent research concern, Ti-containing mesoporous molecular sieves such as MCM-41, MCM-48, SBA-1, SBA-15 and HMS have been utilized as active and selective photocatalysts for the decomposition of NOx. Because such mesoporous materials are possessing high surface area with high pore diameter, high thermal and hydrothermal stability which are well-suitable features as catalyst/catalyst-supported, adsorbent and host-guest encapsulation materials [48–50]. This new family of mesostructure with hierarchal pore network catalyst is expected to demonstrate enhanced reactivity, selectivity and resistance to deactivation. The isomorphous substitution of transition metal ions such as Ti, Mo, Cu, V and Cr, into the framework of mesoporous materials has remarkable catalytic and photocatalytic applications. The incorporation of transition metal atoms into the framework can create both acidic and redox sites which make such materials potentially active photocatalytic materials for the photocatalytic decomposition of NOx. The schematic representation of framework of MCM-41 and isomorphous substitution of metals is illustrated in Fig. 20.11.  i-Containing Mesoporous Materials for Photocatalytic T Decomposition of NOx Ti-MCM-41 and Ti-HMS have been synthesized and used as an efficient photocatalysts. The photocatalytic destruction of NO into N2 and O2 was examined over Ti-MCM-41, HMS and TS-1 catalysts, and the photocatalytic performance with respect of selectivity of N2 and N2O is illustrated in Fig. 20.12. Generally, the photocatalytic destruction of NOx depends on the type of porous materials which can be used as a support. From the figure, it is clearly evident that Ti-MCM-41 material has showed higher selectivity of N2 than that of Ti-HMS and TS-1 materials [51]. This higher activity and selectivity of Ti-MCM-41 is due to the

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Fig. 20.11  Schematic illustration of framework of MCM-41 molecular sieves: (a) neutral, (b) M (divalent metal ions) and (c) M (trivalent metal ions)

formation of well-dispersed Ti-oxide species within the framework and/or cavities of Ti-MCM-41. Hence, the obtained results clearly indicated that isolated and tetrahedrally coordinated Ti-oxide species with well dispersion can act as active sites for the photocatalytic destruction of NO into N2 and O2. The yield and selectivity of N2 and N2O in the direct photocatalytic destruction of NO over various amounts of Ti-loaded Ti-MCM-41, Ti-HMS and Bulk TiO2 catalysts are listed in Table 20.1. Furthermore, Hu et al. [51] have also investigated the mesoporous materials with various Ti metal contents and characterized by using various in situ spectroscopic techniques such as photoluminescence, FTIR, UV-Vis and XAFS spectroscopy. They have demonstrated on the influence of the Ti contents in the Ti-MCM-41 on the selectivity of N2 for the photocatalytic destruction of NO by using FTIR and photoluminescence as shown in Fig. 20.13. From Fig.  20.13, it can be observed that the total amount of the tetrahedral Ti-oxide species exposed at the surface of Ti-MCM-41 can be monitored by the intensity of the δasym NH3 IR band. On the other hand, the total amount of isolated tetrahedrally coordinated Ti-oxide species can be seen from the photoluminescence spectra of Ti-MCM-41. From the above results, it can be concluded that highly

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Fig. 20.12  Conversion of N2 and N2O for the photocatalytic destruction of NO over Ti-MCM-41, Ti-HMS and TS-1 catalysts under UV irradiation at 295 K [51] Table 20.1  The yield and selectivity of N2 and N2O in the direct photocatalytic destruction of NO over various photocatalysts [51] Catalysts Ti-MCM-41 Ti-MCM-41 Ti-MCM-41 Ti-MCM-41 TS-1 Ti-HMS Ti-HMS Ti-HMS Ti-HMS Bulk

Ti content in Wt.% 0.15 0.60 0.85 2.00 0.60 0.60 1.26 2.30 11.6 –

Yield (μmol/g of cat. h) N2 N2O Total 0.60 0.20 0.80 1.50 0.40 1.90 0.40 0.45 0.85 0.25 0.75 1.00 0.85 0.70 1.55 0.95 0.45 1.40 16.0 3.0 19.0 13.0 4.5 17.5 7.5 6.0 13.5 3.5 9.0 12.5

Selectivity (%) N2 N2O 75 25 79 21 47 53 25 75 55 45 68 32 84 16 74 26 56 44 28 72

dispersed and isolated tetrahedrally coordinated Ti-oxide species in the catalyst can act as active sites for the photocatalytic destruction of NO into N2 and O2. Anpo and co-workers [52] have prepared a bimetal substituted mesoporous V-Ti-MCM-41 catalyst by using a simple photo-assisted synthesis method, and they have noticed an absorption spectrum with remarkable red shift compared with Ti-MCM-41. The above-obtained results clearly suggested that bimetal substituted mesoporous V-Ti-­ MCM-41 catalyst can absorb light of longer wavelengths in the near-visible light region. Furthermore, substitution of two different transition metal ions in the framework of mesoporous photocatalysts can utilize the most efficient solar energy resource.

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Fig. 20.13  Photocatalytic destruction of NO into N2 and N2O: (a) infrared band of δasym NH3 adsorbed on Ti(IV5) and (b) photoluminescence spectra of Ti-MCM-41 with different Ti contents [51] CrS-1 λ>450nm

Reaction Time/ h

7 6

N2 Selec, CrS-1 λ>270nm 97% Cr-HMS λ>450nm

d c

5

Cr-HMS λ>270nm

b

4

N2 Selec, 45%

a

3

Light on

2 1 0

0.0

0.2

0.4 0.6 N2 Yields/mol. g-cat-1

0.8

1.0

Fig. 20.14  The photocatalytic destruction of NO by using Cr-HMS and CrS-1 under UV (λ > 270 nm) as well as visible (λ > 450 nm) light irradiations [53]

Anpo et al. [53] have also reported that Cr-HMS also exhibited higher photocatalytic destruction of NO under both UV and visible light irradiation at 275 K. The reaction time profiles of the photocatalytic destruction of NO on Cr-HMS and CrS-1 under both UV and visible light irradiation are shown in Fig. 20.14. It is observed

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from Fig. 20.14 that the yields of N2 increased with respect to irradiation time and the selectivity for N2 (97%) under visible light irradiation showed higher value as compared with UV light irradiation (45%). From the obtained results, it is concluded that Cr-HMS can act as an active photocatalyst under both UV and visible light irradiation. The preparation and photocatalytic destruction of NO over Ti-HMS mesoporous materials were investigated. It was found that Ti-HMS catalyst with highly dispersed isolated tetrahedral Ti species was found to be active species for high selectivity of N2. Nevertheless, the bulk TiO2 and Ti-HMS (10) photocatalysts were found to be favouring the formation of N2O due to the aggregated titanium oxide species with octahedral coordination. The effects of Ti content and product distribution of the photocatalytic destruction of NO on various titanium oxide-based mesoporous catalysts are represented in Table 20.1.  hotocatalytic Oxidation of VOCs over Ti-Containing P Mesoporous-Based Photocatalysts VOCs are very harmful and hazardous compounds, and hence several abatement technologies of VOCs have been developed. However, there is no single method available which focused on high energy consumption, low removal efficiency and environmental friendliness. In this regard, the dual functionality of adsorption and photocatalytic destruction of VOCs is one of the most promising techniques because of the high removal efficiency. The design and developments of innovative TiO2-­ supported nanoporous composite photocatalysts are strongly required to enhance the photocatalytic destruction of VOCs. The photocatalytic activity is mainly associated with the amount of organic compounds adsorbed on the surface of the photocatalysts, and then the adsorbed organic compound is easily decomposed by the photocatalysts under the solar irradiation.

20.4  M  echanism and Activity of Titanium-Based Photocatalysts for Photocatalytic Destruction of NOx and VOCs The photocatalytic active sites and their mechanisms of photocatalytic decomposition reactions are briefly described in the present section. Yamashita and Anpo [54] have reported the relationship between the coordination number of Ti-oxide species and the selectivity of N2 in the photocatalytic destruction of NO by using Ti-oxide photocatalysts as shown in Fig. 20.15. It can be observed from Fig. 20.15 that the selectivity of N2 is clearly depending on the coordination number of the Ti-oxides located within the framework of zeolites. From the obtained results, it is noted that a highly selective photocatalytic

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Fig. 20.15  Correlation between the coordination number of Ti-oxides and the selectivity of N2 in the photocatalytic destruction of NO using Ti-oxide photocatalysts [54]

destruction of NO into N2 and O2 can be achieved by using highly dispersed and isolated tetrahedrally coordinated Ti-oxide species in the zeolite framework. Nevertheless, the aggregated Ti-oxides with octahedral coordination showed the predominant formation of N2O as product. The photocatalytic selectivity of N2 is strongly depending on the nature of active sites and type of catalysts used in the decomposition reaction. The reaction mechanism of the photocatalytic destruction of NO into N2 and O2 on tetrahedrally coordinated Ti-oxide-supported zeolite catalyst under UV light irradiation was proposed by Matsuoka and Anpo [55] as shown in Fig. 20.16. It was found that Ti-oxide species prepared within the framework of zeolites have revealed a unique local structure as well as a high selectivity in the oxidation of organic substances with H2O2. The previous studies clearly demonstrated that the local structure and the dynamics of photochemical reactivities of the well-defined transition metal oxides (titanium, chromium oxides, etc.) incorporated into the cavities and frameworks of various zeolites and zeo-type materials play a vital role in the photocatalytic decomposition reaction. Anpo et al. [56] have investigated the photocatalytic reaction of NO in the presence of various kinds of hydrocarbons such as methane, ethane and propylene. They found that the photocatalytic efficiency of the NO strongly depends on the kind of hydrocarbons used. They also proposed the important role of the intermediate species formed from NO and hydrocarbon radicals such as the propyl radicals, which were subsequently followed by further reaction with NO to produce N2, and then the overall reaction of NO in the presence of hydrocarbons enhances the photocatalytic efficiency.

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Fig. 20.16  Mechanism of the photocatalytic destruction of NO into N2 and O2 on tetrahedrally coordinated Ti-oxide-supported zeolite catalyst under UV and visible light irradiation [54, 55]

20.5  Summary and Prospective Photocatalytic destruction of air pollutants using semiconductor nanostructured TiO2 photocatalyst has been extensively investigated as a potential and effective technology for air purification because it can directly decompose the pollutants into harmless counterparts under ambient atmosphere. Particularly, highly functionalized nanostructured titanium oxide-based photocatalysts can be a perfect candidate for removing low concentration pollutants (even sub-ppm levels) in outdoor as well as indoor environments. Among several proposed technologies, the photocatalytic destruction of NOx/VOCs is recognized as a promising technology owing to the excellent selectivity to N2 with considerable NOx conversion, harvesting based on solar light, energy saving, low-cost technology, environmental energy and conversion of NOx into nitrates (NO3−) which may recover as a conceivable raw material for fertilizers. It has been identified that the dual functional adsorbent/photocatalyst system could be one of the effective and economic techniques for the removal of air pollutants in the gaseous phase because the removal efficiency of pollutant over the combination of adsorption/desorption with photocatalytic oxidation system was three times higher than the normal photocatalytic oxidation system. Thus, these

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adsorbent/photocatalyst composite materials deserve research attention due to their combination properties of adsorption and photocatalysts which enhance the activity and selectivity towards specific air pollutants. A combination of these fascinating and unique features of the photocatalytic materials could provide a better approach in the utilization of solar energy as the most abundant and safe energy source, and these photocatalytic composite materials are promising candidates for the effective reduction of toxic contaminants into harmless products at ambient temperature. Acknowledgement The authors greatly acknowledge the funding support from Technical Research Centre (TRC) project (Ref. No. AI/1/65/ARCI/2014 (c)) sponsored by the Department of Science and Technology (DST), Government of India, New Delhi, India.

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51. Hu, Y., Martra, G., Zhang, J., Higashimoto, S., Coluccia, S., & Anpo, M. (2006). Characterization of the local structures of Ti-MCM-41 and their photocatalytic reactivity for the decomposition of NO into N2 and O2. The Journal of Physical Chemistry B, 110, 1680–1685. 52. Anpo, M., & Thomas, J. M. (2006). Single-site photocatalytic solids for the decomposition of undesirable molecules. Chemical Communications, 3273. 53. Hu, Y., Wada, N., Matsuoka, M., & Anpo, M. (2004). Photo-assisted synthesis of V-MCM-41 under UV light irradiation. Catalysis Letters, 97, 49–52. 54. Yamashita, H., Yoshizawa, K., Ariyuki, M., Higashimoto, S., Anpo, M., & Che, M. (2001). Photocatalytic reactions on chromium containing mesoporous silica molecular sieves (Cr-HMS) under visible light irradiation: Decomposition of NO and partial oxidation of propane. Chemical Communications, 435–436. 55. Yamashita, H., & Anpo, M. (2003). Local structures and photocatalytic reactivities of the titanium oxide and chromium oxide species incorporated within micro- and mesoporous zeolite materials: XAFS and photoluminescence studies. Current Opinion in Solid State & Materials Science, 7, 471–481. 56. Matsuoka, M., & Anpo, M. (2003). Local structures, excited states, and photocatalytic reactivities of highly dispersed catalysts constructed within zeolites. Journal of Photochemistry and Photobiology C: Photochemistry Reviews, 3, 225–252.

Chapter 21

Bandgap Engineering as a Potential Tool for Quantum Efficiency Enhancement Reddy Kunda Siri Kiran Janardhana, Raju Kumar, Tata Narsinga Rao, and Srinivasan Anandan

21.1  Introduction It is a well-known fact that pollution is increasing and humanity is under constant exposure to harmful pollutants. Some studies even suggest the presence of an almost equal level of pollutants in both indoor and outdoor environments. Since, an average person spends almost 80% of his time in indoor environments [1], there is an increase in the use of various daily comforting agents like air fresheners, deodorants, paints, smoke-inducing agents, pesticides for pest control  which  combined with household cooking leads to the formation of harmful by-products. These reasons coupled with poor ventilation only exacerbates the situation. Such conditions manifest themselves through deteriorated physical and mental health introducing a number of symptoms which are collectively defined as sick building syndrome. This leads to a reduction in human productivity, acute discomfort and in worst cases chronic diseases [2]. To tackle this situation, there are some commonly used remedies like (1) dilution, where the indoor air is mixed with disinfected clean air; (2) use of filters like high-efficiency particulate air (HEPA) filters at the openings of heating, ventilation and air-conditioning (HVAC) systems; (3) UV irradiation, where high-energy UV light is used to destroy DNA/RNA making pathogens inactive/ineffective; (4) installation of dehumidifiers with silica gels inside them, which enables them to trap volatile organic compounds (VOC); and (5) anti-bacterial sprays (furthering the use of comforting goods) which are usually alcohol based. Some of the advantages of such methods include their almost zero contact with the running equipment, well-­ established protocols and most importantly a sense of quick remediation or the sentiment of instantly feeling good [3]. However, some of these solutions involve a lot R. K. S. K. Janardhana · R. Kumar · T. N. Rao · S. Anandan (*) Centre for Nano Materials, International Advanced Research Centre for Powder Metallurgy and New Materials, Hyderabad, Telangana, India e-mail: [email protected] © Springer Nature Switzerland AG 2021 S. Balakumar et al. (eds.), Nanostructured Materials for Environmental Applications, https://doi.org/10.1007/978-3-030-72076-6_21

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of initial manpower, very high installation costs, maintenance and a lot of power input making them economically unviable. As a disruptively better solution, semiconductor photocatalysis is proving to address some of above-mentioned issues quite efficiently. The major advantages of this technique are zero human intervention, requiring only sunlight; rapid and economical end-to-end process; and almost zero maintenance.

21.2  How Does Photocatalysis Work? Photocatalysis is a technique in which (Fig. 21.1) sunlight acts as an accelerator to a chemical reaction between encompassed reactants in the presence of semiconductors [4]. All semiconductors (SCs) are characterized by a forbidden region called the bandgap in between the conduction band (CB) and valance band (VB). When the energy of incident light is higher than the bandgap of the SC, the e−/h+ pairs present in the SC separate, and the electrons get excited to the CB of the SC, whereas the holes remain in the VB. These e−/h+ pairs participate in the reduction and oxidation of pollutants, respectively. These e−/h+ pairs have an inherent nature of recombining back, and thus, effective utilization of these e−/h+ is a must for quantum efficiency (QE) enhancement. Sunlight, which is abundantly available has only less than 2% UV light.  Thus, for effective utilization of the available spectrum, visible light-­ active photocatalysts are required [5]. TiO2 and ZnO are borderline UV-active materials, and hence it is easier to convert them to visible light-active materials that can utilize the maximum spectrum of the sunlight for effective pollutant remediation. Since, indoor environment has a very high presence of visible light, modifying these wide bandgap materials to make them visible light-active photocatalysts is an elegant solution to our problem in hand.

Fig. 21.1  Mechanism of photocatalysis

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Fig. 21.2  Band-edge positions of some SCs. Reprinted with permission from [6]. Copyright @ Elsevier 2004

The required bandgap is determined by the redox potential of the chemical entity to be eliminated. For instance (Fig. 21.2), the VB potential of any SC must be higher than 2.6 V (vs NHE, pH = 0) for the generation of •OH radicals which completely mineralizes pollutant species. Similarly, the CB must be at −0.05  V (vs NHE, pH = 0) so that molecular oxygen reacts with the photo-produced electrons. In addition, some overpotentials are needed which increase the bandgap requirement, and thus a bandgap of more than 2.6  eV is required for real-time applications which corresponds to ~480 nm or lesser wavelength in the blue light regime. People are currently using many white LEDs in indoor environments which are composed of red, blue and green wavelength regions. If not the sun, then the blue wavelength region in the LED bulbs can excite photocatalysts in indoor environment. Therefore, researchers are more attracted towards the designing of visible light-driven photocatalyst materials for indoor environment that can effectively work in low-intensity indoor lighting, with minimal human intervention and at the same time cost-­effective for practical uses [7].

21.3  Visible Light-Driven Photocatalysts The conventional strategy of making visible light-active SCs is by doping with cationic or anionic entities, coupling with other SCs and plasmonic excitation. Though generally sol-gel method is used, and numerous methods for doping are also available. Doping of cations (Fig. 21.3) like Ba, Co, Cu, Cr, Ni, Fe, V, Mn, Cd, Ce, Ta, Ag, Bi, La, Nd, Mo, Nb [8–25] and many more into the host lattice is commonly used for reducing bandgap and inducing visible light excitation. This doped entity substitutes the host metal ion and introduces defect states near the CB which reduces

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Fig. 21.3  Formation of defect bands in SCs

the bandgap of the host material. Similarly, many other non-metallic elements like B, N, C, S, P, etc. [26–30] can also be doped. These entities lead to the formation of defect band near the VB. Although, the cationic and anionic defect bands lead to visible light absorption, a trade-off with QE is seen, and these materials usually have lower QE in visible light. At very low dopant concentrations, these defect bands are useful by enhancing charge separation, but after a threshold limit, these defect bands act as recombination centres by reducing mobilities acting as another factor for QE decrement. Electron and holes have an inherent tendency to lower their energy, and hence electron moves towards more positive potentials and the holes towards more negative potentials (vs NHE, pH = 0). Taking advantage of this behaviour, the photocatalytic efficiency can be increased by coupling one photocatalyst (Fig.  21.4) with another photocatalyst such that a small difference in the band-edge positions is present between the photocatalysts. An example of such a system is the anatase and rutile form of TiO2. Anatase has a bandgap of 3.2 eV and rutile has a bandgap of 3.0 eV. When these are coupled in such a way that the anatase/rutile is 80:20, we get enhanced photocatalytic activity as the photo-excited electrons in the CB of anatase can jump onto the CB of rutile increasing the e−/h+ carrier lifetimes by decreasing recombination. Such careful arrangement of the VB and CB of many photocatalysts (not only UV activated but also visible light activated) prevents e−/h+ recombination increasing their lifetimes by facilitating the transfer of electrons and holes from one SC to the other; however these strategies suffer from lower QE [31–34]. Similarly, the free electrons present in the metals oscillate collectively when an incoming light of similar energy is incident on these metals. Also known as surface plasma resonance, it causes excessive charge build-up or higher charge density to be present locally, and this favours electron transfer from the metal to the CB of the photocatalyst from where they are utilized for photocatalytic reactions. However, only some metals like Au, Ag, Pt and Cu etc. [35–38] show this effect, and not all metals are suitable for this effect to be noticed in the visible region. It is, therefore, understood that SC photocatalysts need electrons with high reduction power and holes with high oxidations power for effective and efficient photocatalytic activity. A brief overview of the current trend in the surface modification of the SCs is the matter of investigation for the current book chapter.

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Fig. 21.4 Schematic illustration of a coupled SC

21.3.1  Novel Strategy: Bandgap Engineering The conventional approach for the selection of a SC is based on the positions of the energy levels of VB and CB. For efficient degradation, deeper valance bands and higher conduction bands are needed. In fact, one can increase the degradation activity based on the position of the VB and CB. Pt-grafted WO3 shows excellent oxidation capability of acetaldehyde (pollutant) to harmless CO2 under visible light [39]. Since, the CB of WO3 (0.3–0.5 V vs SHE) is below the single-electron reduction of oxygen (−0.067 V vs SHE), the excited electrons cannot reduce molecular oxygen. However, the reduction potential is high enough for multi-electron reduction of oxygen (O2  +  2H+  +  2e  →  H2O2, 0.68  V; or four-electron reduction, O2  +  2H2O  +  4H+  +  4e  →  4H2O, 1.23  V), and it is this multi-electron reduction (MER) that helps in the degradation of pollutants. Frei et al. [40] first reported the Zr(IV)-O-Cu(I) bimetallic hetero-assemblies where the oxo-bridge acted as a photoinduced redox centre. When excited, the oxo-bridged redox centre promoted the electron transfer from the Zr(IV) to the Cu(I) forming Zr(III)-O-Cu(II) which in the presence of CO2 reduced to form CO.  Nakamura et  al. [41] carefully selected Ti(IV)-O-Ce(III) as a redox centre and reported that after excitation, the formation of Ti(III)-O-Ce(IV), confirmed by X-ray absorption near-edge structure (XANES) analysis, helps in the oxidative decomposition of 2-propanol into acetone and CO2. In addition, the quantum efficiency (QE) was reported to be higher by almost 9 times for acetone and 4.75 times for CO2 formation than that of N-doped TiO2. Since, the mononuclear Ti(IV) or Ce(III) samples did not catalyse any 2-propanol conversion, it was thus hypothesized that the metal-to-metal charge transfer (MMCT) induced by the visible light was responsible for the photocatalytic activity. Based on MMCT, Irie et  al. [42] proposed that the same phenomena, i.e. charge transfer, can be observed from atomic metal ions to the CB of TiO2 which are made up of 3d orbitals, and to prove it, Cr(III) was grafted on TiO2 as a Cr(IV)/Cr(III)

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redox couple which has a E0 of 2.1 V. Rather than the VB electron getting excited, the electron from the Cr(IV)/Cr(III) redox couple can directly be injected to the CB. Thus, the Cr(IV)/Cr(III) sites acted as the oxidation sites oxidizing pollutants, whereas the electron injected to the CB acted in reducing the molecular oxygen, thereby getting consumed. This MMCT is similar to interfacial charge transfer (IFCT) which was formulated by Hush et al. and Creutz et al. [43–45]. The same IFCT phenomenon coupled with MER was later used by Irie et al. to transfer excited VB electrons directly to the surface-grafted metal ions under visible light and reported Cu2+-grafted TiO2 and WO3 [46]. When visible light falls on pristine TiO2, the large bandgap prohibits any absorption and thus exhibits no catalytic activity. But when Cu2+ is grafted on TiO2, the excited valance band electrons are directly injected to the surface Cu2+ ions and lead to the formation of Cu1+. Since, Cu1+ is unstable, it again goes back to its former state by transferring the electron to the oxygen. Similarly, to explain the high photocatalytic activity of Cu2+-grafted WO3, it was reported that MER also takes place since the conduction band potential of WO3 is not enough to reduce oxygen through single-electron reduction. The current discussed strategies (Fig. 21.5) help in increasing the QE for conventional doping strategies from 3.9% of N-doped TiO2 to 8.8% of Cu2+-TiO2 and 17.5% of Cu2+-WO3. Similarly, loading elements like Fe and Ag have also been found to be a suitable candidate which exhibits the IFCT in addition to Cu2+, Cr3+ and Ce3+ which are explained in the following section. The Fe3+/Fe2+ reduction potential (0.771 vs SHE, pH = 0) is much higher than required for MER and follows a pattern which is similar to copper, where the Fe2+ gets converted back to Fe3+, and since it is having very high positive potential, it was expected to be active in the visible light (~500 nm range) [47]. Even Ag+/Ag has a reduction potential that is almost equal to that of Fe3+/Fe2+, and after excitation, the valance band electrons are directly transferred to the surface-grafted metal cations by IFCT, where electrons are used for multi-­ electron reduction (MER) [48]. Unlike doping, grafting entities on the surface does not lead to any structural changes. Surface modification has no influence on

Fig. 21.5  Various excitation mechanisms in (a) N-doped TiO2, (b) Cu2+-grafted TiO2 and (c) Cu2+grafted WO3 [46, 47]

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bandgap, but it significantly improves the charge separation by reducing the charge recombination (unlike doping) process which enhances the photocatalytic ability of the catalyst. As a follow-up to the existing methods, Kumar et al. [49] (Figs. 21.6 and 21.7) reported a Cu2+-modified ZnO photocatalyst. Prior, ZnO nanomaterials were synthesized by flame spray pyrolysis (FSP) followed by Cu2+ modification. Notably, surface-modified SCs showed no structural changes, indicating that no defect bands were formed in the SC after Cu2+ surface modification. The surface modification enables charge separation, in which electrons have high reduction potential and the holes have high oxidation power, which can lead to efficient degradation of pollutants. A similar observation was reported by Yin et al. [50], but under UV light, the presence of Cu2+ grafting on the surface of Nb3O8− enhanced the production of CO. Under light irradiation, photogenerated electrons are excited to the CBs and moved to Cu2+ nanoclusters which act as electron sink and get converted to Cu1+. Since, Cu1+ is highly unstable, electrons are transferred to reduce CO2 to form CO.  On the other hand, generated holes were used for water oxidation to form molecular oxygen. Shoji et al. [51] also reported a CuxO-modified SrTiO3 for photocatalytic CO2 reduction reactions. It was reported that the mix of the valance states of Cu was crucial for the CO2 reduction process. Though the QE was not reported, the CO2 utilization of CuxO-modified SrTiO3 was reported to have been almost double than that of pure SrTiO3 under the same conditions. Kumar et al. [52] reported a Fe3+-modified ZnO (Fig. 21.8) which was used for killing E. coli pathogen under the visible light illumination. The Fe3+ ion-induced IFCT in the visible light coupled with the action of the surface-grafted Fe3+ as a co-­ catalyst promoted multi-electron reduction effectively and improved the photocatalytic activity of these photocatalysts. The reduction reactions with oxygen give rise to many reactive oxygen species (ROS) which helped in the oxidative destruction of the E. coli cell wall. Another point to be noted is that living microbial entities can negate the effect of these ROS because of the presence of antioxidants secreted by them. Thus, a low concentration of ROS may not cause significant damage, but higher concentration of these ROS can cause damage. Since, the E. coli bacteria were completely destroyed under the visible light illumination, it can be inferred that there was a high concentration of ROS present and thus these surface-modified photocatalysts show exceptional activity. Liu et al. [53] and Qiu et al. [54] also reported a Cu2+ ion-modified TiO2 where 2+ Cu was grafted on the TiO2 by a simple impregnation technique. The uniqueness of this method is that unlike the previous reported SCs, the current reported SC can function in the dark as an anti-bacterial agent. During the normal visible light illumination on catalyst, excited electrons are transferred to the surface Cu2+ by IFCT to form Cu1+ species. Since, Cu1+ is highly unstable and to get the stable form Cu2+, electrons are released that react with the outer cell wall of pathogens effectively destroying them. It was also reported that the presence of electrons in the Cu2+ states is more preferred due to their higher positive potential in comparison to the CB of

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Fig. 21.6 (a, b) SEM images, (c, d) TEM images and (e, f) UV DRS studies of ZnO nano-spheres and ZnO nanorods synthesized by flame spray pyrolysis, respectively. Reprinted with permission from [49]. Copyright @ American Chemical Society 2014

TiO2. A small possibility exists that some of amount of these electrons can be trapped at the interface of TiO2 and Cu2+ because of the different bond length at the surface creating defects in the bandgap leading to a small increase in the electron-­ hole pair lifetime [53, 54].

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Fig. 21.7  Schematic mechanism of the inactivation of E. coli bacteria. Reprinted with permission from [49]. Copyright @ American Chemical Society 2014

Fig. 21.8  Visible light activity of (a) control sample, (b) ZnO nanorod, (c) Fe-modified ZnO nanorod, (d) ZnO nano-sphere and (e) Fe-modified ZnO nano-sphere [52]

21.3.2  Surface Grafting of Doped Semiconductors Surface grafting with Cu2+, Fe3+ and Ag1+ is a promising concept, but the IFCT phenomena is limited only to the surface and metal oxide nanocluster interface. As described before, doping leads to bandgap reduction and induces visible light activation. But unintentional high doping forms narrow bandgap which increases recombination centres and results in lower photocatalytic ability. For oxidation reaction like VOC degradation, doping with anions leads to the formation of defect bands above the VB and reduces the QE of the photocatalyst; it is understood that modifying the VB would be an unfavourable choice, and it is of the best interest to not modify the VB of the SC. Thus, only the CB can be modified to induce visible

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light absorption where the cationic dopants promote the absorption of visible light by forming isolated defect states below the CB and also help in  narrowing the bandgap. Maintaining a strict conduction that VB should remain unchanged, Yu et al. [55] have developed a Ga3+- and W6+-doped Ti1−3xWxGa2xO2 where it was anticipated that the W 5d bands contribute to the VB of TiO2 which are composed mostly of 3d orbitals. Other requirements like charge neutrality and similar ionic radii  were almost met for the above two doped entities. Doping of the entities undoubtedly led to enhanced absorption in the visible region of the spectrum, but the QE of the Ti1−3xWxGa2xO2 was very low (