Nanophotocatalysis and Environmental Applications: Detoxification and Disinfection [1st ed.] 978-3-030-12618-6;978-3-030-12619-3

Table of contents :
Front Matter ....Pages i-xiii
Role of Nano-photocatalysis in Heavy Metal Detoxification (Ankita Mazumder, Souptik Bhattacharya, Chiranjib Bhattacharjee)....Pages 1-33
Solar Photocatalysis Applications to Antibiotic Degradation in Aquatic Systems (Margarita Jiménez-Tototzintle, Enrico Mendes Saggioro)....Pages 35-53
Biomass-Based Photocatalysts for Environmental Applications (Yean Ling Pang, Chin Woei Lim, Katrina Pui Yee Shak, Steven Lim, Wai Chong Cheam, Chai Hoon Koo et al.)....Pages 55-86
Application of Bismuth-Based Photocatalysts in Environmental Protection (Ewa Maria Siedlecka)....Pages 87-118
Phosphors-Based Photocatalysts for Wastewater Treatment (Olga Sacco, Vincenzo Vaiano, Diana Sannino)....Pages 119-138
Nanocarbons-Supported and Polymers-Supported Titanium Dioxide Nanostructures as Efficient Photocatalysts for Remediation of Contaminated Wastewater and Hydrogen Production (Kakarla Raghava Reddy, M. S. Jyothi, A. V. Raghu, V. Sadhu, S. Naveen, Tejraj M. Aminabhavi)....Pages 139-169
Investigation in Sono-photocatalysis Process Using Doped Catalyst and Ferrite Nanoparticles for Wastewater Treatment (Sankar Chakma, G. Kumaravel Dinesh, Satadru Chakraborty, Vijayanand S. Moholkar)....Pages 171-194
Magnetic-Based Photocatalyst for Antibacterial Application and Catalytic Performance (Sze-Mun Lam, Jin-Chung Sin, Abdul Rahman Mohamed)....Pages 195-215
Antimicrobial Activities of Photocatalysts for Water Disinfection (Veronice Slusarski-Santana, Leila Denise Fiorentin-Ferrari, Mônica Lady Fiorese)....Pages 217-243
Medicinal Applications of Photocatalysts (Busra Balli, Aysenur Aygun, Fatih Sen)....Pages 245-265
Back Matter ....Pages 267-269

Citation preview

Environmental Chemistry for a Sustainable World

Inamuddin Abdullah M. Asiri Eric Lichtfouse Editors

Nanophotocatalysis and Environmental Applications Detoxification and Disinfection

Environmental Chemistry for a Sustainable World Volume 30

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

Other Publications by the Editors

Books Environmental Chemistry http://www.springer.com/978-3-540-22860-8 Organic Contaminants in Riverine and Groundwater Systems http://www.springer.com/978-3-540-31169-0 Sustainable Agriculture Volume 1: http://www.springer.com/978-90-481-2665-1 Volume 2: http://www.springer.com/978-94-007-0393-3 Book series Environmental Chemistry for a Sustainable World http://www.springer.com/series/11480 Sustainable Agriculture Reviews http://www.springer.com/series/8380 Journals Environmental Chemistry Letters http://www.springer.com/10311

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

Inamuddin • Abdullah M. Asiri • Eric Lichtfouse Editors

Nanophotocatalysis and Environmental Applications Detoxification and Disinfection

Editors Inamuddin Chemistry Department, Faculty of Science King Abdulaziz University Jeddah, Saudi Arabia

Abdullah M. Asiri Chemistry Department, Faculty of Science King Abdulaziz University Jeddah, Saudi Arabia

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

ISSN 2213-7114 ISSN 2213-7122 (electronic) Environmental Chemistry for a Sustainable World ISBN 978-3-030-12618-6 ISBN 978-3-030-12619-3 (eBook) https://doi.org/10.1007/978-3-030-12619-3 © Springer Nature Switzerland AG 2020 This work is subject to copyright. All rights are reserved by the Publisher, whether the whole or part of the material is concerned, specifically the rights of translation, reprinting, reuse of illustrations, recitation, broadcasting, reproduction on microfilms or in any other physical way, and transmission or information storage and retrieval, electronic adaptation, computer software, or by similar or dissimilar methodology now known or hereafter developed. The use of general descriptive names, registered names, trademarks, service marks, etc. in this publication does not imply, even in the absence of a specific statement, that such names are exempt from the relevant protective laws and regulations and therefore free for general use. The publisher, the authors, and the editors are safe to assume that the advice and information in this book are believed to be true and accurate at the date of publication. Neither the publisher nor the authors or the editors give a warranty, express or implied, with respect to the material contained herein or for any errors or omissions that may have been made. The publisher remains neutral with regard to jurisdictional claims in published maps and institutional affiliations. This Springer imprint is published by the registered company Springer Nature Switzerland AG The registered company address is: Gewerbestrasse 11, 6330 Cham, Switzerland

Preface

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Preface

In the recent decades, the world has encountered severe environmental challenges in terms of water pollution, climate change, and increasing infectious diseases. The irregular speed and quantum at which anthropogenic activities are growing have ended up in the wasting of resources, exposure to toxins, and pollution rise all over the world. These ubiquitous problems throughout the globe have been worsened in the developing part of the world due to the industrial revolution and poor monitoring and control of anthropogenic activities. The discharge of wastewater is increasing into water bodies and on land, and varieties of new emerging nonregulated pollutants are entering into the water bodies. The severe shortage of potable water requires more focus on recycling, remediation, and reuse. Conventional pollutant treatment and disinfection methods have their own disadvantages with efficiency and cost issues. The most dignified solutions for various environmental issues are resource management and cost-effective technologies for pollution mitigation and disinfection. Among various cost-effective and viable technologies used for mitigating environmental detoxification, advanced oxidation process-based photocatalysis is the most pioneer one. Principally, it is indeed a “green” technology thriving on light especially on naturally and abundantly available solar energy and oxygen from air under ambient conditions. The phenomena that came into the news with “Honda-Fujishima” effect proved to be one of most effective techniques to degrade persistent and emerging contaminants from aqueous medium utilizing light. After the secondary treatment of polluted water, photocatalysis is trusted for the next stage as it claims to mineralize the complex organic pollutants into water and carbon dioxide using hydroxyl and superoxide radicals. The scientific community, as well as the industrial players, has then utilized this to the maximum with continuous innovations in the designing of photocatalytic materials, pilot-scale experiments, and practical treatment of water using natural solar light. However, various new sources of production, disposals from medicine and plastics industry, agriculture, and household conveniences have led to huge widespread toxic pollutants. It has been an issue of concern for the general public as well as scientists and governments. On a broader line, these are classified as pharmaceuticals, personal care products, and endocrine disruptors. These have challenged the existing “star” photocatalytic materials and methodologies. New and highly advanced photocatalysts and improvement in existing photocatalytic technologies are thus required. Therefore, continuous research is going on various solar-active materials with multipronged capabilities. Another issue of environmental concern is increasing toward biohazards such as bacteria, virus, and other microorganisms in drinking water resources. This has increased the risk of deadly diseases in humans and aquatic species. Conventional disinfection methods are failed as many of these microbes developed resistance to ultraviolet, chlorination, and antibiotics too. Semiconductor-based photocatalysis is proved to be beneficial for water disinfection, i.e., inactivation and photo-killing of microbes via generation of reactive oxygen species. Hence, photocatalysis has been used for environmental detoxification and disinfection which includes organic pollutants, heavy metal, inorganic contaminants removal, and microbial killing.

Preface

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Shifting from single photocatalysts to heterojunctions, nanocomposites, molecularly imprinted catalysts, sensitized nanomaterials, photocatalytic membranes, and heterostructures has led to better results and higher efficiencies. Nanophotocatalysis and Environmental Applications: Detoxification and Disinfection focuses on existing and novel applications of photocatalysts in environmental detoxification as well as disinfection. Photodegradation of organic and inorganic pollutants including dyes, drugs, pesticides, hormones, heavy metals, antibiotics, microbes, bacteria, fungi, and carcinogens is discussed on various novel photocatalysts such as biomass- and phosphor-based photocatalysts, nanocarbonand polymer-supported nanostructures, ferrite nanoparticles, magnetic materials, etc. Based on thematic topics, the book edition contains the following 10 chapters: Chapter 1 intends to explore an overview of the mechanisms and promising research activities on photocatalytic nanoparticle-assisted heavy metal detoxification. Chapter 2 discusses the effect of parameters and pathways (transformation products) of solar photocatalysis of antibiotic groups usually found in aquatic systems such as macrolides, sulfones, lincosamides, and quinolone. Chapter 3 addresses the advancements to overcome the drawbacks of pure semiconductor oxide by the incorporation of biomass-derived carbonaceous materials for the fabrication photocatalysts used to remove organic pollutants from water. This chapter also highlights the types, properties, and conversion of biomass into biochar, activated carbon, or any other carbonaceous materials. The preparation of biomass-based support and the mechanisms of biomass-derived photocatalysis are discussed in detail. Chapter 4 aims to highlight recent advancements in the application of Bi-photocatalysts and its heterostructures used in environmental protection. The review explores photocatalytic degradation of antibiotics, nonsteroidal antiinflammatory drugs, beta-blockers, anticonvulsant, hormones, resorcinol, bisphenol A, and derivatives available in aqueous systems. The applications of Bi-based photocatalysts for treating NOx, water splitting, and CO2 reduction to CO and CH4 are discussed. Chapter 5 describes solutions for enhancing the photon distribution inside the photoreactors using inorganic and organic light-emitting particles (phosphors)) coupled with photocatalysts. This chapter also underlines the difference between inorganic particles having down-conversion, up-conversion, and long afterglow luminescence properties. Additionally, the use of up-conversion organic phosphors is proposed. Finally, some examples concerning the use of semiconductors coupled with different photoluminescent materials in the removal of pollutants from water and wastewater are discussed. Chapter 6 explores the applications of TiO2 nanostructures (0D to 3D) functionalized with various polymeric and nanocarbon hybrid photocatalytic materials for photodegradation of chemical pollutants. Various chemical synthesis methods, surface modification with various polymers and nanostructured carbons, composition, morphological structures, growth mechanism, physicochemical properties, electronic and optical characteristics, and photocatalytic mechanism of

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various heterostructured TiO2-based photocatalysts are discussed in details. The future challenges in the fields of photocatalytic environmental remediation and hydrogen generation are also mentioned. Chapter 7 gives a brief introduction of nanomaterials including their classification, shape and structure, type of nanomaterials, and applications in degradation of recalcitrant organic contaminants. The process intensification using sono-hybrid advanced oxidation processes of sono-photocatalysis and heterogeneous Fentonlike reaction for wastewater treatment is discussed. Chapter 8 explores a brief review of various magnetic-based photocatalytic nanomaterials used for photocatalytic disinfection and degradation processes. The factors influencing the catalytic performance along with the disinfection mechanisms are also discussed. Chapter 9 presents a review of the literature on the various types of photocatalytic materials, their mechanism of action for photocatalytic water disinfection, and photocatalysts with microbial activity. Chapter 10 discusses the importance of photocatalysts and their medicinal application in the daily life of human beings. The properties of photocatalysts in relation to the nanoscale are discussed. The medicinal applications of photocatalysts such as antifungal, antimicrobial, anti-cancerogenic are also discussed in detail. This book is the consequence of the commendable cooperation of authors from various interdisciplinary fields of science. It thoroughly examines the most generous, start to finish, and forefront research and reviews. We are thankful to all the contributing authors and their coauthors for their regarded commitment. We may moreover need to thank all copyright holders, authors, and other individuals who agreed to use their figures, tables, and schemes. Yet every effort has been made to secure the copyright approvals from the individual proprietors to consolidate reference to the imitated materials, we should need to offer our sincere proclamations of disappointment to any copyright holder if unintentionally their benefit is being infringed. Jeddah, Saudi Arabia Jeddah, Saudi Arabia Aix en Provence, France

Inamuddin Abdullah M. Asiri Eric Lichtfouse

Contents

1

Role of Nano-photocatalysis in Heavy Metal Detoxification . . . . . . . Ankita Mazumder, Souptik Bhattacharya, and Chiranjib Bhattacharjee

2

Solar Photocatalysis Applications to Antibiotic Degradation in Aquatic Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Margarita Jiménez-Tototzintle and Enrico Mendes Saggioro

3

Biomass-Based Photocatalysts for Environmental Applications . . . . Yean Ling Pang, Chin Woei Lim, Katrina Pui Yee Shak, Steven Lim, Wai Chong Cheam, Chai Hoon Koo, and Ahmad Zuhairi Abdullah

4

Application of Bismuth-Based Photocatalysts in Environmental Protection . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Ewa Maria Siedlecka

1

35 55

87

5

Phosphors-Based Photocatalysts for Wastewater Treatment . . . . . . 119 Olga Sacco, Vincenzo Vaiano, and Diana Sannino

6

Nanocarbons-Supported and Polymers-Supported Titanium Dioxide Nanostructures as Efficient Photocatalysts for Remediation of Contaminated Wastewater and Hydrogen Production . . . . . . . . . . . 139 Kakarla Raghava Reddy, M. S. Jyothi, A. V. Raghu, V. Sadhu, S. Naveen, and Tejraj M. Aminabhavi

7

Investigation in Sono-photocatalysis Process Using Doped Catalyst and Ferrite Nanoparticles for Wastewater Treatment . . . . . . . . . . . 171 Sankar Chakma, G. Kumaravel Dinesh, Satadru Chakraborty, and Vijayanand S. Moholkar

ix

x

Contents

8

Magnetic-Based Photocatalyst for Antibacterial Application and Catalytic Performance . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 195 Sze-Mun Lam, Jin-Chung Sin, and Abdul Rahman Mohamed

9

Antimicrobial Activities of Photocatalysts for Water Disinfection . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 217 Veronice Slusarski-Santana, Leila Denise Fiorentin-Ferrari, and Mônica Lady Fiorese

10

Medicinal Applications of Photocatalysts . . . . . . . . . . . . . . . . . . . . . 245 Busra Balli, Aysenur Aygun, and Fatih Sen

Index . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 267

Contributors

Ahmad Zuhairi Abdullah School of Chemical Engineering, Universiti Sains Malaysia, Nibong Tebal, Penang, Malaysia Tejraj M. Aminabhavi Soniya College of Pharmacy, Dharwad, Karnataka, India Aysenur Aygun Sen Research Group, Biochemistry Department, Faculty of Arts and Science, Dumlupınar University, Kütahya, Turkey Busra Balli Sen Research Group, Biochemistry Department, Faculty of Arts and Science, Dumlupınar University, Kütahya, Turkey Chiranjib Bhattacharjee Department of Chemical Engineering, Jadavpur University, Kolkata, India Souptik Bhattacharya Department of Chemical Engineering, Jadavpur University, Kolkata, India Sankar Chakma Department of Chemical Engineering, Indian Institute of Science Education and Research Bhopal, Bhopal, Madhya Pradesh, India Satadru Chakraborty Department of Chemical Engineering, Indian Institute of Science Education and Research Bhopal, Bhopal, Madhya Pradesh, India Wai Chong Cheam Department of Chemical Engineering, Lee Kong Chian Faculty of Engineering and Science, Universiti Tunku Abdul Rahman, Kajang, Selangor, Malaysia G. Kumaravel Dinesh Department of Chemical Engineering, Indian Institute of Science Education and Research Bhopal, Bhopal, Madhya Pradesh, India Leila Denise Fiorentin-Ferrari Department of Chemical Engineering, West Paraná State University, Toledo, Paraná, Brazil Mônica Lady Fiorese Department of Chemical Engineering, West Paraná State University, Toledo, Paraná, Brazil xi

xii

Contributors

Margarita Jiménez-Tototzintle Sanitation and Environment Health Department, National School of Public Health, Oswaldo Cruz Foundation, Rio de Janeiro, RJ, Brazil M. S. Jyothi Department of Basic Sciences, School of Engineering and Technology, JAIN Deemed-to-be University, Bangalore, Karnataka, India Chai Hoon Koo Department of Civil Engineering, Lee Kong Chian Faculty of Engineering and Science, Universiti Tunku Abdul Rahman, Kajang, Selangor, Malaysia Sze-Mun Lam Department of Environmental Engineering, Faculty of Engineering and Green Technology, Universiti Tunku Abdul Rahman, Kampar, Perak, Malaysia Chin Woei Lim Department of Chemical Engineering, Lee Kong Chian Faculty of Engineering and Science, Universiti Tunku Abdul Rahman, Kajang, Selangor, Malaysia Steven Lim Department of Chemical Engineering, Lee Kong Chian Faculty of Engineering and Science, Universiti Tunku Abdul Rahman, Kajang, Selangor, Malaysia Ankita Mazumder Department of Chemical Engineering, Jadavpur University, Kolkata, India Abdul Rahman Mohamed School of Chemical Engineering, Universiti Sains Malayisia, Nibong Tebal, Pulau Pinang, Malaysia Vijayanand S. Moholkar Department of Chemical Engineering, Indian Institute of Technology Guwahati, Guwahati, Assam, India S. Naveen Department of Basic Sciences, School of Engineering and Technology, JAIN Deemed-to-be University, Bangalore, Karnataka, India Yean Ling Pang Department of Chemical Engineering, Lee Kong Chian Faculty of Engineering and Science, Universiti Tunku Abdul Rahman, Kajang, Selangor, Malaysia A. V. Raghu Department of Basic Sciences, School of Engineering and Technology, JAIN Deemed-to-be University, Bangalore, Karnataka, India Kakarla Raghava Reddy School of Chemical & Biomolecular Engineering, The University of Sydney, Sydney, NSW, Australia Olga Sacco Department of Industrial Engineering, University of Salerno, Fisciano, SA, Italy V. Sadhu School of Physical Sciences, Banasthali Vidyapith, Banasthali, Rajasthan, India

Contributors

xiii

Enrico Mendes Saggioro Sanitation and Environment Health Department, National School of Public Health, Oswaldo Cruz Foundation, Rio de Janeiro, RJ, Brazil Diana Sannino Department of Industrial Engineering, University of Salerno, Fisciano, SA, Italy Fatih Sen Sen Research Group, Biochemistry Department, Faculty of Arts and Science, Dumlupınar University, Kütahya, Turkey Katrina Pui Yee Shak Department of Chemical Engineering, Lee Kong Chian Faculty of Engineering and Science, Universiti Tunku Abdul Rahman, Kajang, Selangor, Malaysia Ewa Maria Siedlecka Department of Chemistry, University of Gdańsk, Gdańsk, Poland Jin-Chung Sin Department of Petrochemical Engineering, Faculty of Engineering and Green Technology, Universiti Tunku Abdul Rahman, Kampar, Perak, Malaysia Veronice Slusarski-Santana Department of Chemical Engineering, West Paraná State University, Toledo, Paraná, Brazil Vincenzo Vaiano Department of Industrial Engineering, University of Salerno, Fisciano, SA, Italy

Chapter 1

Role of Nano-photocatalysis in Heavy Metal Detoxification Ankita Mazumder, Souptik Bhattacharya, and Chiranjib Bhattacharjee

Contents 1.1 1.2

1.3 1.4 1.5

1.6 1.7

1.8 1.9

1.10

1.11 1.12

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Heavy Metals and Their Toxicological Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2.1 Cadmium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2.2 Chromium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2.3 Copper . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2.4 Lead . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2.5 Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2.6 Nickel . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2.7 Zinc . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Overview of Photocatalysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Mechanism of Photocatalysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Types of Photocatalysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.5.1 Homogeneous Photocatalysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.5.2 Heterogeneous Photocatalysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Overview and Mechanism of Nano-photocatalysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Photocatalytic Nanoparticle Synthesis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.7.1 Organic Synthesis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.7.2 Chemical Synthesis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.7.3 Physical Synthesis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Mode of Operation on Nano-photocatalysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Parameters Affecting the Photocatalytic Efficiency . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.9.1 Effect of pH of the Reaction Solution . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.9.2 Effect of Photocatalyst Concentration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.9.3 Effect of Substrate Adsorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.9.4 Effect of Dissolved Oxygen . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Application . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.10.1 Chromium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.10.2 Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.10.3 Arsenic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.10.4 Uranium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Disadvantages of Photocatalysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Photocatalyst Modifications . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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A. Mazumder · S. Bhattacharya · C. Bhattacharjee (*) Department of Chemical Engineering, Jadavpur University, Kolkata, India e-mail: [email protected] © Springer Nature Switzerland AG 2020 Inamuddin et al. (eds.), Nanophotocatalysis and Environmental Applications, Environmental Chemistry for a Sustainable World 30, https://doi.org/10.1007/978-3-030-12619-3_1

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A. Mazumder et al.

1.12.1 Dye Sensitization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.12.2 Ion Doping . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.12.3 Composite Semiconductor . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.13 Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract Heavy metal elimination from contaminated waterbodies is a worldwide concern. Their inherited property of causing toxicity with the metabolism of both aquatic and terrestrial life forms has been revealed. These metals are highly poisonous for flora and fauna even if they are present at very low concentrations due to their bioaccumulative capabilities. The detoxification of such contaminants is considered to be a prime concern in ecological remediation. Surface chemistry, photocatalysis and chemical engineering fundamentals in combination with advanced nanoengineered technologies spark up attractive prospects towards the treatment of water resources with heavy metal contamination. Products obtained through photocatalysis-assisted engineered nanotechnology have successfully proven upgraded performance characteristics like increased surface area, easy recovery, reuse capability, better volumetric potential, reduced mechanical stress, enhanced shelf life, excellent sorption behaviour, stability under operational conditions, easy separation, no secondary pollution and many more. The present chapter is contemplated to display an overview of the mechanisms and promising research activities on photocatalytic nanoparticles assisted heavy metal detoxification. This will thereby provide researchers and individuals with a profound intuition and bridge the research gaps of the exquisite nanomaterial-induced heavy metal removal techniques. Keywords Nano-photocatalysis · Semiconductor photocatalysts · Bandgap energy · Photocatalytic nanoparticle synthesis · Photocatalyst modifications · Photocatalytic efficiency · Heavy metal toxicity · Bioaccumulation · Neurologic degeneration · Environmental remediation

1.1

Introduction

Recently water contamination and water scarcity because of the rapid ongoing urbanization and industrialization are worldwide extreme environmental issues. As water is the most essential and fundamental requirement of human sustenance, the insufficient and limited availability of fresh water is enhancing the environmental burden, which is in turn negatively affecting the economic development, environmental balance and human wellbeing. To address the adverse environmental effects of the accelerated utilization of natural resources in a developing economy, the European Union has adopted a sustainable strategy on the use of water, with the objective of mitigating the gradual depletion and contamination of water (European

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Commission 2005). Approximately 40% of the total world population is severely affected by the water scarcity, which therefore triggers the foremost necessity of treating polluted water and producing clean reclaimed water (Calzadilla et al. 2011). It has been reported that above 600 varieties of organic, inorganic and biological pollutants were found in the different effluents each having different toxicity levels (Gupta and Ali 2013). Amidst these pollutants, inorganic heavy metals are highly toxic in nature and are associated with alarming health risk affecting the ecological balance even if present in very trace concentrations. The heavy metal is the naturally occurring element having higher density, atomic weight in the range of 63.5–200.6 and high specific gravity (>5.0) (Fergusson 1990; NK Srivastava and Majumder 2008). As metal ions are non-biodegradable and generally non-biotransformable, they persist in the environment for a long duration. Moreover, even if these heavy metals exist in water in very trace level, bioaccumulation and bioconcentration can result in their accumulation at high concentration in different organisms of the food chain. Therefore, the heavy metal pollution is a serious and vital environmental concern because of their high flexibility, non-biodegradable, persistence and accumulation properties (Peligro et al. 2016). These pollutants are mainly found in industrial effluents though sometimes remain present in drinking water as well. The major contributors of heavy metals in the environments are industries such as pesticides, paper, tanneries, mining, metal plating, etc. Among these heavy metals, some of them are quintessential for the metabolic activities of living organisms if consumed in very less amount only. However, in most cases, the intake of these metals in larger quantity causes incurable and severe toxicity ultimately leading to the malfunctioning of the human physiology (Koedrith et al. 2013; Nordberg et al. 2007). The toxicity of these contaminants can eventually diminish the energy level and the proper functionality of prime organs such as the brain, kidney, lungs, liver, etc. Moreover, the toxicity level is highly dependent on the absorbed dosage, exposure duration and route of exposure. The prolonged exposure to elevated concentrations of potentially dangerous heavy metals results in continuous degeneration of physical, neurological and muscular ability causing chronicle diseases like Alzheimer’s, multiple sclerosis, Parkinson’s, cancer and muscular dystrophy (Jaishankar et al. 2014). When these metals are exposed to the aqueous medium, aquatic organisms can easily absorb those metals either from water or consume sediments or other aquatic lives (plants or microorganisms) already contaminated with heavy metals. Following the subsequent mechanism of bioaccumulation and bioconcentration, the concentration of those metals gradually becomes very high in those organisms in comparison with the water (Abida et al. 2009). These toxic components ultimately show adverse effects critically affecting the aquatic organisms’ survival rate, growth, metabolic activities or reproduction ability. In this context, stringent laws and regulations are being forced by Environmental Protection Agency (EPA) and World Health Organization (WHO) to control these metals’ toxicity so that their discharge in the surroundings is mandatorily kept well below the fixed permissible limit. Thus the need to develop a promising water treatment method for efficient heavy metals removal becomes a bloom in the worldwide research domain.

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Several technologies studied for either removal or recovery of toxic heavy metals include coagulation-flocculation, chemical precipitation, adsorption, membrane filtration, ion exchange, electrochemical treatment, etc. But these conventional technologies investigated for detoxification of heavy metals are expensive and highly energy intensive and generate high loads of waste by-products. In recent times, photocatalysis has gained enormous research attention and emerged as a proficient technology for reduction of heavy metal to non-toxic state. This technology involves the use of light energy to excite the semiconductor photocatalyst for the formation of electron-hole pairs that eventually facilitates the detoxification of heavy metals. The main positive attribute involved with photocatalysis is the requirement of readily available sustainable energy source (solar energy) for carrying out the photocatalytic activity. Moreover, photocatalytic reactions take place in moderate temperature and atmospheric pressure, aided with an additional advantage when the reactants or generated products show highly heat sensitiveness and explosive nature (Hennig and Billing 1993). However, the main intricacy of photocatalysis is the very fast electron-hole pairs recombination in huge quantity, resulting in the reduced photocatalytic activity (Rothenberger et al. 1985). Henceforth, worldwide research efforts have been executed to mitigate this limitation. One of the recent advancements in this field is to reduce the size of semiconductor photocatalysts in nanoscale for sufficiently elevating the specific surface area and a total number of active sites, which will eventually improve the photocatalytic activity. Thus, the concept of implementing nanostructured semiconductor photocatalysts evolved for better photocatalytic efficiency.

1.2

Heavy Metals and Their Toxicological Effects

Some of the most abundantly available poisonous heavy metals in the environment are arsenic, nickel, chromium, cadmium, zinc, cobalt, antimony, copper, etc. which are highly detrimental posing the toxic threat to the environment. The level of adverse toxic effect largely depends on the degree of biomagnification. Generally, different metal ions such as Pb2+, Cd2+, Hg2+, As3+ and Ag+ get reacted with biomolecules to produce harmful and poisonous components that show intricacy in case of kidney filtration. The probable health hazards of these aforesaid toxic heavy metals are discussed briefly in this section.

1.2.1

Cadmium

Cadmium appears as a poisonous heavy metal, mostly discharged in nature through electroplating industries, metal-finishing industries, metallurgical activities, batteries, solders, stabilizers, insecticides, fertilizers, alloy, etc. The human exposure to carcinogenic cadmium even at trace concentration can cause hepatic toxicity,

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respiratory problem, lung cancer, kidney failure, liver damage, reproductive disabilities, etc. (Alslaibi et al. 2013; Filipič 2012). The contamination of cadmium in the river even leads to a disease named ‘itai-itai’ which shows symptoms like bone degeneration, the bones softening and fractures (Nishijo et al. 2017).

1.2.2

Chromium

The major contributors of chromium to the environments are both the natural sources (volcano, weathering phenomenon of rocks, soils and sediments) and anthropogenic ones (fossil-fuel burning, plastic production, chromates production, metals electroplating and overutilization in the leather, tannery and textile industries). Chromium coexists in the industrial effluents in hexavalent and trivalent form (Miretzky and Cirelli 2010). Though Cr(III) is essential in the metabolism of fat and poses a less toxic threat, Cr(VI) is around 300 times more toxic and harmful for living plants and organisms compared to Cr(III) (Krishnani and Ayyappan 2006). The primary source of Cr(VI) is the production of chromate salt. Skin inflammation, ulcer, vomiting, kidney and liver failure and pulmonary congestion are some of the diseases caused by chromium (Cieślak-Golonka 1996; Kotaś and Stasicka 2000).

1.2.3

Copper

The use of copper is very common since ancient days. However, recently, the elevated usage of toxic copper sulphate (CuSO4) in both industrial and agricultural activities (as a fungicide, pesticide and pathogen killer) has significantly enhanced its concentration in waterbodies (Özer et al. 2004). It is a necessary element, required at very trace quantity, for the proper functional activity of some specific protein and enzymes (Fraga 2005). But if being taken at high concentrations, its negative impact is severely detrimental for human health. It can exist in various forms including metal (Cu0), cuprous ion (Cu+) and cupric ion (Cu2+). The most available form of its existence in the environmental mediums is in cupric ion state, which is comparatively more toxic and harmful than other forms (Awual et al. 2014). The toxicity level of copper sulphate varies largely with the type of aquatic species and is very dangerous for fish. In the case of humans, overexposure can result in irritation, vomiting, nausea, loss of physical strength, diarrhoea, stomach damage and kidney failure (Uriu-Adams and Keen 2005). Continuous exposure for long duration leads to the accumulation of copper in different organs of the human body such as the brain and liver, ultimately causing death.

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Lead

Lead is a heavy, dense and highly poisonous metal. It is discharged in significant amount in the effluents of different types of industries such as electrical, electroplating, explosive, steel industries, etc. (Acharya et al. 2009; Jalali et al. 2002). It is considered that the most dangerous source point of lead contamination is lead-acid battery effluent. The lead poisoning can cause serious health conflicts including vomiting, physical weakness, abdominal cramping, pain, seizures, nervous system damage, mental retardation, coma, encephalopathy, cancer and reproductive system failure to the human being (Gupta et al. 2001; Mudipalli 2007; Sanborn et al. 2002). The toxicity of lead can also critically affect the animals and living plants.

1.2.5

Mercury

The mercury pollution incident in Minamata Bay of Japan has created worldwide alertness. Mothers of that region who were overexposed to mercury pollution because of the consumption of mercury-accumulated fish are giving birth to physically disabled, deformed and mentally challenged babies as a consequence of mercury pollution (Harada 1995). Mercury usually exists in three main forms, namely, elemental mercury (Hg0), mercuric ion (Hg2+) and mercurous ion (Hg22+). It can get easily transported in waterbodies and accumulate in food chain organisms of an ecosystem. In this context, Minamata Convention is adopted in 2013 to save ecological species from the harmful and dangerous effect of mercury. According to this convention, the products containing mercury as its constituents have been listed, and stringent emission standards have been formulated (Mackey et al. 2014; Wu et al. 2016). The substantial quantity of mercury is discharged with the wastewater of industries such as chloro-alkali, pharmaceutical, paper and pulp, plastic, oil refineries, etc. (Boening 2000). Some of the most probable negative health effects of mercury are kidney failure, brain damage, respiratory system malfunction, reproductive system dysfunction, etc. (Parham et al. 2012).

1.2.6

Nickel

Attributing to the beneficial properties of nickel such as the ability to form an alloy, high resistivity to corrosion, high strength and durability and good thermal and electrical conductivity, it is enormously used in different types of industries. Out of the total nickel consumption, the huge fraction (around 75%) of it is solely utilized for alloy production (Deevi and Sikka 1996). Some of the well-known health effects caused by the overutilization of nickel include breathing problem, asthma, dry

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cough, skin irritation, chest pain, nausea, vomiting, diarrhoea, gastrointestinal problem, pulmonary fibrosis, etc. (Das et al. 2008; Mobasherpour et al. 2012).

1.2.7

Zinc

Zinc is essential in nutritional supply at a very trace concentration for the proper physiological functioning of living tissues. However, if it is present in high amount in the nutritional diet, its effect becomes adverse causing several health issues such as vomiting, nausea, pain, fever, skin irritation and inflammation, anaemia, etc. (Carolin et al. 2017; Cristian et al. 2015). Zinc is majorly utilized in steel industries, paper and pulp industries, brass metal works industries and electroplating industries. Therefore, zinc enters in the ecological environment chiefly from human activities like mining, coal combustion, waste burning, steel production, etc. Only a little fraction of it is released from natural sources. In the environment, it gets adsorbed to soil, sediments and dust particles in the air, and then some zinc compounds consequently transfer into the aqueous medium (groundwater and surface water). When zinc concentration gets slightly enhanced, bioaccumulation occurs in fish and other aquatic organisms and thus badly affects the aquatic ecology by even bringing the death of organisms (Emsley 2011; Megateli et al. 2009).

1.3

Overview of Photocatalysis

The word ‘photocatalysis’ consists of the affix ‘photo’ (Greek phos, light) and ‘catalysis’ (Greek katalyo, break apart). A catalyst is a material which has the ability to reduce the free activation enthalpy of a chemical reaction to accelerate the reaction rate. In an illumination-induced reaction or photoreaction, application of a catalyst to enhance the reaction rate is defined as photocatalysis. Much like the normal catalysts, a photocatalyst does not get involved in the chemical transformation, but it promotes the reaction. Unlike the conventional thermal catalysts which are activated by heat, the photocatalyst is activated by photons of desired energy. The photocatalysis may work in two different modes when the catalyst (C) is activated by photons, (1) energy transfer and (2) electron transfer. In energy transfer mode, the photon-activated photocatalysts (C*) transfer its energy into the reactant/ substrate (S) which, in turn, activates the reactant to undergo oxidation (Fig. 1.1). This is difficult to achieve when the reactant is in its ground state. In electron transfer mode, the photon-activated photocatalysts work either as an electron acceptor or a donor. In this case, oxidation and reduction both take place with the activated photocatalysts to achieve reaction equilibrium (Fig. 1.2). Increased surface area, reusability, moderate stability, proper bandgap and material morphology are significant attributes of the photocatalytic systems. Photocatalysis is widely used to degrade hazardous organic compounds to natural elemental

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Fig. 1.1 Energy transfer mechanism

Fig. 1.2 Electron transfer mechanism

compounds such as CO2 and H2O and to reduce toxic metallic ions into non-toxic states. Apart from these, eradication of water-borne germs, decomposition of the air pollutants such as NOx and CO, synthesis of significant industrial chemicals and plastic degradation are the major points of application of photocatalytic systems (Singh and Ikram 2016). Organic pollutant degradation by the photocatalytic system was studied thoroughly. However, inorganics especially heavy metals in their various oxidation states have infinite lifetimes and are not degradable at all. Chemical and biological treatments of heavy metals are restricted due to not being environmentally friendly and also economically prohibitive. Moreover, less attention was given to transform toxic heavy metals into less toxic forms by photocatalysis.

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1.4

9

Mechanism of Photocatalysis

Any photocatalyst, upon exposure to the light energy (UV-VIS) in the form of ultraviolet radiation, visible light or fluorescent illumination, gets activated. The energy within the light excites the electrons of the catalyst’s valence band (VB) and promotes them towards conduction band (CB). This creates the negative-electron (J. Yang & Lee) zone at conduction band and positive-hole (h+) zone at the valence band of the photocatalyst. This is termed as semiconductor’s photoexcitation state. The difference in the energy within these two bands is known as ‘bandgap’ or ‘forbidden energy zone’. Pairs of h+ and e
generated by photo-induced energy react with water or atmospheric oxygen and form aggressive radical such as hydroxyl (OH
) or superoxide. These radicals are capable of oxidizing organic and inorganic materials. However, to generate OH
radicals, the redox potential of the valence band hole has to be enough positive, and on the other hand, highly negative redox potential of the conduction band electron is demanding for the formation of superoxide radicals (Samsudin et al. 2015). The wavelength of the light required to supply the bandgap energy is found by the following equation: E ¼ hν ¼ hc=λ

ð1:1Þ

(Here, E is the energy of bandgap; ν, frequency of light; h, Planck’s constant; λ, wavelength of light; and c, speed of light.) The steps of photocatalysis include (i) formation of electron-hole pairs by absorbing light; (ii) separation of excited charges; (iii) transfer of electrons and holes to photocatalysts surface; and (iv) redox reactions by surface charge (Fig. 1.3).

Fig. 1.3 Schematic diagram of photocatalytic redox processes on the photoexcited semiconductor. (Reproduced from Lim and Goei (2016) with permission from the Royal Society of Chemistry)

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Photocatalytic degradation of hazardous organic compounds, toxic inorganic materials and microbial contamination has been proven as an effective alternative with respect to conventional processes.

1.5

Types of Photocatalysis

Based on the reactant and photocatalyst’s state, photocatalysis can be divided into two subgroups (Ameta and Ameta 2016). When the substrate is already photochemically active, it is known as homogeneous photocatalysis, and if the substrate is otherwise photochemically inactive, then it is known as heterogeneous photocatalysis.

1.5.1

Homogeneous Photocatalysis

When reactant and photocatalyst reside in the identical phase, the reaction is called homogeneous photocatalysis. Dyes, pigments and coordination compounds are the major examples of such photocatalysts (Wu and Chang 2006). Ozone (O3) and photo-Fenton systems (Fe+ and Fe+/H2O2) are widely known as common homogeneous photocatalysts. •OH is the major reactive species and is employed for several applications. The reaction of ozone is denoted below: O3 þ hν ! O2 þ O ð1DÞ

ð1:2Þ

O ð1DÞ þ H2 O ! • OH þ • OH

ð1:3Þ

O ð1DÞ þ H2 O ! H2 O2

ð1:4Þ

H2 O2 þ hν ! • OH þ • OH

ð1:5Þ

Moreover, the mechanism for Fenton system reaction is the following: Fe2þ þ H2 O2 ! HO • þ Fe3þ þ OH


ð1:6Þ

Fe3þ þ H2 O2 ! Fe2þ þ HO • 2 þ Hþ

ð1:7Þ

Fe2þ þ HO • ! Fe3þ þ OH


ð1:8Þ

Furthermore, in the photo-Fenton process, photolysis of H2O2 produces additional sources of OH radicals, and Fe3+ ion reduction takes place under ultraviolet light. The specific conversions are mentioned below:

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H2 O2 þ hν ! HO • þ HO •

ð1:9Þ

Fe3þ þ H2 O þ hν ! Fe2þ þ HO • þ Hþ

ð1:10Þ

However, operating parameters such as H2O2 concentration, pH and ultraviolet intensity enforce influence on the efficiency of Fenton processes.

1.5.2

Heterogeneous Photocatalysis

When the photocatalysts and the reactant reside in a different phase, the reaction is called heterogeneous photocatalysis. Transition metal chalcogenides are very wellknown compounds that could perform heterogeneous photocatalysis. Heterogeneous photocatalysis can be augmented in the gas phase, liquid phase and aqueous solution. The overall heterogeneous catalysis process is composed of the following steps: 1. 2. 3. 4. 5.

Reactants transfer from the fluid phase to the surface. Adsorption of the reactants. The reaction takes place in the adsorbed phase. Desorption of the product(s). Removal of the products from the interface area.

Heterogeneous photocatalysis includes a variety of reactions such as oxidation, dehydrogenation, exchange of deuterium-alkane isotope, metal deposition, polluted water detoxification, atmospheric pollutant removal and environmental remediation procedures (Peternel et al. 2007). Metal oxides such as TiO2, ZnO, ZrO2, MoO3, Fe2O3, WO3, SnO2 and SrTiO3 and chalcogenide metals such as ZnS, ZnO, CdS, WS2, CdSe and MoS2 are some common photocatalysts (Fig. 1.4). They can generate a bulk amount of electron-hole pairs. These pairs will eventually undergo dissociation to form free photoelectrons in the conduction band and photo holes in the valence band when illuminated with photons from the light source. This phenomenon can only take place when the light energy is either equivalent to or greater than their respective bandgap energy (hν  bandgap energy) (Zhu and Wang 2017).

1.6

Overview and Mechanism of Nano-photocatalysis

Solid-phase transformation and electron/hole (e
/h+) dynamics of any photocatalyst are significantly dependent upon the particle size. Small particle size reduces volume recombination of (e
/h+), which increases interfacial charge carrier transfer. So, maximum photocatalytic activity lies in the nanometre range of the material. This

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Fig. 1.4 Band positions of various semiconductors and redox potentials of the various processes associated with photocatalysis in water at various pH. (Reproduced from Lim and Goei (2016) with permission from the Royal Society of Chemistry)

phenomenon is known as quantum confinement. Nanophotocatalysts have some assertive positive qualities over the catalysts with large particle size: 1. The ratio of surface area to volume is very high, which results in an increased amount of active sites on the photocatalyst surface. 2. By changing the size of the nanocatalyst, it is possible to tune and modify the catalyst’s absorption wavelength and electronic and optical properties (Radhika et al. 2016).

1.7

Photocatalytic Nanoparticle Synthesis

1.7.1

Organic Synthesis

1.7.1.1

Plant Extracts Aqueous Solutions

Metal ions can be reduced to nanoparticles by biomolecules present in plant extracts. It is a short-step nature-friendly synthesis process and environmentally benign. The reducing agents present in the plant extract are various enzymes, coenzymes and plant metabolites (e.g. alkaloids, terpenoids, phenolic compounds, etc.). This biogenic process is rapid and can be done in room temperature at normal atmospheric pressure (Fatimah et al. 2016; Goutam et al. 2018; Mittal et al. 2013).

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Fig. 1.5 Overview on different types of synthesis methods to produce a variety of nanoparticles. (Reproduced from Dhand et al. (2015) with permission from the Royal Society of Chemistry)

Plant extracts of Polyalthia longifolia, Pelargonium graveolens, Crossandra infundibuliformis, Cinnamomum zeylanicum, Datura metel, Eclipta prostrata, Eucalyptus hybrid, Jatropha curcas, Desmodium triflorum, Cinnamomum camphora and Mimosa pudica were used to create photocatalytic metal nanoparticle reported by various researchers (Figs. 1.5, 1.6 and 1.7).

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Fig. 1.6 Schematic representation of radiative (left) C and nonradiative (right) decay of particle plasmons in noble metal nanoparticles. The nonradiative decay occurs via excitation of electronhole pairs either within the conduction band (intraband excitation) or between the d-band and the conduction band (interband excitation). (Reprinted with permission from Sönnichsen et al. (2002). Copyright 2002, by the American Physical Society) Fig. 1.7 Pictorial representation of dye sensitization mechanism for visible light-induced photocatalytic reaction. (Figure reproduced with permission from Maeda et al. (2008). Copyright 2008, American Chemical Society)

1.7.1.2

Microorganisms

Microbes are capable of reducing metal ions into nanoparticles. This process is economic. A bacterial innate ability such as the reactive nature of intracellular proteins and other biomolecules possesses the potential to control inorganic crystal agglomeration at the time of biomineralization processes. Recently, bacterial species such as Bacillus cereus, Zooglea ramigera, Pseudomonas aeruginosa and Rhizopus oryzae; fungal species like Aspergillus flavus, Bryophilous and Rhizoctonia; and various other species have been explored for producing metal nanoparticles (Nishant Srivastava and Mukhopadhyay 2014).

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1.7.2

Chemical Synthesis

1.7.2.1

Sol-Gel Method

15

The sol-gel process uses precursors like metal alkoxides and metal chlorides and undergoes several types of hydrolysis and polycondensation reaction mechanisms. In the first step, metal-oxo or metal-hydroxo polymers in the solution are generated. Afterwards, when the solution turns into a gel-like diphasic system with various morphologies, the remaining solvent phase is removed by the drying process. Finally, a thermal treatment is needed to polycondensate and to intensify the mechanical properties and structural integrity of the product. The sol-gel method is cost-effective and requires low temperature which gives it an advantage for wide application. Sol-gel process-derived materials are used in energy, optics, electronics, biosensors, space, medicine (nanodrug), reaction material and separation (membrane, chromatography) processes (Tseng et al. 2010).

1.7.2.2

Hydrothermal Method

This method uses varying pressure and temperature conditions of supercritical water to control the characteristics of nanoparticles. It can be performed in both batch hydrothermal process, which allows undergoing a system with the desired ratio phases, and in the continuous hydrothermal process which achieves a high reaction rate within a short period of time.

1.7.2.3

Polyol Synthesis

Polyethene glycols hold qualities such as excellent solvent, stabilizing agent, reducing agent and complexing agent properties, which can be used as a reaction medium for synthesis of metallic nanoparticles. This process is extensively utilized to synthesize a vast range of metal oxide nanoparticles like ZnO, Gd2O3, Cu2O and indium tin oxide, metal nanoparticles and hybrid metal nanoparticles.

1.7.2.4

Precipitation Method

In this method, precipitation technique controls the aggregation of small-scale crystallites. In this synthesis procedure, constituent material reacts in a suitable solvent to form nanoparticles. Nonaqueous solvents generally control the Oswald ripening and coagulation due to heat at lower temperatures of crystallites. Before precipitation, a suitable dopant is added to the parent solution. Addition of favourable surfactant maintains dissociation of the particles formed. Centrifugation

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followed by washing and vacuum dry process is used to separate newly formed nanocrystals.

1.7.3

Physical Synthesis

1.7.3.1

Ball Milling

Ball milling process involves the transfer of kinetic energy of the moving balls to the milled materials. This produces smaller particles of nanorange with newly created surfaces.

1.7.3.2

Melt Mixing

In this method, a polymer is mixed mechanically with modified nanofillers using extrusion or kneading. This method is convenient for modern industrial practices due to its eco-friendly nature. This process is one of the oldest which designs polymer composites with nanoparticles as the fillers for achieving desired material characteristics.

1.7.3.3

Physical Vapour Deposition (PVD)

PVD is an eco-friendly method related to vacuum deposition process. There are three fundamental steps of this technique: (1) material vaporization from a solid source, (2) transportation of the vaporized material and (3) nucleation and growth to create thin films and nanoparticles.

1.7.3.4

Laser Ablation

The laser beam (continuous or pulsed) of high energy can evaporate particles from a solid source. Nanoparticles formed using this technique are of very elegant quality. This technique is employed for synthesis of nanoparticles of metal oxide (TiO2, Al2O3, SiO2, Fe2O3), non-oxide (Si, Si3N3, SiC, MoS2) and ternary composites such as Si/Ti/C and Si/C/N.

1.7.3.5

Sputter Deposition

Vacuum-based PVD process is known as sputtering. Sputter deposition technique undergoes the below-mentioned steps:

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1. Generation of plasma of neutral gasses (argon). 2. Ions present in the plasma are directed towards the target material using the potential difference between two electrodes. 3. The collision of the ions with the target material leads to ejection of tiny particles from the target. 4. The transportation and deposition of ejected nanorange materials into the desired substrate for collection (Dhand et al. 2015).

1.8

Mode of Operation on Nano-photocatalysis

The change in the particle size changes scattering to absorption ratio to a large extent. Particles of larger size show elevated efficiency in scattering light. Absorption of light by particles of smaller size is the reason behind the colour we see. The two major excitation models for this operation are Localized surface plasmon resonance (LSPR) and interband transition. In the localized surface plasmon resonance model, light interacts with particles much smaller than the incident wavelength. This results in a plasmon that oscillates locally around the nanoparticle with a certain frequency. Interband transition occurs because of transitions between deeper bands (d-bands of noble metals) and conduction band. Coinage metal nanoparticles generally possess characteristic localized surface plasmon resonance absorption in or near the visible-light region. However, morphology modification and increasing particle size to hundreds of nanometres can promote non-coinage metals to express localized surface plasmon resonance absorption. The population decay for localized surface plasmon resonance occurs through the transformation of particle plasmons into photons (radiation damping) and via nonradiative decay into electron-hole excitations. The latter process falls into two categories: intraband excitations within the conduction band and interband excitations. Normally, large nanoparticles decay mostly through radiative damping (e.g. larger than 50 nm for dipolar plasmon resonances), and for small nanoparticles, nonradiative damping is the major decay mechanism (Halas et al. 2011; Liu et al. 2017).

1.9

Parameters Affecting the Photocatalytic Efficiency

Various parameters that affect the reaction kinetics and the efficiency of photocatalysis are discussed here.

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Effect of pH of the Reaction Solution

pH is a very significant factor influencing the surface charge, particle size, band edge position of TiO2 and the energy of both valence and conduction bands. For example, in case of Degussa P25 TiO2, the determined point of zero charges (PZC) is at a pH value of 6.9. If the pH is greater than pHPZC, photocatalyst surface gets negatively charged, and thereby, the interaction between photocatalyst and cationic electron donor and acceptor gets facilitated. On the other hand, if pH is less than pHPZC, photocatalyst surface is positively charged and is favourable for anionic electron donors and acceptors (Dutta et al. 2004).

1.9.2

Effect of Photocatalyst Concentration

Determination of the optimum photocatalyst concentration for any photocatalytic reaction is essential for both maximizing the reaction rate and avoiding photocatalyst wastage. The initial reaction rate of the photocatalytic process is directly proportional to the photocatalyst concentration up to a certain optimized concentration, beyond which this relationship does not hold good. Moreover, the concentration of photocatalyst is also depended on the shape, geometry, type and working mechanism of photocatalytic reactors. As observed from previous studies, the maximum concentration of TiO2 required for different reactors such as fixed bed reactor and the batch reactor is not same (1.3 mg/cm2and 2.5 g/L, respectively). At very high photocatalyst loading, illumination of the maximal quantity of TiO2 takes place, leading to light scattering and reduced photon efficiency (Dutta et al. 2004). Henceforth, it is recommended to optimize the photocatalyst concentration for the photocatalytic process.

1.9.3

Effect of Substrate Adsorption

Since heterogeneous photocatalysis is a surface phenomenon occurring at the photocatalyst surface, the photocatalytic efficiency largely depends on the substrate’s ability to adsorb or get adsorbed onto the semiconductor photocatalyst surface. The efficiency of the photocatalysis process will be very low if the substrate displays incapability in getting adsorbed onto the photocatalyst surface. On the contrary, when the substrate gets easily adsorbed onto the photocatalyst surface, oxidation/reduction of more substrate will occur because of the electron/hole at the surface of the photocatalyst, resulting in high efficiency of the photocatalytic process (Kormann et al. 1991).

1 Role of Nano-photocatalysis in Heavy Metal Detoxification

1.9.4

19

Effect of Dissolved Oxygen

In photoreduction processes, the presence of dissolved oxygen significantly reduces the reduction rate. When oxygen is present in the system, it competes with the metal ions for the photogenerated electrons, resulting in a limited electron availability for the metal reduction as compared to the metal reduction in the absenteeism of dissolved oxygen (Litter 1999). Henceforth, it is important to ensure the absence of dissolved oxygen to enhance metal reduction process by purging the solution with inert gasses such as nitrogen or argon.

1.10

Application

Advanced oxidative or reductive processes (AOPs/ARPs) are simple, less expensive treatment procedures for removal of heavy metals. Amidst these processes, heterogeneous photocatalysis is a well-established method. As discussed earlier, this procedure involves the promotion of electrons from valence band to conduction band during irradiation of the semiconductor with a light source of sufficient excitation energy. This transfer of electrons leaves holes behind. Therefore the photocatalytic reduction of an acceptor by the electrons of conduction band (e
conduction band) occurs when its redox potential becomes more positive than e
conduction band, whereas the oxidation of donor is possible by the holes of the band (h valence band+) if its redox potential is more negative than h valence band+. Application of heterogeneous photocatalysis for the removal of heavy metals is briefly illustrated here with some suitable examples.

1.10.1 Chromium Since 1979 several research studies have been reported for chromium reduction employing photocatalysis under UV/visible light illumination (Yoneyama et al. 1979). Titanium dioxide (TiO2) is the extensively explored semiconductor due to its inherent suitable properties such as relatively less expensive, high stability, efficient photocatalytic characteristics and photostable and supportive chemical properties. Both pure and modified TiO2 have been investigated for Cr reduction (Mu et al. 2010; Wang et al. 2008). Some more semiconductors like ZnO, CdS, ZnS and WO3 were also explored for heterogeneous photocatalytic activities. Recently photocatalytic degradation using the modified TiO2 under the influence of solar light has gained significant research attention. The redox potentials of TiO2 for different chromium species are shown below. It has been observed that multiple e–conduction band s have the sufficient potential for direct reduction of Cr(VI), Cr(V) and Cr(IV).

20

A. Mazumder et al.

However, beyond Cr(IV) further reduction is not feasible. The overall reaction of Cr (VI) degradation is mentioned below (Khalil et al. 1998). 2Cr2 O7 2
þ 16Hþ ! 4Cr3 þ þ 8H2 O þ 3O2 Reactions involved in the photocatalytic reduction of Cr using TiO2 photocatalysts are given below (Litter 2017): CrðVIÞ þ eCB
! CrðVÞ E0 ðCrðVIÞ=CrðVÞÞ ¼ 0:55 V

ð1:11Þ

CrðVIÞ þ R • ! CrðVÞ þ Rox

ð1:12Þ

CrðVÞ þ eCB
! CrðIVÞ E0 ðCrðVÞ=CrðIVÞÞ ¼ 1:34 V

ð1:13Þ

CrðVÞ þ R • ! CrðIVÞ þ Rox

ð1:14Þ




CrðIVÞ þ eCB ! CrðIIIÞ

0

E

ðCrðIVÞ=CrðIIIÞÞ

¼ 2:10 V

ð1:15Þ

CrðIVÞ þ R • ! CrðIIIÞ þ Rox

ð1:16Þ

2 CrðVÞ ! CrðVIÞ þ CrðIVÞ

ð1:17Þ

2 CrðIVÞ ! CrðVÞ þ CrðIIIÞ

ð1:18Þ

CrðIVÞ þ CrðVÞ ! CrðVIÞ þ CrðIIIÞ

ð1:19Þ

CrðVÞ=CrðIVÞ=CrðIIIÞ þ

• hþ VB =HO

! CrðVIÞ=CrðVÞ=CrðIVÞ

CrðIIIÞ þ R • ! Rox þ CrðIIÞ !! Crð0Þ E0 ðCrðIIIÞ=CrðIIÞÞ ¼
0:42 V

ð1:20Þ ð1:21Þ

The rapid advancement of nanotechnology has promoted the use of TiO2 nanoparticles and nanofibres enriched with novel functionalities that shows potential enhanced photocatalytic activity for Cr(VI) reduction (Mu et al. 2010). Though synthesis of the TiO2 nanoparticles and nanofibres are mainly carried out by electrospinning technology, an alternative, versatile and cheap synthesis method named electrophoretic deposition (EPD) can be successfully utilized for depositing the dense layer of TiO2 onto the surface of electrodes required in the photocatalytic Cr(VI) reduction (Pifferi et al. 2013). It has been observed that deviation in the photocatalytic efficiency of Cr(VI) reduction are strongly dependent on structural morphology and composition of the TiO2 nanofilm at varying heating temperatures and coating cycles (Kajitvichyanukul and Amornchat 2005). Shaham Waldmann and Paz 2010 achieved approximately 86% of Cr(VI) photoreduction by utilizing semiconductor TiO2 in the modified form of nanofibres under the influence of ultraviolet light at acidic pH after 2 hours of irradiation (Shaham Waldmann and Paz 2010). Various parameters that significantly affect the efficiency of photocatalysis for Cr reduction are discussed here.

1 Role of Nano-photocatalysis in Heavy Metal Detoxification

1.10.1.1

21

pH

The photocatalysis mechanism of Cr (VI) reduction is highly pH sensitive. The highest efficiency of Cr(VI) reduction was evident below pH 3 when it gets converted to Cr(III). However, with the increase in pH, the reduction rate gradually decreases. The large deviation in reduction rate has been evident in between pH 3 and pH 4, which is probably attributed because of pH-dependent alterations in surface or substrate interactions and a conversion from the dichromate anion to chromate. The relatively higher and significant reaction efficiency was achieved at acidic pH than at neutral one. At pH greater than 4, fouling of the photocatalysts occurred because of the deposition of stable precipitate of Cr (III) (Prairie et al. 1993; Schrank et al. 2002; Yang and Lee 2006). Moreover, the thermodynamic investigation of Cr(VI) photoreduction employing TiO2 at varying pH in the range of 0–10 also manifested the similar result concluding that thermodynamic driving force for Cr(VI) reduction was higher in acidic solutions than in basic one (Lin et al. 1993).

1.10.1.2

Light Intensity

The photon-induced photocatalytic reduction of Cr(VI) is significantly affected by the intensity of light. The reduction rate of Cr(VI) was observed linearly proportional with the intensity of light till certain optimum intensity of light was achieved. After reaching certain optimum light intensity, reduction rate does not increase further with the light intensity. This occurs mainly due to the fact that the light intensity becomes extremely amplified in comparison to less number of TiO2 particles present in the solution to excite extra electrons and holes on the TiO2 surface (Khalil et al. 1998; Ku and Jung 2001).

1.10.1.3

Photocatalyst Dosage

The amount of the photocatalyst present in the reaction solution is a dominant parameter controlling the photocatalytic removal of Cr(VI). Generally, the reaction rate for reduction of Cr(VI) elevates with the rise in photocatalyst dosage. As the dosage of photocatalysts elevates, both the number of adsorption sites on the photocatalyst and the formation of free electrons in conduction band also increases, thus facilitating the photocatalytic activity. Literature shows that there is a certain optimal value for the catalyst dosage, beyond which decrease in reduction rate is observed. For the photocatalytic reduction process of Cr(VI) utilizing UV/CdS, the optimum dosage of CdS was around 5 g/l. However, Cr(VI) reduction rate starts to decline gradually because of the addition of CdS at a higher dosage than optimum dosage (Ku and Jung 2001). This negative effect is caused as the layer of excessive photocatalyst particles can, in turn, reduce the penetration of UV light.

22

1.10.1.4

A. Mazumder et al.

Presence of Organic Compounds

It is very well known that the photocatalyst has the ability to carry out oxidation and reduction of organic compounds and metal ions, respectively. To verify the practical application of photocatalysis in water treatment, the effect of organic materials present in the reaction solution on the photocatalytic reduction of Cr(VI) was evaluated. It was found that because of the presence of both the organic compounds (humic acid, ethylenediaminetetraacetic acid (EDTA), oxalate, phenol, nitrilotriacetic acid, etc.) and Cr(VI) in water, the reduction rate for both organic and metal gets elevated to some extent (Prairie et al. 1993). The photocatalytic reduction of Cr(VI) in association with the simultaneous photocatalytic oxidation of added organics results in the facilitation of the Cr(VI) reduction rate, which is majorly attributed because of the synergistic effect between organic matter and Cr (VI) (Wang et al. 2008). Organic contaminants act as the donor of electrons and even as a-holes scavenger in the valence band of the semiconductor. Though the rate of adsorption of Cr(VI) onto the photocatalyst surface could be hindered by the organic matter, elevated reduction rate and photocatalytic efficiencies are the results of the decreased possibility of recombination between electrons and the positive holes. Therefore, this above-explained fact nullifies the negative impact on the photocatalytic removal of Cr(VI) (Yang and Lee 2006). However, there is specific optimal value for the number of organic compounds up to which the acceleration of Cr(VI) reduction takes place. Beyond that limiting value of the concentration of organics, reduction rate shows reverse behaviour. At high concentration of organic matter, absorption of the UV light is very intense so strongly and acts as an internal filter. The reduction of Cr(VI) depends substantially on the specific nature of organic compounds. The photocatalytic reaction systems comprising alcohols, acids and aldehydes of low molecular weight showed almost equivalent reduction rates to those systems without any organic components. On the contrary, the presence of organics such as EDTA, salicylic acid and citric acid in the reaction system displayed very accelerated reduction rate of Cr(VI) (Lin et al. 1993).

1.10.2 Mercury Numerous studies have been carried out using several photocatalysts such as metal oxides (TiO2, ZnO) and metal sulphides (cadmium sulphide (CdS), pyrite (FeS2) and zinc sulphide (ZnS)) under visible, UV and solar irradiation. Among them, TiO2 photocatalyst exhibits the best photocatalytic efficiency in mercury removal. This mechanism is accomplished by reducing the state of Hg(II) to a less toxic state. Then it is either deposited or adsorbed by TiO2, followed by the extraction mechanism using mechanical or chemical methods to such a form which can be safely disposed or treated subsequently (Zlamal et al. 2007).

1 Role of Nano-photocatalysis in Heavy Metal Detoxification

23

The overall stoichiometric reaction for the phenomenon of mercury deposition is given below (De Lasa and Serrano-Rosales 2009): 1 HgðIIÞ þ H2 O ! 2Hþ þ O2 þ Hgð0Þ 2

ð1:22Þ

The reductive pathway followed for the reduction of Hg engaging consecutive electron charge transfer reactions is depicted below: HgðIIÞ þ e
! HgðIÞ

ð1:23Þ

HgðIÞ þ e
! Hgð0Þ

ð1:24Þ

However, the main intricacy of the photocatalytic reduction of mercury is the high chances of reoxidation of reduced Hg species by photogenerated positive holes or hydroxyl radicals are possible, as mentioned below (De Lasa and Serrano-Rosales 2009): Hgð0=IÞ þ holeþ =HO • ! HgðI=IIÞ

ð1:25Þ

Some of the well-established approaches employed to reduce and mitigate the recombination rate of the charge carriers (positive holes and photogenerated electrons) are doping of novel materials on the surface of TiO2 photocatalyst or surface deposition of metal oxide and the incorporation of an electron scavenger or hole to the reaction system. These approaches mainly aim to create repulsion and thereby diverging both the photogenerated electrons and holes away from the metal oxide surface. Thus, the lifespan of charge carriers gets enhanced, providing sufficient time to initiate the oxidation and reduction reactions, respectively (Tan et al. 2003). Research investigations revealed that the mercury reduction efficiency is dependent on various significant parameters, namely, illumination intensity, pH, catalyst dosage, irradiation time, initial contaminant concentration, substrate adsorption, intermediates reactivity, the addition of organic donors, the presence of oxygen, etc. (De Lasa and Serrano-Rosales 2009). Litter et al. carried out a comparative study of photocatalysis for treating three different mercuric salts (Hg (NO3)2, HgCl2 and Hg(ClO4)2). From the kinetic study, the initial reduction rate of Hg(II) over irradiation time was very fast, followed by a decline in the reduction rate. Moreover, the highest reduction rate was achievable at higher pH levels for all the three types of Hg salts (De Lasa and Serrano-Rosales 2009). It has been deduced from numerous research work that the addition of sacrificial donors or hole scavengers to the reaction system can substantially enhance the photocatalytic efficiency. Amidst various investigated chemicals as hole scavengers, EDTA showed promising reducing ability as it can make highly stable complexes in association with Hg(II). Some other studied hole scavengers are sodium dodecyl sulphate (SDS), ethanol, citrate, formic acid, cetyltrimethylammonium bromide, etc.

24

A. Mazumder et al.

(De Lasa and Serrano-Rosales 2009). Some of the additional advantages of formic acid use are its small size, ease to get adsorbed onto the photocatalyst for direct oxidation and the generation of water and carbon dioxide as photoproducts of oxidation of formic acid not causing any considerable environmental hazards.

1.10.3 Arsenic In this section, the detailed mechanism of the treatment of arsenic by photocatalysis will be discussed. Photocatalysis of both As(V) and of As(III) with the most general nano-photocatalyst TiO2 has been highlighted. The transformation of As(III) can be carried out either by oxidative or reductive photocatalytic mechanism. Oxidation process follows the reaction mechanism as mentioned in Eq. (1.26). This involves the consecutive one-electron steps, involving hVB+ or HO as the major oxidant in the oxidation reaction for converting As(III) to As(IV) (E0 » +2.4 V) (Klaening et al. 1989). However, it has been observed from stopped-flow studies that thermodynamically it is not possible to directly reduce As(V) to As(IV) by e–CB since E0(As(V)/As (IV)) » –1.2 V (Klaening et al. 1989). The reduction of As(V) can be done in the presence of methanol (MeOH) as shown in Eqs. (1.29) and (1.30) (Litter et al. 2014; Yang et al. 1999). Equations (1.31), (1.32), (1.33) and (1.34) show the generation of As(III). It has been studied that As(III) reduction by TiO2 e–CB (as shown in Eq. 1.35) is feasible both in the presence and absence of MeOH. Further reduction of As may result in the generation of As(0) and AsH3 and other products, which were not precisely distinguished and identified (Levy et al. 2012). Therefore, reduction of As(V) to solid zerovalent arsenic (As(0)) was accomplished which marks a very promising and auspicious achievement in terms of removing As from water. However, for scaling up of this process, proper optimization of reaction conditions of the photocatalytic system should be carried out to escape the generation of AsH3. Mnþ þ hVB þ =HO • ! Mðnþ1Þþ CH3 OH þ hVB þ fHO • g! • CH2 OH þ Hþ fH2 Og

ð1:26Þ

E 0 ð • CH2 OH=CH3 OHÞ

¼ 1:45 V •

ð1:27Þ

CH2 OH ! CH2 O þ Hþ þ eCB


E0 ð • CH2 OH=CH2 OÞ  -0:9 to-1:18 V

AsðVÞþ • CH2 OH ! AsðIVÞ þ CH2 O þ Hþ

E 0 ðAsðVÞ=AsðIVÞÞ  -1:2 V

AsðVÞ þ CO2 •
! AsðIVÞ þ CO2


AsðIVÞ þ eCB ! AsðIIIÞ E

0

ðAsðIVÞ=AsðIIIÞÞ

¼ 2:4 V

AsðIVÞþ • CH2 OH ! AsðIIIÞ þ CH2 O þ Hþ

ð1:28Þ ð1:29Þ ð1:30Þ ð1:31Þ ð1:32Þ

1 Role of Nano-photocatalysis in Heavy Metal Detoxification

25

AsðIVÞ þ CO2 •
! AsðIIIÞ þ CO2

ð1:33Þ

2AsðIVÞ ! AsðVÞ þ AsðIIIÞ

ð1:34Þ

AsðIIIÞ þ eCB
! AsðIIÞ

E0 ðAsðIIIÞ=AsðIIÞÞ ¼ ?

ð1:35Þ

1.10.4 Uranium Direct photocatalytic reduction of U(VI) to U(V) by e–CB (Eq. 1.36) is thermodynamically possible. Moreover, in the presence of formic acid (FA), the formation of CO2
takes place because of hole attack, which ultimately facilitates in U (VI) reduction (Eq. 1.37). However, this mechanism is not possible in the presence of 2-PrOH (Litter 2017; Salomone et al. 2015). A further change of state of U(V) can be carried out by either disproportionation (Eq. 1.38) or by re-reduction to U (IV) (Eqs. 1.39 and 1.40). Formed U(IV) can either precipitate or exist in solution. U3O8, UO2 + x and other reduced uranium oxides are some of the insoluble final products, which result in the photocatalyst deactivation (Amadelli et al. 1991). Furthermore, if the resulted suspension comes in contact with oxygen, reoxidation can occur leading back to the generation of U(VI). The photocatalytic efficiency of the photocatalyst can be regained by removing the insoluble uranium products. Consecutive reduction of U(IV) to U(III) by TiO2 e–CB is not permissible as the redox potential (E0(U(IV)/U(III))) is equal to (
0.52 V). The generation of U(III) formation by CO2
is possible if FA is present in the reaction solution (Eq. (1.41) (Salomone et al. 2015). Any further reduction of U(III) to U(II) is not thermodynamically feasible as the redox potential (E0(U(III)/U(II))) couple is negative (¼
2.53 V) (Kulyukhin et al. 2006). It has been observed that the effect of O2 is adverse for the reduction of U(VI). This detrimental effect is due to the very rapid reaction of O2 with uranium reduced species, but not because of the competition between O2 and U(VI) for e–CB (Salomone et al. 2015). Reactions involved in the photocatalytic reduction of uranium using TiO2 photocatalyst are given below (Litter 2017): UðVIÞ þ eCB
! UðVÞ E 0 ðUðVIÞ=UðVÞÞ ¼ 0:16 V UðVIÞ þ CO2

•


! UðVÞ þ CO2

ð1:36Þ ð1:37Þ

2 UðVÞ ! UðVIÞ þ UðIVÞ

ð1:38Þ

UðVÞ þ eCB
! UðIVÞ E 0 ðUðVÞ=UðIVÞÞ ¼ 0:38 V

ð1:39Þ

UðVÞ þ CO2

•


! UðIVÞ þ CO2

ð1:40Þ

26

A. Mazumder et al.

UðIVÞ þ CO2 •
! UðIIIÞ þ CO2

1.11

E0 ðUðIVÞ=UðIIIÞÞ ¼
0:52 V

ð1:41Þ

Disadvantages of Photocatalysis

Apart from the positive aspects of photocatalysis, some intricacies associated with their implementation for heavy metal removal are the rapid charge carrier recombination reactions that restrict the photocatalytic efficiency, the least utilization of visible light due to the limitation posed by band edge absorption threshold and the rupture of the used catalyst terminating further reactions (Hennig and Billing 1993; Rothenberger et al. 1985; Serpone et al. 1994). To mitigate these drawbacks, some schemes have been proposed to increase the capacity of light absorption and augment the catalyst lifespan. The widened wavelength absorption spectra can be accomplished by using strategies such as dye sensitization, bandgap tailoring or external surface modification (Anandan et al. 2010).

1.12

Photocatalyst Modifications

1.12.1 Dye Sensitization Electromagnetic radiation produced by fusion process in the sun consists of three types of radiation, namely, visible, ultraviolet (UV) and infrared radiation. The bulk fraction of the solar spectrum entering the earth’s surface is in the visible range with only a narrow range in the near-infrared and near-ultraviolet region. As discussed previously, the photocatalyst in the heterogeneous photocatalysis process is constantly irradiated with light inheriting higher energy level compared to the photocatalysts bandgap energy. When no more recombination of the electron-hole pair is occurring, their migration takes place to the surface of the catalyst. Consequently, they participate in the degradation of pollutant through redox reactions. Since only 4% of the broad solar spectrum is in the UV range, the photocatalyst used in the solar light irradiated heterogeneous photocatalytic processes can use only a limited portion of the total solar light. To address this problem, the research interest in the future of semiconductor photocatalysis mainly aims towards the usage of visible light (approximately 46% of the total solar spectrum) for the excitement of semiconductors. The two major drawbacks associated with the use of visible light are poor activation of the photocatalyst and rapid recombination of electron-hole pairs. To mitigate these disadvantages, some of the techniques employed include doping with either anions or cations, implantation of metal ions and coupling with a semiconductor having a narrow bandgap (Hodes and Graetzel 1984). However, these techniques are not cost-effective and time-consuming (Chowdhury 2012).In

1 Role of Nano-photocatalysis in Heavy Metal Detoxification

27

this context, another technique named dye sensitization has been developed to promote the photocatalytic activity of the catalyst. Dye sensitization is a novel, simple, inexpensive and efficient technique, which enables photocatalyst to utilize longer wavelengths than UV (i.e. visible light). In dye sensitization, the dye gets chemically adsorbed on the surface of the semiconductor and thus generates a dye-sensitized semiconductor film which has the capacity to adsorb the visible light efficiently. Here, chemisorbed dyes act as the spectral sensitizer promoting the visible light adsorption. On the exposure of dye to the visible light, it gets excited, and eventually oxidation reaction proceeds, injecting an electron into the conduction band of the semiconductor. This phenomenon is called anodic sensitization. These released electrons in the conduction band are then utilized by the metal ions for carrying out the reduction process. The mechanism of dye sensitization is illustrated using a suitable example. In this example, TiO2, eosin Y (EY) dye and triethanolamine (TEOA) are considered as the photocatalyst, sensitizer and electron donor, respectively. During sensitization process, EY dye should remain in its ground state, as the dye excitation is only feasible from its ground state. The release of electrons to the conduction band of semiconductor TiO2 takes place only from the excited state of the dye (EY*). To regenerate the dye from its oxidized state (EY+), TEOA is utilized as an electron donor. The main action of TEOA is to reduce the eosin Y dye to its original ground state while getting itself oxidized. The overall dye sensitization process is represented in few steps with respective reactions. 1. Excitation of dye by the adsorption of visible light as shown in Eq. (1.42). 2. Sensitization of TiO2 because of the excited dye species (EY*) as depicted by Eq. (1.43). 3. Regeneration of the oxidized dye (EY+) using TEOA as the electron donor (see Eq. 1.44). 4. The last step is the change of state of metal ion to its non-toxic form as shown in Eq. (1.45). visible light source

EY








! EY EY ! TiO ðe - Þ þ EYþ

ð1:42Þ

EYþ þ TEOA ! EY þ TEOAþ

ð1:44Þ

Cd2 þ 2e
! Cd0

ð1:45Þ

2

CB

ð1:43Þ

28

A. Mazumder et al.

1.12.2 Ion Doping Doping pure semiconductor (TiO2) with either cations or anions is an alternative approach to elevate the ability of TiO2 for improved response in the visible range and to increase its photocatalytic efficiency. In this route, an electron donor level is created in between valence band and conduction band by the dopant, which in turn reduces the bandgap. Thus, the light absorption range gets elongated, and semiconductor becomes active in the visible region. Cation doping of TiO2 generally involves doping of noble metals, rare earth metals, transition metals, poor metals or post-transition metals. However, anion doping is carried out by introducing anionic non-metals such as N, C, I and S (Daghrir et al. 2013).

1.12.3 Composite Semiconductor In this route of modifying semiconductor, the coupling of semiconductors of varying bandgap is carried out. Here, a semiconductor of the wide bandgap is coupled with a narrow-bandgap semiconductor where electrons are impregnated to the conduction band of wide-bandgap semiconductor from the conduction band of a narrowbandgap semiconductor (Marschall 2014).

1.13

Conclusion

As discussed in the chapter, the utilization of photocatalytic nanoparticles for the detoxification of heavy metals from wastewater is a considerable progress for research domain. The photocatalysis can be implemented effectively for detoxification of various heavy metals from contaminated waters. Major mechanisms and synthesis procedures are well documented in the recent researches. However, several technical glitches are involved with this process and have to overcome those challenges in making nano-photocatalysis a competent wastewater treatment technology in the near future. A huge emphasis is being given on scaling up the technology from laboratory-scale experiments to real-life application, biocompatibility improvement of the nanoparticles and making it eco-friendly as well as costeffective. Acknowledgement The authors acknowledge the Council of Scientific and Industrial Research (CSIR), New Delhi, India, for providing financial support in terms of Senior Research Fellowship (SRF) scheme (file no. 09/096(0879)/2017-EMR-I).

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Chapter 2

Solar Photocatalysis Applications to Antibiotic Degradation in Aquatic Systems Margarita Jiménez-Tototzintle and Enrico Mendes Saggioro

Contents 2.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2 Solar Photocatalysis Process . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3 Solar Photocatalysis Treatment for Antibiotic Degradation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3.1 Trimethoprim . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3.2 Sulfamethoxazole . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3.3 Erythromycin . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3.4 Ciprofloxacin . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.4 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract Contaminant such as pesticides, veterinary products, industrial compounds, food additives, personal care products, and pharmaceuticals may cause a negative effect when comes into contact with the environment. The presence of these compounds can be generated toxic effects in the aquatic system and cause an irreparable ecological alteration. Antibiotic compounds are one of the main pollutants found in the aquatic systems, because they are often inadequately prescribed and as a part of antibiotics is not completely consumed or degraded in human and animal bodies. Their residues can be entered in aquatic systems by wastewater treatment plants. The presences of antibiotics in aquatic systems have been linked to increasing microorganism antibiotic resistance through different mutations. Advanced oxidation processes have been proposed for the treatment of antibiotic in aqueous systems, including solar photocatalysis. Several parameters are necessary to take into account in solar photocatalysis treatment to eliminate antibiotics, since these compounds display different physicochemical and biological properties. This chapter discusses the effect of parameters and pathways (transformation products) of solar photocatalysis of antibiotic groups usually found in aquatic systems as macrolides, sulfones, lincosamides, and quinolone. M. Jiménez-Tototzintle · E. M. Saggioro (*) Sanitation and Environment Health Department, National School of Public Health, Oswaldo Cruz Foundation, Rio de Janeiro, RJ, Brazil e-mail: enrico.saggioro@ensp.fiocruz.br © Springer Nature Switzerland AG 2020 Inamuddin et al. (eds.), Nanophotocatalysis and Environmental Applications, Environmental Chemistry for a Sustainable World 30, https://doi.org/10.1007/978-3-030-12619-3_2

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Keywords Antibiotics · Antibiotic resistance bacteria · Aquatic systems · Advanced oxidation processes · Solar photocatalysis · Wastewater treatment plants

2.1

Introduction

In recent years, increasing attention has been paid to pollution of aquatic systems, which is affected by the continuous discharge of wastewater from of urban, industrial, hospital, etc. (Batt et al. 2006; Lam et al. 2004). Pharmaceutical industry is considered an emerging market, since sales have doubled in the last 5 years, leading to a market share of approximately 20% in developing countries (Koenig and Dillon 2017; Lindsley 2017; Tannoury and Attieh 2017). It is estimated that approximately half of the pharmaceutical wastewaters generated worldwide are discharge without correct treatment (Fleeger et al. 2003). Several factors have contributed to increased drug demands, including population growth, population aging and changing health, and disease perspectives (Dąbrowska et al. 2018). Accordingly, the World Health Organization has reported that high doses of antibiotics are consumed around the world (World Health Organization 2017). Antimicrobial compounds are often inadequately prescribed, and their misuse has had a negative effect on the environment, as a part of antibiotics is not completely consumed or degraded in human and animal bodies, and their residues enter aquatic ecosystems by wastewater treatment plants (Bielen et al. 2017; Hu et al. 2018). Antibiotics can be grouped either by their chemical structure or by mechanism of action and, depending on the chemical group, can be divided into different subgroups, such as ß-lactams, quinolones, tetracyclines, macrolides, and sulphonamides, among others (Bielen et al. 2017; Hu et al. 2018). Recent advances in analytical techniques allow for the sensitive multi-residue analysis of numerous antibiotics found in the environment in different matrices, such as water, soil, sediment, and sludge, as well as in the drinking waters of many countries (Rivera-Jaimes et al. 2018; Zhu et al. 2017). Antibiotic groups usually found in aquatic systems are macrolides, sulfones, lincosamides, and quinolone, in the order of ng.L
1. The summarized occurrence of these compounds is exhibited in Table 2.1. Detection of antibiotic pollutants is a complex step, since they display different physicochemical properties and can be present in several matrices, such as soil, water, food, and wastewater. Recent advances in analytical techniques for determining antibiotic residues in aquatic matrixes include gas chromatography and liquid chromatography coupled with mass spectrometry (Bottoni and Caroli 2015). In 2017, Mokh et al. investigated the occurrence of pharmaceuticals in surface waters sampled from Lebanese water systems. The analytical method was based on a solid-phase extraction followed by liquid chromatography with electrospray ionization coupled to tandem mass spectrometry in the positive ion mode, and the method was validated according to European Union guideline requirements. A critical point

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Table 2.1 Summarized studies of antibiotic occurrence in environment matrices worldwide Antibiotic classes Macrolides

Compounds Erythromycin

Sulfones

Sulfamethoxazole

Trimethoprim

Trimethoprim

Lincosamides

Lincomycin

Quinolone

Ciprofloxacin

Detected concentration (ng L
1) 346 3.0 70 21,278 120

Antibiotic source Raw wastewater Tap water

Sample location Spain

Surface water Effluent wastewater Rivers

UK

USA

Asia European

References Rosal et al. (2010) Zhu et al. (2017) Ebele et al. (2017) Tran et al. (2018) Johnson et al. (2015)

to identify and quantify 63 pharmaceuticals and some metabolites in the aqueous samples from the Lebanese water system, for example, rivers, wells, fountains, and lakes, was pretreatment and solid-phase extraction optimization, leading to increased rate of recovery, while at the same time decreasing the matrix effect. The results indicate that the use of Oasis cartridges allows for extraction recoveries of 54% of the evaluated compounds in the acceptable range (75–125%) (Mokh et al. 2017). On the other hand, the presence of antibiotics in aquatic systems has been linked to increasing microorganism antibiotic resistance through different mutations. As a result, potential burdens or ecological effects on the development of antimicrobial resistance are noted (Pruden et al. 2012). All species worldwide have either a direct or an indirect contact with antibiotic residues (Fig. 2.1). Thus, the spread of antibiotics in the environment puts our ability to treat common infectious diseases at risk (Alexander et al. 2016). Antibiotics and metabolites are introduced into the environment by different pathways but primarily from wastewater treatment plant discharges. However, conventional processes, such as biologic treatment, are not enough to biodegrade these substances, leading to persistence and entry into the environment when the treated wastewater is discharged into aquatic systems (Guo et al. 2017). Consequently, microorganisms come into direct contact with antibiotics and promote antibiotic resistance through different mechanisms (Lüddeke et al. 2015; Rodriguez-Mozaz et al. 2015; Sharma et al. 2016; Thayanukul et al. 2013; Zanotto et al. 2016). Antibiotic residues and metabolites are a risk to health and the environment, and, as such, it is necessary to propose efficient treatment methods to eliminate this source of contamination before it enters the environment (Pruden et al. 2012; Rizzo et al. 2013; Tahrani et al. 2015). In this scenario, advanced methods should be applied to deal with this environmental concern, including absorption by activated carbon, membranes, anaerobic digestion processes, bioremediation, or chemical advanced oxidation, as recently reviewed in the specialized literature (Bernabeu et al. 2011).

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Fig. 2.1 Route concerning the direct or indirect contact with antibiotic residues or antibiotic metabolites

Among advanced oxidation processes, solar photocatalysis deserves significant attention. In general, advanced oxidation processes can be used as a tertiary treatment or in increasing influent biodegradability, prior to its biological treatment (Jiménez-Tototzintle et al. 2015). The most important advanced oxidation processes characteristic is their in situ generation of free hydroxyl radicals. In this respect, solar photocatalysis has been proposed as an unconventional treatment for pathogenic microorganism inactivation and pollutant mineralization (Chun Hu et al. 2007; Martínez-Costa et al. 2017; Moosavi and Tavakoli 2016).

2.2

Solar Photocatalysis Process

In general, photocatalysis processes are based on the direct or indirect absorption of irradiant energy (visible or ultraviolet light) by a semiconductor, which is excited with higher energy. Thus, the bandgap bands an energy, and their electrons exit from the valence band to the conduction band, promoting positive holes in the valence band, i.e., electron-hole pair (e
–h+) generation (Fig. 2.2) (Nakata et al. 2012; Ohtani 2010). During the process, electron-hole pair can recombine on the surface or in the matrix of the particle in a few seconds (and the energy is then dissipated as heat) or can be trapped in surface states, where they can react with adsorbed donor or acceptor species (Fig. 2.2) (Litter 1999).

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Fig. 2.2 Photoabsorption by electron transition in a semiconductor valence band to the conduction band. (Carp et al. 2004)

The semiconductor is usually solid, such as oxides or chalcogenides, as they display adequate bandgap energies to be excited by ultraviolet or visible light and the redox potentials of the edges of the valence band and conduction band can generate a series of oxidative or reductive reactions (Litter 1999). However, the most applied semiconductor in the photocatalysis processes is TiO2. The benefits of using TiO2 photocatalysis for eliminating pollutants from water include its high photocatalytic efficiency, physical and chemical stability, and low cost and low toxicity. When TiO2 is irradiated with wavelength lower than 390 nm light, electrons migrate from the valence band to the conduction band, leaving a positively charged hole in the valence band and creating an electron-hole pair (Eq. 2.1) (Calza et al. 2003). TiO2

þ

hν ðλ < 390 nmÞ ! e


þ

hþ

ð2:1Þ

The different possible catalytic reactions in TiO2 photocatalysis are complex and include the water, the dissolved oxygen, and the catalyst groups in the media, summarized by Eqs. (2.2) and (2.3) (Fernández et al. 2005). • þ H 2 O þ hþ VB ! OH þ H

ð2:2Þ

•
O 2 þ e
CB ! O2

ð2:3Þ

As mentioned, the light intensity determines the extent of light absorption by the semiconductor catalyst at a given wavelength. In this context, the strategy to use solar light as a source of natural energy is a constant challenge (Byrne et al. 2017). Many authors have researched catalysis synthesis to promote the solar photodegradation. Also, the researchers are focused to design of solar photocatalytic reactors in order to obtain greater benefits in the treatment of pollutants in aquatic systems through solar photocatalysis (Ahmadi et al. 2016; Ani et al. 2005; Ayala et al. 2014; Binas et al. 2017; Binh et al. 2018; García-Fernández et al. 2015).

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2.3

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Solar Photocatalysis Treatment for Antibiotic Degradation

Solar photocatalysis treatment has been proposed for the treatment of antibiotic groups usually found in aquatic systems. Many parameters are necessary to take into account in solar photocatalysis treatment to eliminate antibiotics, since these compounds display different physicochemical and biological properties, such as, log Kow, sorption behavior, photoreactivity, antibiotic activity, and toxicity group. The next section exemplifies solar photocatalytic treatments for antibiotic degradation, frequently detected in various surface waters or/and different aquatic matrices.

2.3.1

Trimethoprim

Trimethoprim is a synthetic antibiotic, comprising a 2-aminopyrimidine scaffold (Fig. 2.3), usually prescribed in combination with sulfamethoxazole for the treatment of respiratory, urinal, and gastrointestinal infections. Up to 10 to 20% of trimethoprim is processed, primarily in the liver, while the rest is excreted, unchanged, in urine (Abellán et al. 2009). Trimethoprim is highly water-soluble, so that it is commonly found in aquatic resources. In addition, trimethoprim is not eliminated by conventional processes, such as biological treatment (Batt et al. 2006). Solar photocatalysis treatment has been considered the main trimethoprim transformation pathway in the environment by different authors (Dedola et al. 1999; Luo et al. 2012; Martínez-Costa et al. 2017; Michael et al. 2012; Prieto-Rodriguez et al. 2012; Ryan et al. 2011; Sirtori et al. 2010). Trimethoprim solar photocatalysis treatment has been evaluated under different conditions in order to study mechanism degradation and photoproduct yield. Abellán et al. (2009) studied the effect of TiO2 suspension concentrations on the photocatalytic process of 100 mg L
1 of trimethoprim in deionized water using a solar simulator and tubular quartz reactor with a Xenon lamp (1500 W). The results indicate that the photocatalysis efficiency is directly proportional to the amount of catalyst, where 0.2 gL
1 of TiO2 led to over 95% trimethoprim degradation, and values above the optimal concentration decreased photocatalytic degradation, due to the scattering phenomena between the catalyst particles, since trimethoprim can absorb radiation up to a wavelength of 310 nm. Fig. 2.3 Trimethoprim structure

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Irradiation effects were evaluated, and the results demonstrated that trimethoprim could be degraded by photolysis in 300 min. However, photocatalysis was more efficient in trimethoprim mineralization than photolysis, where trimethoprim mineralization percentages reached 35% using 0.2 g.L
1 of TiO2 and only 7% in photolysis. Thus, irradiation can degrade the parent compound to some extent, but mineralization of the parent compound is higher when the catalyst is light activated. Some oxidation products reported during photocatalysis processes were diastereoisomeric tetraarylethanes, compounds more aromatic than the parental antibiotic (Abellán et al. 2009). The development of solar pilot plants is an important achievement in solar photocatalysis, increasing photon absorption efficiency and enabling the treatment of large amounts of water. Many authors have described pilot compound parabolic collector plants designed for solar photocatalytic applications (Carbajo et al. 2016; García-Fernández et al. 2015; Jiménez-Tototzintle et al. 2015; Malato et al. 2009; Rizzo et al. 2013; Saggioro et al. 2014). The photoreactor is composed of Pyrex glass tube modules mounted on a fixed platform tilted 37 (local latitude), where the solar radiation is reflected and concentrated in the absorber tube. The intensity of solar ultraviolet radiation absorbed by the photoreactor is measured by a radiometer, as ultraviolet radiation is an important parameter for result evaluation during solar photocatalytic treatment in a pilot plant. Consequently, the irradiation time is often normalized. Malato et al. (2003) described Eq. (2.4) for the combination of data from several days of experiments and their comparison with other photocatalytic experiments. t 30W , n ¼ t 30W ,n
1 þ Δt n

UV V i ; Δt n ¼ t n
t n
1 30 V T

ð2:4Þ

where VT is the total volume, (Vi) is the irradiated volume, tn is the experimental time for each sample, UV is the average solar ultraviolet radiation (wavelength lower than 400 nm) measured during the interval tn, t30W is a “normalized illumination time,” and 30 Wm
2 refers to a constant solar ultraviolet power (typical solar ultraviolet power on a perfectly sunny day around noon) (Malato et al. 2003). Sirtori et al. (2010) used the compound parabolic collector pilot plant to evaluate the degradation of 20 mg L
1 trimethoprim using 200 mg L
1 of catalyst in two different water matrices: demineralized water and simulated seawater, where the pH was adjusted to 6.0 and 7.5 for demineralized water and seawater, respectively. The results indicate that degradation efficiencies were affected by the presence of inorganic species, which can act as hydroxyl radical scavengers. In this sense, the Cl
species contained in the seawater matrix reacted with hydroxyl radicals and generated less reactive species than the hydroxyl radicals (Eqs. 2.5, 2.6 and 2.7): Cl
þ HO ! ClOH 

ð2:5Þ

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M. Jiménez-Tototzintle and E. M. Saggioro

ClOH  þ H þ ! H 2 O

ð2:6Þ

Cl þ Cl
! Cl
2

ð2:7Þ

However, trimethoprim was completely degraded in the different water matrices, following apparent first-order kinetics. The percentage of mineralization for trimethoprim in the different matrices was of 65% in 107 min for demineralized water and 400 min in simulated seawater. Trimethoprim transformation products during solar photocatalysis treatment were identified by liquid chromatography-time-of-flight mass spectrometry. The generation of inorganic nitrogen compounds, such as ammonia and nitrate, and propionic and oxalic acid during trimethoprim degradation was detected in demineralized water. Similar intermediates were identified during trimethoprim photocatalysis in both matrices (demineralized water and seawater), where trimethoxybenzylpyrimidine (m/z 305; C14H17N4O4, protonated molecule) was identified during the first irradiation times, and from that point on, a series of photoreactions involving hydroxyl radicals were generated. These reactions included hydroxylation, demethylation, and, finally, cleavage of the trimethoprim molecule. However, most of the hydroxylated derivatives detected in demineralized water were not present in seawater. Incomplete mineralization in seawater could be due to the presence of ions in the matrix that inhibit reaction with hydroxyl radicals (Sirtori et al. 2010).

2.3.2

Sulfamethoxazole

Sulfonamides are among the most frequently detected antibiotics in soils and water, due to their physicochemical properties (Chen et al. 2017; Kang et al. 2018; Park and Huwe 2016). Sulfonamides are weak acids and highly polar compounds, and a noteworthy compound belonging to this class is sulfamethoxazole (Zhang et al. 2017). Sulfamethoxazole is a bacteriostatic antibacterial agent that interferes with folic acid synthesis in susceptible bacteria and is often used to treat urinary tract infections (Fig. 2.4). Sulfamethoxazole exhibits high-water solubility (610 mg.L
1 at 37  C) and low KOW ¼ 0.89, meaning that the compound can leach into both groundwater and surface. Moreover, sulfamethoxazole has acid-base properties in aqueous solution, as it displays two protonated states (pKa1 ¼ 1.7 and pKa2 ¼ 5.6). This

Fig. 2.4 Sulfamethoxazole structure

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43

parameter is important to consider when evaluating photocatalysis degradation (Chen et al. 2017). Abellán et al. (2007) evaluated the influence of different variables, such as pH and titanium concentrations, in the photocatalysis degradation of sulfamethoxazole and total organic carbon reduction rates. The photocatalytic reaction was carried out in a tubular quartz reactor, which was placed in the axis of two parabolic mirrors in a solar simulator. The presence of dissolved oxygen was measured during the photocatalytic reaction, and result indicates that the initial oxygen concentration (8 mg.L
1) decreased during the first hour of reaction and then maintained constant (6 mg.L
1). The presence of oxygen is necessary for heterogeneous photocatalysis reactions, as it reacts with the generated electrons on the surface of the catalyst, reducing electron-hole recombination. Different catalyst concentrations were evaluated, and the results demonstrate that 0.2 g.L
1 of TiO2 eliminated 91% of the initial sulfamethoxazole concentration, although 0.5 g.L
1 of TiO2 led to the highest mineralization percentage. Colored intermediates were observed during the photocatalytic degradation of sulfamethoxazole. These intermediates compete with the catalyst for the available radiation, and as a result, photocatalytic efficiencies were affected. Four experiments at pH 2, 5, 7, and 11 were carried out to evaluate pH effects during photocatalytic degradation. The result indicates that mineral acids appear throughout the experiment due to gradual solution mineralization, causing a pH drop. Total organic carbon elimination rate was affected at pH 2, but when the pH of the solution increased from 5 to 11, the rate remained around 30%. The chromatograms registered during pH evaluation indicate that the intermediates were quite similar at pH 5, 7, and 11. However, at pH 2 the formed intermediates displayed different residence time and area, indicating that the sulfamethoxazole photocatalytic degradation pathway is affected by pH. This behavior could be caused by adsorption of the ionic form of sulfamethoxazole on the catalysis surface, which is affected by the solution pH. TiO2 has a point of zero charge between 5.6 and 6.4, so the TiO2 charge depends on the solution pH, with positive charges at pH < 6 and negative at pH > 6. The liquid chromatography mass spectrometry study indicates that the hydroxyl radicals first attack the isoxazole ring and, subsequently, the aniline ring (Abellán et al. 2007). An essential step to develop photocatalytic reactions is to evaluate the catalyst concentration, in order to reduce catalyst loading and increase photon absorption. The catalyst concentration depends on many factors, such which ions are present, total organic carbon, nature of the pollutant, and turbidity of the aqueous matrix. Prieto-Rodriguez et al. (2012) evaluated the photocatalysis operation at low TiO2 concentrations for the treatment of different emerging contaminants identified in secondary biological wastewater treatment plants effluents and detected sulfamethoxazole, among others. Solar photoreactors with different diameters and optical thicknesses were used to increase photon absorption efficiency. The tubular reactor consisted in a Pyrex glass tube (170 mm long, 32 mm optical density, Øinner ¼ 29.2 mm, glass wall thickness

44

M. Jiménez-Tototzintle and E. M. Saggioro

1.4 mm, light transmission wavelength less than 400 nm 91%). The wastewater effluents were pretreated with an inorganic acid to remove carbonates, which can act as a hydroxyl radical scavenger, and the pollutants were spiked into the wastewater treatment plants. The result demonstrates that the reactor was 20.8% efficient regarding adsorbed radiation when the photocatalyst concentration was maintained at 20 mg.L
1 of TiO2, while efficiency at 50 mg.L
1 was of 40.6%. The sulfamethoxazole degradation rate was doubled when the photocatalyst concentration varied from 20 to 50 mg. L
1, corresponding to double the rate of photon absorption when the photocatalyst concentration was increased from 20 to 50 mg.L
1. An experiment performed by modifying the radiant fluxes demonstrated no significant differences in photocatalytic sulfamethoxazole degradation when the photocatalyst concentration is low. Another series of experiments were carried out with the aim of comparing photocatalytic efficiency when the diameter of the reactor tube was increased. The result indicated that the larger diameter photoreactor (Øinner ¼ 46.4 mm) was more efficient than smaller diameter photoreactor (Øinner ¼ 29.2 mm), with radiation absorption efficiencies of 30.3 and 55.8%, respectively. In addition, the compound parabolic collector photoreactor was more efficient than the laboratory-scale reactor. The compound parabolic collector exhibited, on average, 70% higher photon absorption efficiency than a tubular reactor, requiring 39% less catalyst to operate under optimal conditions. In the larger diameter photoreactor, 90% of total pollutants were eliminated after 300 min of irradiation, while the smaller diameter photoreactor eliminated 85% of total pollutants after 480 min of irradiation. Thus, compound parabolic collector with a larger tube diameter increases solar radiation absorption and, consequently, leads to an increase in the solar photocatalytic degradation of pollutants present in the water sample (Prieto-Rodriguez et al. 2012).

2.3.3

Erythromycin

Erythromycin is a glycosidic 14-membered ring macrolide antibiotic widely used for the treatment of several bacterial infections. In addition, erythromycin is an alternative to patients presenting allergy to penicillin (Vignesh et al. 2014). It is also used in cattle raising, to promote animal growth and prevent bacterial infections (Gao et al. 2016). The extensive use of erythromycin, the third most frequently applied antibiotic, leads to the common detection of this compound in environmental matrices like municipal and industrial wastewater (Luiz et al. 2010). Commercial photocatalysts, such as Degussa P25, Millennium PC, Aldrich, and Tronox, can be applied in solar photocatalysis to remove antibiotics and antimicrobial activity. Xekoukoulotakis et al. (2010) investigated several operational parameters for different commercially manufactured TiO2. The Degussa P25 catalyst resulted in high erythromycin degradation, with the mixture of the crystalline phase (anatase/rutile) providing more activity than individual phases, such as

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45

Hombikat UV100 – anatase higher than 99%, as a phase mixture avoids the electro/ hole recombination. The electron recombination of the conduction band to the valence band is limited in improving the efficiency of solar photocatalysis. Thus, solar photocatalysis requires doping methods with specific elements to improve the solar spectrum. To avoid the electron/hole recombination, Karaolia et al. (2018) performed the synthesis of graphene-based TiO2 composite photocatalysts and achieved 84% of erythromycin removal after 60 min. The strong interaction between TiO2 and graphene promotes the charge connection, making the hole/electron recombination difficult (Karaolia et al. 2018). This recombination was also prevented by Vignesh et al., (2014), who synthesized zinc phthalocyanine modified TiO2 nanoparticles. Zinc phthalocyanine is easily photoexcited on the TiO2 surface, promoting mobile electrons in the conduction band, which react to surface adsorbed oxygen (Vignesh et al. 2014). These strategies provide the highest photocatalytic removal efficiency, explained by the reduction of the electron-hole recombination rate, increased specific surface area, and decreased particle size of the catalyst (Fakhri et al. 2016). In this way, Fakhri et al. (2016) promoted the synthesis of a γ-Fe2O3/SiO2 nanocomposite and evaluated optimal operational parameters. The sol-gel doping method promoted a morphological spherical shape and a blue shift in the short wavelength of the catalyst. Consequently, high erythromycin removal was achieved (87.17%) at 500 mg.L
1 of catalyst, 6 mg.L
1 of initial erythromycin concentration, and 6 min of irradiation time under basic conditions (pH ¼ 8). The pH effect is complex, since it depends on the equilibrium of water dissociation, the surface charge of the catalyst regarding its point of zero charges and the ionization of erythromycin molecules (pKa ¼ 8.8) and its transformation products (Xekoukoulotakis et al. 2010). Erythromycin transformation pathways were studied by Gao et al., (2016), who promoted erythromycin photocatalysis using ZnIn2S4 as catalyst. The complex erythromycin structure, comprising a 14-membered lactone ring and cladinose and desosamine deoxy sugars, is presented in Fig. 2.5 (Luiz et al. 2010). Fig. 2.5 Erythromycin structure

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M. Jiménez-Tototzintle and E. M. Saggioro

Erythromycin degradation begins with sugar molecule cleavage (desosamine and cladinose). However, these sugars are found at the end of photocatalysis processes, indicating that these intermediates do not undergo oxidation. Subsequently, degradation promotes demethylation of the tertiary amine in the secondary amine. Finally, via radical ion attack, deprotonated intermediates are detected, due to the open lactone ring (Gao et al. 2016). These intermediates formed during photocatalysis probably do not present antimicrobial activity. According to Xekoukoulotakis et al., (2010), after 10 min of photocatalysis at a residual erythromycin concentration of 10 mg.L
1 and only 10% mineralization, antimicrobial activity was eliminated.

2.3.4

Ciprofloxacin

Ciprofloxacin (Fig. 2.6) is a second-generation of antibiotic quinolone that represents most of the total antibiotic intake worldwide (Adriaenssens et al. 2011). This antibiotic class acts through the inhibition of DNA replication in bacteria via breakdown of DNA topoisomerase I and IV (Von Döhren 2009). Most quinolones are excreted in unaltered form or as a pharmaceutically active metabolic and released into the environment through wastewater effluents and can be very persistent in the aquatic environment (Hassani et al. 2015). Ciprofloxacin has been detected in different environment compartments, such as hospital (Hartmann et al. 1998) and municipal treatment plant (Le-Minh et al. 2010) wastewater, surface water (Watkinson et al. 2009), groundwater (Hu et al. 2010), and drink water (Cheng et al. 2015). The pH strongly influences ciprofloxacin removal rates, since the differences in pH determine ciprofloxacin speciation in the aqueous medium (Salma et al. 2016). The effect of pH on ciprofloxacin photodegradation is complex, since ciprofloxacin has two pKa values (5.9 and 8.89) and, under different pH, occurs as a cation (CIP0,+), zwitterion (CIP
,+) and anion (CIP
,0) (Hassani et al. 2015). Salma et al. (2016) found the highest removal rates at pH 9. An et al. (2010) found similar results, where the rate constant of ciprofloxacin degradation at pH 3.0 was 0.06 min
1, while at pH 9.0 the rate constant was 0.38 min
1. Defluorination is the first step in ciprofloxacin degradation (Silva et al. 2016). In acidic conditions, the fluorine atom is preserved, while transformation products lose this ion in basic media. Furthermore, decreased efficiency of photodegradation at low pH is attributed to the scavenger Cl
ion effect and hydrogen ion adsorption on the catalyst surface that compete with ciprofloxacin for active sites (Hassani et al. Fig. 2.6 Ciprofloxacin structure

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2015; Wang et al. 2018). The level energy of the triplet state of ciprofloxacin in basic media undergoes a photo-induced chemical reaction, promoting the fluorine reaction (Salma et al. 2016). On the other hand, Hassani et al. (2015) found high ciprofloxacin photodegradation at pH 5 applying synthesized TiO2 nanoparticles on montmorillonite. The determined pHzpc was 8.4, so the absorption between ciprofloxacin and the catalyst was maximum, as the TiO2 nanoparticles on montmorillonite surface is in its positive form and ciprofloxacin is in its neutral form. Silva et al. (2016) studied ciprofloxacin photodegradation by TiO2, where ciprofloxacin was highly absorbed onto the catalyst surface in the first 15 min. After 45 min of photodegradation, ciprofloxacin was not detected, but toxicity was increased, with maximum luminescence inhibition in Vibrio fischeri. The ciprofloxacin structure breakdown pathway can occur by direct photo-hole attack and by the addition of hydroxyl radicals (An et al. 2010). The by-products formed from the ciprofloxacin reaction are due to the breakdown of the piperazine ring and the quinolone moieties. Ciprofloxacin photodegradation begins with defluorination by superoxide radical attack on carbon 12 and hydroxyl radical substitution followed by piperazine ring cleavage (Silva et al. 2016; Wang et al. 2018). The reasonable addition of hydroxyl radicals can be at positions of carbon 8, 5, and 2, respectively, in the quinolone ring (Haddad and Kümmerer 2014). Furthermore, superoxide radicals can attack the quinolone group (peroxide formation), with subsequent cleavage of the benzene and generation of the ketone products from benzene (Wang et al. 2018). Salma et al. (2016) attributed the hydroxyl radical attack to the secondary aliphatic amine group of the piperazine ring. The transformation of the piperazine ring can occur for either H2O or an
OH group insertion (Silva et al. 2016) or oxygen addition with the formation of a peroxy-piperazine ring (Salma et al. 2016). Applying the same approach, Durán-Álvarez et al. (2016) proposed that the ciprofloxacin photocatalytic transformation occurs mainly via the piperazine ring cleavage instead of quinolone moiety cleavage. Consequently, a loss of antibacterial activity was observed during the photocatalysis process, due to the cleavage of the piperazine ring. The piperazine ring is easily attacked at the nitrogen1 and 4 atoms by positive hole (Wang et al. 2018). Piperazine transformation of the carbon7 substituent affects the acid-base speciation of the ciprofloxacin molecule, which alters cell permeation and reduces the recognition and affinity to DNA topoisomerases (Paul et al. 2010). Moreover, defluorination is related to the decrease of antibacterial activity, since the fluorine substituent increases ciprofloxacin cell penetration and DNA binding (Paul et al. 2010).

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M. Jiménez-Tototzintle and E. M. Saggioro

Conclusions

The presence of antibiotic residues and metabolites in environmental matrices is an increasing concern worldwide. Pharmaceutical industries exhibit an increase in antibiotic percentage sales, as these compounds are often inadequately prescribed and used indiscriminately. This profile favors antibiotic resistance and is a risk to health and to the environment. Solar photocatalysis is a powerful technology for antibiotic degradation and decreases the antibacterial activity of trimethoprim, sulfamethoxazole, erythromycin, and ciprofloxacin, the main antibiotic compounds found in aqueous environment matrices. However, the electron recombination of the conduction band to the valence band is the main drawback to improving the efficiency of solar photocatalysis. Trimethoprim can be degraded during solar photolysis. The pH strongly influences sulfamethoxazole and ciprofloxacin photocatalysis processes, since these compounds exhibit two pKa values. Sulfamethoxazole is primarily degraded to hydroxyl radicals that cleave the isoxazole and aniline rings. Ciprofloxacin photocatalysis occurs via positive holes and hydroxyl radicals, which lead to breakdown of the piperazine ring and the quinolone moieties. Primary, erythromycin oxidation is due to sugar cleavage, such as desosamine and cladinose. Acknowledgments Saggioro, E.M. would like to thank FAPERJ for financial support (E-26/ 010.002117/2015 and E-26/203.165/2017).

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Chapter 3

Biomass-Based Photocatalysts for Environmental Applications Yean Ling Pang, Chin Woei Lim, Katrina Pui Yee Shak, Steven Lim, Wai Chong Cheam, Chai Hoon Koo, and Ahmad Zuhairi Abdullah

Contents 3.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2 Background of Biomass-Derived Carbon . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2.1 Biochar . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2.2 Activated Carbon (AC) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.3 Synthesis Methods of Biomass-Derived Carbon . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.3.1 Pyrolysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.3.2 Hydrothermal Carbonization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.3.3 Physical and Chemical Activation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.4 Photocatalysts and Photocatalysis Reactions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.5 Functionalized AC and Applications . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.5.1 Types of Functionalized AC . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.5.2 Functionalized AC Photocatalysts and Its Application . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.6 Future Challenges and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

56 58 58 60 60 60 62 64 66 69 69 72 79 80

Abstract In recent years, advanced oxidation process has shown a new frontier environmental friendly and sustainable wastewater treatment technology to replace the conventional treatment methods. Heterogeneous photocatalysis is a promising approach in environmental remediation. Among the popular photocatalysts, semiconductor oxide photocatalysts have been widely explored due to their durability and chemical and physical properties. However, the practical application of semiconductor oxide is limited because of its expensive cost, large bandgap energy and Y. L. Pang (*) · C. W. Lim · K. P. Y. Shak · S. Lim · W. C. Cheam Department of Chemical Engineering, Lee Kong Chian Faculty of Engineering and Science, Universiti Tunku Abdul Rahman, Kajang, Selangor, Malaysia e-mail: [email protected] C. H. Koo Department of Civil Engineering, Lee Kong Chian Faculty of Engineering and Science, Universiti Tunku Abdul Rahman, Kajang, Selangor, Malaysia A. Z. Abdullah School of Chemical Engineering, Universiti Sains Malaysia, Nibong Tebal, Penang, Malaysia © Springer Nature Switzerland AG 2020 Inamuddin et al. (eds.), Nanophotocatalysis and Environmental Applications, Environmental Chemistry for a Sustainable World 30, https://doi.org/10.1007/978-3-030-12619-3_3

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rapid recombination of the photoinduced electron-hole pairs. This chapter addresses the main advancements in overcoming the barriers accompanied by pure semiconductor oxide by incorporating with biomass-derived carbonaceous materials for fabrication of effective visible light-responsive semiconductor oxide-based photocatalysts. The utilization of these biomass-derived photocatalysts provides a greener synthesis route for environmental purifications. This chapter firstly highlights the types, properties and conversion of biomass into biochar, activated carbon (AC) or any other carbonaceous materials. The general methods for preparing biomass-based support and the mechanisms are also presented in details. Finally, the latest research papers on modification or functionalization of a biomass-based photocatalyst for the removal of organic pollutants have also been comprehensively discussed. There is still a lack of published information on the fundamental knowledge of biomassderived photocatalyst to understand the complex photocatalytic mechanisms especially on the reactions occurring on the heterogeneous biomass surfaces. Therefore, further theories and hypotheses need to be investigated to fill these gaps in order to address all full potentials of the biomass-derived photocatalyst for environmental applications. Keywords Biomass · Activated carbon (AC) · Biochar · Photocatalyst · Modification · Functionalization · Heterogeneous · Advanced oxidation process · Photocatalytic degradation · Organic pollutants

3.1

Introduction

Clean water is one of the most important elements to support all the living organisms. The surface of the Earth is covered by about 70% of water. However, it has been reported that only 2.5% is available for drinking, agriculture and domestic and industrial consumption (Malathi et al. 2018). The rapid industrial development exacerbates environmental pollution, global warming and climate changes. This has polluted the environment by releasing large amount of hazardous wastes and toxic organic pollutants to the wastewater stream. These are very serious problems currently facing in the whole world. Various physical, biological and chemical methods have been applied to treat these contaminants. Therefore, water and wastewater treatment technologies are important for human health and sustain the global environment. Heterogeneous photocatalysis appeared as an effective, efficient and environmentally friendly technology to degrade various types of recalcitrant organic pollutants. In the past two decades, photocatalytic degradation of organic pollutants in the presence of semiconductor oxide such as titanium dioxide (TiO2), zinc oxide (ZnO), cadmium sulphide (CdS), zinc sulphide (ZnS), iron oxide (Fe2O3), nickel oxide (NiO), tin oxide (SnO2) and carbon nanotubes has attracted increasing attention. This is due to their low-cost, low toxicity and high physical and chemical stability to

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degrade and mineralize organic pollutants into carbon dioxide (CO2), water (H2O) and harmless inorganic anions (Szczepanik 2017). Unfortunately, most of the semiconductors possess wide energy bandgap (Eg) such as TiO2 (3.2 eV) (Szczepanik 2017) and ZnO (3.37 eV) (Qi et al. 2017). The catalytic efficiency of these semiconductors is generally determined by the sufficient light absorption which is able to excite electrons (e
) in the valence band to the conduction band, leaving holes (h+) in the valence band to induce the photoredox reactions. This implies the necessity of employing stronger UV light which only comprises about 3–5% from the natural solar light (Uyguner-Demirel et al. 2017). Moreover, the fast recombination of photo-generated electron-hole pairs also minimizes the performance of these common photocatalysts. Therefore, several possible modification techniques have been investigated to extend their visible light response and restrain the recombination of electron-hole pairs through doping with metals/nonmetals/noble metals and constructing heterojunctions using coupling carbon materials. Carbon-based nanostructured materials offer unique advantages such as chemical inertness, high stability, porosity and tunable structures (Khalid et al. 2017). These materials usually possess large specific surface area and high mobility of charge carriers that are feasible to act as support for metal oxide particles. The common carbon materials include carbon nanotubes, carbon doping, diamond, graphite, graphene and C60 fullerenes. Several reviews on the development of carbon nanostructures and photocatalytic applications had been published previously (Khalid et al. 2017; Xiang et al. 2012; Jayaraman et al. 2018). Recently, the utilization of biomass waste as carbonaceous materials for producing biochars or activated carbon (AC) is gaining interest because it solves the waste disposal problem and reduces the cost of raw materials and the properties of the final products can be tailored for different applications. Thus, a lot of attention has been focused on the application of these biomass resources for biochar production via pyrolysis under oxygen-limited conditions and solvothermal/hydrothermal and physical/chemical activation methods. The product biochar usually exhibits porous structure, multi-surface functional groups and mineral components due to the removal of moisture and volatile matter contents during the thermochemical process (Tan et al. 2017). These favourable properties are possible to be used as alternative carbon materials. The average price for biochar was highly dependent on the origin of biochar production sites, and its prices were ranging between $0.09 kg
1 (Philippines) and $8.85 kg
1 (UK) (Tan et al. 2017). Therefore, it is of great importance to understand the development of this type of activated biochar as a new potential cost-effective and environmentally friendly carbon material in order to optimize and scale up the process conditions. This book chapter initially describes the background of biomass followed by the synthesis methods to produce activated biochar. Meanwhile, the basic principles of photocatalysis reaction are presented. The performances of activated biochar applied in various fields have also been reported. This is then followed by the discussion on the current progress and approaches for the enhancement of semiconductor photocatalytic activity by biomass-based carbonaceous nanomaterials. Finally, latest

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development on the literature and future challenges are summarized in the concluding remarks. It is anticipated that this chapter will provide new insights into the design of biomass-based carbonaceous-semiconductor composites for environmental applications.

3.2

Background of Biomass-Derived Carbon

In general, biomass can be explained as the biodegradable fraction of products, waste and residues from various origins such as agriculture. Biomass has a complicated combination of carbon (C), oxygen (O), sulphur (S), nitrogen (N), hydrogen (H) and a minor portion of other substances such as alkali and heavy metals. The carbon element usually has the highest portion within biomass, followed by oxygen and hydrogen elements (Tripathi et al. 2016). The constituent will be different depending on the feedstock and reaction condition. Broadly, biomass is ample in nature and can be grouped under natural or anthropogenic based on their origin. The former is the biomass found in nature, whereas the latter can be obtained via miscellaneous processing technologies on natural biomass. Specifically, biomass can be further categorized into five small groups based on the location of the source, including woody, aquatic, human and animal, industrial as well as agricultural biomass. There are several types of biomass (Urbaniec and Bakker 2015). Woody biomass covers the various parts of trees such as stem, branch, leaves and lumps. Several examples of woody biomass are oaks, pines, redwoods and maples which are mainly from the forest area. Meanwhile, aquatic biomass covers the species of microalgae, plants and microbes present in water. For example, blue algae, fungus and variant water seeds. Animal and human biomass comprises animal manure, cooked or uncooked foods and fruits. On the other hand, industrial biomass includes the waste generated from different industries such as paper sludge from the paper industry, food waste from food processing industry as well as sugar cane residue from sugar mill industry. Industrial biomass is distinguished from animal and human biomass due to the containment of various toxic chemicals and additives in industrial biomass. Last but not least, agricultural waste consists of the waste generated as the result of the agricultural operation. This type of biomass can be obtained from coconut, cocoa, coffee, sugar cane and others.

3.2.1

Biochar

Biochar is a carbon-rich product of biomass generated from thermal decomposition under limited oxygen (O2) supply at a relatively low temperature (less than 700  C) (Yu et al. 2017). The composition of biochar varies with the types of feedstock. As a rule of thumb, increasing the pyrolysis temperature of biochar will increase carbon

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Fig. 3.1 A porous biochar containing several types of functional groups. Numerous surface functionalities such as aromatic, hydroxyl (–OH), –COOH, C–O and C¼O exist on biochar. Reprinted from (Lee et al. 2017) with permission of (Elsevier)

content, aromaticity, porosity, ash content, pH, surface area and microporosity while at the same time decrease hydrogen content, oxygen content and polarity (Kah et al. 2017). Biochar can be distinguished among the carbonaceous compounds such as AC and carbon nanotubes as it consists of various functional groups including OH-, carboxyl (-COOH) and phenolic group. Biochar possesses numerous surface functionalities such as carbon single bond (C-O) and carbon double bond (C¼O) to oxygen. Figure 3.1 depicts several common functional groups attached to a porous biochar. However, it is difficult to control the desired type and content of the surface functionality through heat treatment process. Thus, a post-synthetic surface functionalization/doping is required to introduce specific functionality for different applications (Cheng et al. 2017) such as surface oxidation, surface amination and surface sulfonation. Biochar’s pore size varies among nanopores (< 0.9 nm), micropores (50 nm). The formation of the pore is a result of liberating the volatiles from the backbone of carbonaceous feedstock (You et al. 2017). Different pore sizes are important for different applications. For example, large pore size is good to promote soil breathability and raise the number of water traps in the soil to maintain the spacing for microbial growth and reproduction (Tan et al. 2017). Biochar can also be employed as adsorbent/catalyst due to its microporous structure, high carbon content and specific surface area. Moreover, biochar offers a low-cost alternative ($246/ton un-activated biochar) as compared to a traditional adsorbent such as granular ACs ($1500/ton), thereby reducing the water treatment costs (Lin et al. 2017).

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Activated Carbon (AC)

AC is charcoal that has been treated (i.e. activated) with oxygen (in general) to increase its microporosity and surface area. The term “activated” is referred to the enhanced surface area of charcoal upon thermal or chemical treatment. The characteristics of AC include high pore volume, well-established internal porous structure and numerous functional groups (Tran et al. 2017). AC is actually similar to biochar in terms of preparation method, source materials and high surface area properties (Ahmad et al. 2014). The distinctions between them are obvious. The production temperature of biochar is usually less than 700  C, which is lower than that of AC. Besides, the activation process is unnecessary for biochar production, while it is crucial for AC production. Thus, the production price of biochar is usually about 1/6 of that of commercial AC that is around the US $246/ton (Zhang et al. 2017). It has been reported that AC usually possesses specific surface areas in the range of 500–3000 m2/g, while biochar possesses variable specific surface areas from 2 to 750 m2/g (Kah et al. 2017).

3.3

Synthesis Methods of Biomass-Derived Carbon

Generally, there are three common techniques employed to produce biomass-derived carbon materials, namely, pyrolysis and solvothermal/hydrothermal and physical/ chemical activation methods. These methods are utilized by many types of research to produce high quality and yield of biomass-derived carbon materials.

3.3.1

Pyrolysis

Pyrolysis is a thermal decomposition process of organic (carbon-based) materials at high temperature in the absence of oxygen or inert atmosphere (Dhyani and Bhaskar 2017; Wang et al. 2017b). Pyrolysis of biomass usually produces char (solid residue), bio-oil (liquid) and syngas (vapour) in different proportions, and its process depends on the operating parameters such as temperature or heating rate. The oxygen-free environment can prevent the materials from complete combustion to carbon dioxide. Consequently, biomass is forced to break down into its smaller constituents and generates more stable products instead of combusting into water and carbon dioxide. According to Tripathi, Sahu and Ganesan (2016), pyrolysis usually proceeds as a two-stage process: primary and secondary pyrolysis. In primary pyrolysis, biomass is broken down and devolatilized into its main constituents through dehydration, decarboxylation and dehydrogenation of the biomass. At the initial temperature less than 100  C, moisture will be vaporized. Upon reaching 220–315  C, hemicellulose degenerates easily. When the temperature is further

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increased up to 315–400  C, cellulose is pyrolysed, while lignin decomposition covers a wider range of temperature (Deng et al. 2016). Upon the completion of primary pyrolysis, the secondary pyrolysis occurs by cracking large organic molecules into smaller ones to produce biochar and vapour products such as methylene, methane, carbon monoxide and carbon dioxide. Therefore, the major component of the residual solid is carbon. In general, six different types of pyrolysis have been developed: fast, catalytic fast/flash, intermediate, slow, vacuum and hydropyrolysis (Tripathi et al. 2016). Slow pyrolysis is characterized by a temperature between 400 and 500  C with a slow heating rate of about 0.1 to 1  C/s for 5–30 min. Meanwhile, fast pyrolysis involves a higher temperature of 850–1250  C with a heating rate of 10–200  C for 1–10 s. Fast pyrolysis favours bio-oil production, while slow pyrolysis favours the biochar production. Liquid product instead of biochar is formed due to the insufficient time of the biomass to react and convert into biochar in fast pyrolysis process. The modified fast pyrolysis can be considered as flash pyrolysis. The temperature involved in this flash pyrolysis is between 900 and 1200  C for a rather short duration 0.1–1 s. Vacuum pyrolysis operated under a low pressure of about 0.05–0.2 MPa and a temperature range between 450 and 600  C. Intermediate pyrolysis is generally utilized to balance the proportion between liquid and solid products. Lastly, hydropyrolysis which occurs at a higher pressure of about 5–20 MPa is usually used to produce high-quality bio-oil. On the other hand, gasification is a partial combustion of solid and able to convert biomass into syngas by controlling the amount of oxygen. Gasification is capable of producing a gas fuel that can be used for direct heat generation or electricity generation. Several important parameters such as temperature, duration, particle size and pressure can be controlled in order to obtain the desired fractions of solid, liquid and gaseous products (Mohan et al. 2014). Table 3.1 illustrates the fraction of each product formed by using pyrolysis and gasification processes. Heat transfer in the biomass pores involves three mechanisms, including conduction inside the biomass particle, convection within the pores of the biomass material and radiation from the surface of the final product (Thines et al. 2017). The conventional heating process employs heat transferred from the material surface towards the centre through convection, conduction and radiation. This type of pyrolysis may encounter issues when the moisture content in the biomass is too

Table 3.1 Comparison of biomass conversion by using pyrolysis and gasification

Name of the process Slow pyrolysis (carbonization) Fast pyrolysis Gasification Mohan et al. (2014)

Product distribution (wt%) Solid Liquid (biochar) (bio-oil) 35 30

Gas (a mixture of gaseous product) 35

10 10

20 85

70 5

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high. In fact, the biomass is naturally a bad heat conductor which both disrupts and impedes the heat transfer mechanism. Besides, conventional method involves longer duration for heat transfer due to numerous mediums passing through and a larger amount of energy loss. To avoid the mentioned problems in conventional pyrolysis, an alternative heating method can be delivered through microwave heating which utilizes heat transfer from electromagnetic energy to thermal energy. Microwave heating has been demonstrated as a promising method due to its faster heating rate, selective heating wavelength and uniform heating (Li et al. 2016). Consequently, it could accelerate the reaction rates and increase energy efficiency. Microwave heating is primarily depended on two molecular effects induced by the microwave, which are the dipolar polarization and interfacial polarization. Dipolar polarization occurs when polar molecules (i.e. water) are forced to rotate, vibrate and move around in an accelerated rate. The increment of molecular rotation/movement will lead to the increment of molecular friction and the tendency of molecular collision. As a result, the presence of a large concentration of polar molecules in the biomass can be favourable in microwave heating operations.

3.3.2

Hydrothermal Carbonization

In 1913, Bergius reported that hydrothermal carbonization was suitable to convert biomass with high water content into products of higher energy density. Then, Bergius (1928) devised a method to carbonize cellulose into coal-like products by the introduction of steam to avoid superheating under hydrothermal condition. According to Hu et al. (2010), the hydrothermal condition was applied when subjected to an aqueous medium at a temperature and pressure exceeding 100  C and 0.1 MPa, respectively. Water serves as the hydrothermal solvent for the conversion of biomass in order to avoid the requirement of the predrying stage for wet biomass feedstock. Many minerals, peat and coal can be formed under these circumstances. Typical hydrothermal carbonization processes can be classified into two main categories which are high-temperature (300  C to 800  C) and low-temperature process (less than 300  C). High-temperature hydrothermal carbonization process is commonly used to produce carbon nanotubes and three-dimensional carbon structures. Meanwhile, low-temperature hydrothermal carbonization process is used to produce various carbonaceous materials with tailored morphological structure, carbon-based nanocomposites and porous carbon materials. Tuning the hydrothermal conditions can facilitate varying physical and chemical interaction between reagents and the solvent, resulting in the formation of carbon materials with different chemical properties and morphologies. Hydrochar term is used instead of biochar since it is a water slurry product. Hydrochar particle is usually uniform in spherical shape, and its porosity and functional surfaces such as hydroxyl ions (OH
) can be controlled. The chemical structure of hydrochar produced is quite similar to biochar generated from pyrolysis

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or gasification. However, hydrochar has higher hydrogen/carbon (H/C) ratio and oxygen/carbon (O/C) ratio as compared to biochar (Yuan et al. 2017). Biochar with dense aromatization is more resistant to oxidation and microbial degradation. Furthermore, hydrochar can be treated as an alternative precursor for producing highly porous ACs due to the high content in carbon, low ash content, limited aromatization as well as plenty of oxygenated functional groups (Islam et al. 2017). Hydrothermal treatment of biomass usually consists of five procedures, i.e. hydrolysis, dehydration, decarboxylation, polymerization and aromatization (Deng et al. 2016). As a result, carbonization processes are significantly complex and comprise of many different reactions which lead to the production of not only the hydrochar but other liquid and vapour products such as water vapours, oxides of carbon, hydrogen and various acids (Jain et al. 2016; Yue and Economy 2017). Additionally, the production of these acids will enhance the hydrolysis reactions and promotes the decompositions of oligomers and monomers into smaller constituents which suggested that solution pH plays a tremendous role during hydrothermal process. According to Titirici et al. (2015), factors such as the types of biomass, pH, temperature, reaction time and the solid content of the biomass were crucial in determining the range of products for hydrothermal carbonization. According to Cao et al. (2013) and Jain et al. (2016), the hydrochar product was found to increase with a temperature below 300  C. The further increment of temperature will lead to the production of more liquid and vapour products. It can be concluded that higher temperature and longer reaction time will transform the hydrochar to become chemically and spectroscopically similar to lignite. Besides, Yue and Economy (2017) stated that in order to produce primarily solid hydrochar only, the temperature should be around 200  C. Meanwhile, hydrothermal carbonization of biomass at low temperature (180–250  C) under pressure in water will produce primarily hydrochar (Xiong et al. 2017). Hence, the hydrothermal temperature might vary according to the properties of the hydrochar and the biomass feed. Cao et al. (2013) found that biomass with a high content of lignin typically required higher hydrothermal carbonization temperature. Cellulose-based hydrochar in the presence of high lignin content will acquire greater stability in terms of its crystalline structure due to higher-temperature resistance during hydrolysis reaction. Furthermore, they had also reported that high ratio of H/C and O/C indicated a higher degree of carbonization in the hydrochar produced and the presence of primary plant-based macromolecule (i.e. cellulose). As a result, lignin-rich feed typically resulted in hydrochar which has a lower degree of carbonization. Jain, Balasubramanian and Srinivasan (Jain et al. 2016) reported that a good quality of hydrochar contains a great amount of oxygenated functional groups (OFG) due to the higher surface area and high-porosity carbon materials. Since OFG is produced by the hydrolysis of biomass, a higher hydrothermal carbonization temperature will be favourable in OFG production. However, the increment of OFG will reach optimum amount at about 275  C and decrease with further increment in temperature due to its decomposition process.

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Y. L. Pang et al.

Physical and Chemical Activation

The limitations of biochar in its application include insufficient surface area, porosity as well as functional groups (Chen et al. 2017b). Biochar possesses less surface area as compared to AC before any treatment or modification (Gupta et al. 2015). Activation of biochar has been conducted to overcome these issues. The common activation method used is either through the physical or chemical activation process. After activation, the pore structure, specific surface area, functionality and sorption capacity of biochar would be enhanced significantly (Cha et al. 2016). Meanwhile, the hydrochar produced from hydrothermal carbonization process usually does not possess sufficient qualities such as porosity to become an efficient carbonaceous support for the catalyst. As a result, further treatment step needs to be taken to enhance the porosity and surface area of the hydrochar while preventing the formation and deposition of tar-associated products in the voids of the carbon skeleton. This process is known as the activation of carbon. Carbon activation can be conducted in two ways, (i) physical (or thermal) activation using air, carbon dioxide or steam at 800–900  C and (ii) chemical activation using acid activator such as phosphoric acid (H3PO3), sulphuric acid (H2SO4) and hydrochloric acid (HCl), whereas the alkali activator could be sodium hydroxide (NaOH), potassium hydroxide (KOH) and ammonia (NH3) typically in the temperature range of 450–650  C. Molina-Sabio et al. (1996) compared the physical activation by carbon dioxide and steam in terms of the microporosity of the product. They stated that activation using carbon dioxide would result in a larger volume of micropore and a narrower micropore size distribution as compared to steam due to microporosity being widening at the early stage of the activation process. The release of excess syngas such as H2 would increase the surface area and pore volume. Among the parameters, activation time and the amount of steam applied are the important parameters to control the physical and sorption capacity of biochar (Rajapaksha et al. 2016). During the activation process, steam or carbon dioxide (CO2) will react with carbon to produce carbon monoxide (CO), carbon dioxide (CO2), hydrogen (H2) or methane (CH4). This process is referred to as burn-off. Burn-off is defined as the weight differences between the char and the AC divided by the weight of the original char based on the dry basis (Mohamed et al. 2010). The development of materials porosity can be achieved through the burn-off process. Ergun (1956) and Reif (1952) proposed a reasonable reaction mechanism for the carbon gasification in the presence of carbon dioxide (Eqs. (3.1) and (3.3)) and steam (Eqs. (3.2)–(3.3)). C þ CO2 ðActivation agentÞ ! CðOÞ þ CO

ð3:1Þ

C þ H2 O ðActivation agentÞ ! CðOÞ þ H2

ð3:2Þ

CðOÞ ! CO þ C

ð3:3Þ

where C represents the free active sites on the carbon material surface and C (O) represents the oxygen surface complex. Both mechanisms involve the adsorption

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of the activating agent onto the surface of the carbon material and the generation of carbon monoxide. Meanwhile, Ahmadpour and Do (1996) investigated the effect of physical and chemical activation on AC with the desired surface area and pore size distribution. They found that chemical activation was effective to prepare highly porous materials, while physical activation might cause the excessive destruction of the carbon structure, carbon gasification, burn-off and reduction of porosity and surface area due to the formation of macropores. Among the chemical agents used for chemical activation, KOH is applied extensively due to its ability to produce AC with the high specific surface area and pore volume (Huang et al. 2015). Marsh et al. (1984) reported that chemical activation was an irreversible process where the porosity and surface area of the carbon material could be enhanced. The typical chemical agent utilized was KOH and other various alkaline metal compounds. The reaction between the carbon materials and KOH was found to be able to disturb and alter the lamella-layer-like crystalline structure of carbon materials into wrinkle-like structures. The unevenness of this structure caused the formation of voids between carbon layers that will contribute to the enhancement of overall surface area and porosity. It was also reported that KOH activation will lead to the formation of the hydrophilic surface due to the formation of hydroxyl groups (Viswanathan et al. 2009). Guo et al. (2002) proposed the reaction mechanism between KOH and carbon materials. The reaction began with the formation of potassium oxide (K2O) and water by two KOH molecules at high temperature (Eq. (3.4)). Meanwhile, the reaction between KOH and carbon precursor can result in the formation of functional groups such as –OK using oxygen of the alkali salt (Eq. (3.5)). Subsequently, K2O will react with carbon dioxide to produce potassium carbonate (K2CO3). This eventually lead to the gasification of carbon, where carbon molecules can escape from the overall structure via the formation of carbon monoxide (Eqs. (3.7)–(3.8)). 2KOH ! K2 O þ H2 O

ð3:4Þ

4KOH þ C ! K2 CO3 þ K2 O þ 2H2

ð3:5Þ

K2 O þ CO2 ! K2 CO3

ð3:6Þ

K2 O þ C ! 2 K þ CO

ð3:7Þ

K2 CO3 þ 2 C ! 2 K þ 3 CO

ð3:8Þ

Additionally, potassium metal generated from the activation process will intercalate to the carbon matrix. The carbon material was washed with water after activation process to remove the additional potassium salts present in the carbon particles. At the same time, it would create interlayer voids, which was beneficial to the enhancement of surface area and porosity (Viswanathan et al. 2009). Viswanathan et al. (2009) stated that alkaline metal carbonates were good substitutes to highly corrosive metal hydroxides and transition metal salts. This was due to their ability to remove hydrogen and oxygen atoms from the carbon structure instead of

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gasification of carbon to produce high carbon content material. These reactions involved chemical transformations such as dehydration, degradation and condensation. However, the excess phosphoric acid will promote gasification of char and increased the total weight loss of carbon. Caturla, Molina-Sabio and RodríguezReinoso (Caturla et al. 1991) reported that impregnation with zinc chloride (ZnCl) would result in degradation of the cellulosic material, carbonization and dehydration that promoted aromatization of the carbon skeleton and creation of the pore structure. The types and amount of activating agent, as well as the operating temperature, are the important variables to the activation rate of biochar. Chemical activation was found to possess higher efficiency than physical activation (Cha et al. 2016). It was identified that higher carbon yield, larger surface area and better microporosity were obtained during chemical activation (Chen et al. 2017b). Although there are many advantages in the chemical activation process, it also has several disadvantages such as corrosion on the equipment, intricate chemical recovery and higher cost of chemicals.

3.4

Photocatalysts and Photocatalysis Reactions

Semiconductor photocatalysis has attracted an increasing interest to solve the environmental issues such as water and air purification. Semiconductors such as TiO2, ZnO, Fe2O3, WO3, g-C3N4, CdS and ZnS can act as photosensitizers due to their electronic structure characteristic which consists of a filled valence band and an empty conduction band. An electron (e
) will be promoted from the valence band to the conduction band, leaving a hole (h+) at conduction band if a photon with an energy of hv matches or exceeds the bandgap energy (Eg) of the semiconductor (Hoffmann et al. 1995). This electron and hole may recombine and dissipate as heat or move towards the semiconductor surface to induce the desired redox reactions. The schematic of the overall photocatalysis mechanism is shown in Fig. 3.2. The electrons and holes will initiate a series of oxidation and reduction reactions to produce radicals that have tremendous oxidizing strengths. These radicals will then participate in the degradation of harmful organic pollutants. The mechanism for the water-splitting technology using photocatalyst to generate oxidizing radicals is summarized from Eqs. (3.9), (3.10), (3.11), (3.12), (3.13), (3.14), (3.15), and (3.16) (Pawar and Lee 2015). Eqs. (3.17), (3.18), (3.19), and (3.20) show the degradation of organic pollutants. Photocatalyst þ hv ! e
þ hþ

ð3:9Þ

hþ þ H2 O ! • OH þ Hþ

ð3:10Þ

þ




h þ OH ! • OH

ð3:11Þ

3 Biomass-Based Photocatalysts for Environmental Applications

67

Energy

Energy Level

Reduction Reactions

Recombination

Electron e

e

h+ e

Energy

Excitation

Recombination

Conduction Band

Band Gap

h+

h+

Valence Band

e

Catalyst Particle Surface

Hole E=hv ≥ Band Gap Oxidation Reactions

Fig. 3.2 Schematic diagram for photocatalysis process. An electron (e
) will be promoted from the valence band to the conduction band, leaving a hole (h+) at conduction band if a photon with an energy of hv matches or exceeds the bandgap energy (Eg) of the semiconductor

e
þ O2 ! • O2


ð3:12Þ

• O2
þ Hþ ! • OOH

ð3:13Þ

• OOH ! O2 þ H2 O2

ð3:14Þ

H2 O2 þ • O2
! • OH þ OH
þ O2

ð3:15Þ

H2 O2 þ hv ! 2 • OH

ð3:16Þ

hole þ Pollutants ! Degraded pollutants

ð3:17Þ

• OH þ Pollutants ! Degraded pollutants

ð3:18Þ

• OOH þ Pollutants ! Degraded pollutants

ð3:19Þ

O2
þ Pollutants ! Degraded pollutants

ð3:20Þ

Generally, the major problem for the development of photocatalysts for industrialization usage is related to the quantum efficiency of the photocatalyst. The major drawbacks include (1) high photo-generated electron-hole pair recombination rate, (2) photocatalyst only able to absorb a narrow portion of solar irradiation due to the

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wide bandgap energy, and (3) other physical constraints such as the surface area to volume ratio (Manoj et al. 2011). For instance, titanium dioxide (TiO2) is one of the most famous explored photocatalysts due to its low toxicity, cheap and special photo-oxidative properties. Despite its superiority in many desired qualities of a photocatalyst, the photocatalytic property of TiO2 can only be induced when the incident photons are in the ultraviolet (UV) spectrum due to its large energy bandgap (Bora and Mewada 2017). Furthermore, high electron-hole recombination rate is also another factor that deteriorates its catalytic activity. In order to synthesize an efficient and proficient TiO2 photocatalyst, many types of research have investigated the structural aspect improvement of the TiO2 such as doping and coupling techniques. One of the most popular improvements to the TiO2 photocatalyst is by introducing TiO2 particles to carbonaceous materials. Carbonaceous materials are no doubt one of the most abundant materials in the global ranging from domestic household organic-based waste to freshly made charcoal. In recent years, the utilization of carbon-based materials has gained wide attention among researchers, particularly on environmental remediation purposes due to their intriguing properties and enhancements compared to conventional metal oxide catalysts. Carbon-based supports have several advantages such as hydrophobic surfaces, high surface area, large pore volume, chemically inert, thermally and mechanically stable, easy to handle and cheap manufacturing cost (Colmenares et al. 2016). Moreover, biomass also possesses good electrical conductivity and outstanding chemical stability. The adsorption efficiency of pollutants onto the surfaces of the TiO2 photocatalyst was also found to improve when paired with carbonaceous based materials (Perera et al. 2012). Photocatalytic properties of TiO2 only allow it to absorb light from the UV spectrum which only accounts for around 5% of the incident sunlight. By pairing TiO2 catalyst with carbonaceous nanomaterials, the light absorption range of TiO2 can be expanded to the whole visible light range due to the increased absorption by the Ti-O-C bonds found in between the TiO2 and the carbon-based material (Khalid et al. 2017). As a result, this has widened the application of TiO2-based photocatalyst and simultaneously improved its usability and flexibility under ambient conditions. Another significant quality of coupling carbonaceous materials with the TiO2 photocatalyst is that carbonaceous materials are able to enhance electron-hole separation by acting as an electron reservoir which is able to reduce the rate of electron-hole recombination (Lin et al. 2011). Based on the several merits of the application of carbonaceous materials with TiO2 photocatalyst stated above, carbon-based materials may play an important role in the future of photocatalysis as it is one of the promising solutions to the problems revolving around the synthesis of an industrially suitable photocatalyst.

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3.5 3.5.1

69

Functionalized AC and Applications Types of Functionalized AC

In recent years, the development of AC derived from biomass is making headway as a carbonaceous material used mostly for adsorption. The physical and chemical structures of AC can be modified to enhance its adsorption capability for higher uptake of target pollutants. Apart from adsorption, AC is also being functionalized to cater for different applications such as capacitance enhancement (Li et al. 2017a). In general, functionalization of AC is known to improve its surface properties through the introduction of various functional groups on its surface (Mohammad et al. 2016). The existing surface functional group can be modified through thermal or chemical post-treatments. To date, AC can be modified through acid and alkali treatment as well as metal or metal oxide impregnation. The acid treatment utilizes strong acids such as nitric acid (HNO3), sulfuric acid (H2SO4) and hydrochloric acid (HCl) under various temperature and reaction conditions to increase the acidity of oxidative functional groups such as carboxylic acids, alcohols and lactones (Sweetman et al. 2017). On the other hand, alkali or base treatment of AC using sodium hydroxide (NaOH), ammonia (NH3) and potassium hydroxide (KOH) tends to generate amides and aromatic amines with basic functionality. Also, oxygen species are removed when the high temperature is used with basic catalyst treatment of AC. The selection of modification or functionalization methods depends on the nature of application since AC will become more selective towards specific species after modification. To date, many studies on surface modification or functionalization have been done on biomass-derived AC to enhance their adsorption ability for pollutant compounds. Table 3.2 lists several sources of biomass used to develop functionalized AC and their applications in environmental remediation. Majority of the functionalized AC were used as adsorbents for water remediation. Typically, AC contains atoms ranging from oxygen, nitrogen, sulphur and hydrogen. However, the presence of oxygen plays the biggest role in determining the properties of an AC (Marsh and Rodríguez-Reinoso 2006). Usually, surface oxygen groups can be increased through oxidation and decreased through heating under inert condition (Figueiredo et al. 1999). In this case, gas phase oxidation increases the concentration of hydroxyl and carbonyl surface groups of AC, while liquid phase oxidation increases the concentration of carboxylic acids. Further modification of surface oxygen groups is commonly done after the activation step, where nitric acid oxidation can be used to incorporate acidic oxygen functional groups, which was found to be effective for cadmium removal (Rodríguez-Estupiñán et al. 2017). Usually, carboxyl, quinone and phenol groups can be detected after the oxidation process. According to Fu et al. (2015), oxidation of AC using strong acids could lead to the insertion of surface carboxyl groups, which can be utilized in the subsequent surface chemistry of AC. These functional groups, especially carboxyl and phenol groups, are useful for the adsorption of heavy metals. In this situation, phenolic groups form complexes with heavy metal, while carboxyl groups boost the uptake of

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Table 3.2 Functionalized/modified activated carbon (AC) derived from biomass Source of AC Almond shells

Broom sorghum stalk

Coconut shells

Functional group Total basicity increased: Decreased carboxylic acids, phenols and lactones Surface amide groups

The oxygencontaining groups: Phenolic group Latone group Carboxylic group

Increased surface basicity:

Corncob

Olive palm bunch

Incorporation of nitrogen functionality Reduced oxygen content Acidic oxygen groups

Treatment Ammoxidation

Findings 78% removal of total organic carbon from Tunisian industrial phosphoric acid

References Omri et al. (2013)

Modified with triethylenetetramine (TETA) and urea

CO2 adsorption (mmol.g
1): (1) Unmodified: 1.66 (2) TETA modified: 3.20 (3) Urea modified: 2.33 O-xylene uptake capacity (mg.g
1):

Mehrvarz et al. (2017)

Treated with

(1) Ammonia (2) Sodium hydroxide (3) Nitric acid (4) Phosphoric acid (5) Sulphuric acid Treated with NH3 at 700  C and under argon condition at 1000  C

Li et al. (2011)

(1) 305.70 (2) 295.40 (3) 241.69 (4) 220.74 (5) 189.46 Perfluorooctane sulfonic acid uptake capacity (mg/g): (1) Native: 1.72

Zhi and Liu (2016)

(2) Modified: 17.0 Nitric acid oxidation

Phosphate and phenolic groups

Microwave-assisted phosphoric acid activation

Nitrogen species

Ammonia

Improved removal of phenol, Pb2+ and methylene blue Generated a large number of surface groups with high capacity towards copper ions Removal of 2,4-dichlorophenol (2,4-DCP) (mg/g): 285.71

El-Hendawy (2003) Trofymenko et al. (2015)

Shaarani and Hameed (2011) (continued)

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Table 3.2 (continued) Source of AC Rambutan peel

Functional group –

Treatment Microwave-assisted acid treatment using KOH

Rice husk

Oxide content increased

Acid treatment using HNO3

Wood (eucalyptus)

Increased basic functional groups

Heat treatment with ammonia solution

Increased surface basicity:

Treated with NH3 at 700  C and under Ar condition at 1000  C

Incorporation of nitrogen functionality Reduced oxygen content

Findings Removal of acid yellow 17 (mg/g): (1) Native: 133.3 (2) Modified: 215.05 Adsorption of Pb (II) (mg/g): (1) Native: 66.23 (2) Modified: 95.24 CO2 adsorption capacity (mmol/g): (1) Modified: 3.22 (2) Olive stone: 0.77 (3) Almond shell: 0.80 Perfluorooctane sulfonic acid uptake capacity (mg/g): (1) Native: 2.70

References Njoku et al. (2014)

Yao et al. (2016)

Heidari et al. (2014)

Zhi and Liu (2016)

(2) Modified: 85.2

heavy metals through the formation of complexes on the AC surface (González and Pliego-Cuervo 2014). When comparing to the aforementioned surface functional groups, nitrogen functional groups are relatively scarce in AC (González-García 2015). Activation or modification methods are needed to integrate nitrogen functional groups either through doping or functionalization for specific applications. To date, several chemical post-treatment methods have been introduced to incorporate nitrogen on AC, which includes strong oxidation using nitric acid (Salem and Ebrahim Yakoot 2016) as well as thermal treatment in the presence of ammonia, urea or melamine (Alabadi et al. 2016). To promote effective adsorption of negatively charged target pollutants, especially organic species in liquid, alkaline treatment using urea and ammonia atmosphere is ideal as it imparts positive surface charges on AC (González-García 2015). Phenol uptake capacity was found to be higher in such cases (Stavropoulos et al. 2008) and could be due to the better electrostatic interaction between adsorbateadsorbent. In addition, studies pertaining to the functionalization of almond shells (Omri et al. 2013), broom sorghum stalk (Mehrvarz et al. 2017), coconut shells (Zhi and Liu 2016; Li et al. 2011), olive palm bunches (Shaarani and Hameed 2011), rice husk

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(Yao et al. 2016) and eucalyptus wood (Heidari et al. 2014) utilized this concept to remove gas and water pollutants ranging from acids, carbon dioxide, o-xylene and phenol-containing compounds. Their findings are summarized in Table 3.2. All of the studies showed the enhanced removal of pollutants after modification. Furthermore, functionalization of AC to increase the surface amide groups for better carbon dioxide adsorption was proven to be better with triethylenetetramine (TETA) (3.20 mmol.g
1) and urea (2.33 mmol.g
1) (Mehrvarz et al. 2017). Conversely, HNO3 and oxygen gasified treated AC showed lower phenol adsorption capacity due to the nature of its acidic surface functionality. For compounds such as o-xylene, base treatment of AC (305.70 mg.g
1) would be preferable over acid treatment (241.69 mg.g
1) due to better performance. According to Wang et al. (2017a), this occurrence could be attributed to the presence of additional nitrogen which plays an important role in enhancing carbon dioxide adsorption. On the other hand, sulphur functionalities of AC could be generated through its reaction with carbon disulphide (CS2), hydrogen sulphide (H2S) and sulphur dioxide (SO2) followed by interaction with oxygenated acidic groups or through placement on empty sites via adsorption, chemisorption and capillary condensation (GonzálezGarcía 2015). Activated carbon fibres (ACF) impregnated with a sulphur monolayer were found to be an excellent adsorbent for mercury uptake (Feng et al. 2006). However, an increase in sulphur content would block the access to the first layer and fills the sorbent pores. Therefore, the pore volume would reduce which led to lower mercury uptake as it became less accessible. Phosphorus functionalities can also be incorporated on AC after polymerization of carbon during chemical activation using phosphoric acid. Phosphorus compounds are usually found in AC through the incorporation of nitrogen using urea and melamine as precursors (HulicovaJurcakova et al. 2009) or from phosphorus doping using phosphoric acid (H3PO4) (Chen et al. 2014). However, surface phosphorus functionalities of AC are mostly used to boost electrochemical capacitance (Hasegawa et al. 2015) or serve as an efficient oxygen reduction catalyst for microbial fuel cells (Lv et al. 2018). Conventionally, many post-treatment methods to functionalize AC use harsh chemicals (acids) and ozone treatment techniques. In recent years, microwaveassisted chemical activation methods are used to functionalize AC, which could be beneficial in terms of lower power consumption and treatment time. Recent studies which include the functionalization of corncob (Trofymenko et al. 2015; El-Hendawy 2003) and rambutan peels (Njoku et al. 2014), had successful modified the surface groups to enhance adsorption of heavy metal (copper ions) and Acid yellow 17 dye, respectively.

3.5.2

Functionalized AC Photocatalysts and Its Application

The usage of AC on its own as an adsorbent is limited by its adsorption capacity. In addition, organic pollutants which remain trapped on AC after adsorption requires further thermal and chemical regeneration or disposal. In this content, there are three

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issues involved: (1) thermal regeneration degenerates the carbon and changes its structure; (2) chemical regeneration can produce additional waste streams from chemical consumption; and (3) hazards are not eliminated since pollutants remain in its dangerous state after being adsorbed onto the AC (Bagheri et al. 2015). Apart from utilizing acid and alkali treatment as a modification method, AC can also be modified by impregnating metal or metal oxide nanoparticles onto its porous surface. This technique which incorporates the application of adsorption and heterogeneous photocatalysis has shown progress in recent years. The integration of functionalized AC and semiconductor catalyst as a hybrid photocatalyst has proven to be a beneficial combination which provides a synergistic effect for adsorption and photodegradation in an advanced oxidation process system (Zawawi et al. 2017). With AC, many shortcomings of metal oxides such as TiO2 can be lifted. Its application is commonly limited by its poor adsorption ability, easy aggregation and high filtration cost (Fu et al. 2015). Furthermore, AC is dubbed as a good lightabsorbing compound which is suitable to be used as a photoactive species (Smits et al. 2014), making it a potential material for the development of photocatalysts. In general, metal oxide semiconductors with sufficient bandgap energy such as TiO2 and ZnO are suitable candidates for the hybrid system (Bagheri et al. 2015). For example, a previous study (Velasco et al. 2012) had displayed improved photooxidation of phenol with AC compared to native TiO2, proving it to be a compatible match. In the past, bare TiO2 photocatalysts had shown successful photo degeneration trials towards pollutants such as dyes, phenols and organic compounds which were found in wastewater (Bagheri et al. 2015). According to Chekin et al. (2013), TiO2 is a potential semiconductor for the photocatalytic destruction of organic pollutants due to its good stability and catalytic performance in aqueous media. As an alternative to TiO2, ZnO supported AC had also been applied for photodegradation of acid red 14 dye (Byrappa et al. 2006), alizarin cyanine green (Muthirulan et al. 2013) and Congo red dye (Joshi and Shrivastava 2011). In another study, AC had displayed synergistic effect of decreasing and eliminating ciprofloxacin (CIP) during photocatalytic degradation by ZnO (Nekouei et al. 2018). Furthermore, AC also was capable of increasing catalytic efficiency of ZnO in secondary degradation of intermediates. A photocatalyst should be efficient, non-toxic, stable and inexpensive. Both TiO2 and ZnO are qualified based on these criteria. The impregnation of TiO2 onto AC can be performed in several different techniques: (1) chemical vapour deposition (CVD), (2) hydrothermal, (3) aerosol pyrolysis, (4) direct air-hydrolysis, (5) dip coating, (6) precipitation and (7) sol-gel (Bagheri et al. 2015; El-Sheikh et al. 2004). The selection of impregnation techniques is highly dependent on the properties of AC and types of targeted pollutant. For better adherence of TiO2 to AC surface, the wet synthesis method is recommended, while chemical vapour deposition using water and titanium tetraisopropoxide vapour is more ideal for nanosized metal oxides (Bagheri et al. 2015). The hybrid system between AC and TiO2 works in a way that AC acts as an efficient adsorbent to bring target pollutants closer to TiO2 for fast, immediate and

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effective photodegradation. Although functionalized AC in hybrid photocatalysts is known to possess better surface properties, Mohammad et al. (2016) observed a relatively lower specific surface area for the hybrid with functionalized AC when compared to its native counterpart. The smaller specific surface area could be attributed to the presence of surface oxygenated groups. Furthermore, the photocatalytic activity was lower as repulsion between the functionalized surface and dye molecules could be present. It is evident that the surface chemistry of AC (modified through functionalization) should be made under the jurisdiction of the targeted pollutant’s characteristic. This is further supported by a study done by Gulyas and co-workers (2013) who had combined adsorption and photocatalysis to treat gray water and tetraethylene glycol dimethyl ether (aliphatic compound). Although the lower organic content was found in gray water after treatment, the researchers found that powdered AC contributed to photocatalyst shading when dealing with photocatalytic oxidation of aliphatic compounds as compared with aromatic compounds (pure aqueous phenol). As a result, it is evident that targeted pollutants are sensitive to specific properties of AC which will affect the synergistic effect of photocatalytic oxidation and adsorption due to their complex interaction. It is also worthy to note that the utilization of real wastewater to study the synergistic effect of photocatalysis-adsorption could provide a better assessment of treatment efficacy as compared to using a pure aqueous solution of a single target pollutant. Conversely, Fu et al. (2015) reported better removal of Cr(VI) (28.4 mg g
1) with functionalized TiO2/AC-DETA photocatalyst compared to TiO2/AC or plain TiO2. The presence of DETA on the AC surface was proven to enhance the photoactivity of the photocatalyst by accelerating the mass transfer of aqueous Cr(VI) to the AC. In this case, the acceleration of mass transfer could be attributed to (1) the electrostatic interactions, (2) coordinative interaction and (3) hydrogen bonding interaction between surface amino groups and Cr(VI) ions (Fu et al. 2015). Furthermore, acid treatment and consecutive chloridization of –COOH on AC have provided carboxyl groups and acyl chlorides to facilitate closer contact between TiO2 and AC, which in turn promotes the interfacial transfer of reactive species to effectively remove Cr(VI) and convert it into a less toxic form, Cr(III). To date, limited studies have been carried out to develop hybrid metal oxide/AC photocatalysts using biomass as precursors for its carbonaceous matrix. Table 3.3 shows several examples of metal oxide photocatalyst supported on carbonaceous materials that could be derived from biomass. Nozawa et al. (2001) demonstrated that TiO2 paired with fibre AC was able to enhance the adsorption rate of malodorants leading to better degradation of methyl mercaptan into sulphates and ammonia into nitrates under ultraviolet irradiation. The performance of AC as reactant absorber was investigated by Ao and Lee (2005). They found that the degradation rates of nitrous oxide and toluene were significantly enhanced when TiO2/AC filters were used compared to pure TiO2. In addition, El-Sheikh et al. (2017) demonstrated that AC coupled with nitrogen co-doped TiO2 was able to degrade ibuprofen with 100% efficiency under visible light. The AC and nitrogen were doped onto TiO2 to modify the bandgap energy. Hence, the charge separation

Granular AC

(2 wt% AC)

TiO2

AC

AC derived from Garcinia mangostana

Basic red 46 dye (BR46)

AC

N-TiO2

Basic red 18 dye (BR18)

AC

AC and nitrogen co-doped mesoporous TiO2 Fe ions doped TiO2 (0.3% Fe- TiO2/AC) TiO2 nanoparticles

Chitosan decorated magnetic AC

Brilliant red dye (K2G)

Spherical AC

Er3+: YAlO3 doped TiO2

TiO2 co-doped with iodine and Nitrogen

Ibuprofen

AC filter

TiO2

Humic acid

Remazol brilliant blue dye

Salicylic acid

Methyl orange dye (MO)

Toluene, nitrous oxide and other oxides of nitrogen (NOx) in air

Methyl mercaptan (MM), ammonia and hydrogen sulphide (H2S)

Targeted pollutants Malodorants:

Carbonaceous materials Fibre AC

Catalyst TiO2

Degradation rate constant: 0.0229 min
1 Chemical oxygen demand removal percentage of BR18: 81.2% (150 min) BR46: 76.2% (150 min) Degradation efficiency: 89.71% Degradation efficiency: 80% (6 h) Degradation efficiency: 99.5%

Significant findings Degradation rate constant of MM: 13.1 h
1 Ammonia: 4.3 h
1 H2S: 4.0 h
1 Degradation efficiency: Toluene: 89.5% Nitrous oxide: 97.0% NOx: 95.4% Degradation efficiency: 90.8% Degradation efficiency: 100%

Table 3.3 Some examples of metal oxide photocatalyst with the carbonaceous support that can be derived from biomass

UV (365 nm)

Visible light

Visible light

UV (200–280 nm)

UV (320–400 nm)

Visible light

LED (>400 nm)

UV (254 nm)

UV (254 nm) UV (365 nm)

Wavelength UV (254 nm)

(continued)

Xue et al. (2011)

Leong et al. (2017)

Wang et al. (2016a)

Mahmoodi et al. (2011)

Li et al. (2009)

El-Sheikh et al. (2017)

Dong et al. (2012)

Ao and Lee (2005)

References Nozawa et al. (2001)

3 Biomass-Based Photocatalysts for Environmental Applications 75

Methylene blue dye

AC

Hollow AC fibres (HACF)

Waste biomass AC nanocomposites (WBAC) Konar bark AC

Iraqi date palm seeds AC

Acid-treated AC

AC

Fe-doped TiO2

TiO2-MnTiO3

ZnO (3:1 ZnO to WBAC weight ratio)

ZnO

Pt/TiO2

F-ag-β-cyclodextrin co-doped TiO2

ZnO

Vehicle exhaust: Hydrocarbons (HC), CO, CO2 and NOx

Sugar cane bagasse made AC

TiO2 (20 wt% TiO2)

Celestin blue b dye (CBB), C17H18ClN3O4 Ethylenediaminetetraacetic acid disodium salt (EDTA-2Na) Naphthalene

Ciprofloxacin (CIP)

Orange G dye

Methylene blue dye

Basic yellow 28 dye (BY28)

AC

Barium ferrite (10 g/L barium ferrite/AC)

Targeted pollutants Acid orange 7 dye (AO7)

Carbonaceous materials AC

Catalyst Ag2CrO4

Table 3.3 (continued)

Degradation efficiency: ~40% Cumulative removal efficiency: 59% Degradation efficiency: 98.4%

CIP removal efficiency: 95% (pH 6.5)

Degradation efficiency: 96% (180 min) Degradation efficiency: HC: 10.9% (15.23% Ti) CO: 5.8% (9.86% Ti) CO2: 7.7% (15.23% Ti) NOx: 67.6% (19.07% Ti) Degradation efficiency: 99.1% (60 min) Degradation efficiency: >90% (pH 6.45)

Significant findings Degradation rate constant: 0.00304 s
1 Degradation efficiency: 100% (30 min)

Visible light

UV light (420 nm and above) UV light (642 nm) UV

UV

Visible light

Visible light

Visible light

Visible light

Wavelength Visible light

Chen et al. (2017a)

Mohammad et al. (2016) Su et al. (2015)

Nekouei et al. (2018)

Vinayagam et al. (2016)

Li et al. (2017b)

Hu et al. (2017)

Roonasi and Mazinani (2017) El-Salamony et al. (2017)

References Azami et al. (2018)

76 Y. L. Pang et al.

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was enhanced due to the synergistic effect provided by different crystalline phases in TiO2 as well as the heterojunction between TiO2 and AC or nitrogen. Furthermore, Li et al. (2009) discovered that TiO2 would be dispersed more evenly on AC support when doped with Fe ions as it would reduce the size of TiO2 nanoparticles and enhance photocatalytic activity by increasing the rate of charge separation. They found that when Fe-doped TiO2 was immobilized on AC surface, the rate of transport of the adsorbed pollutant to the catalyst was also improved, thus leading to a greater brilliant red K2G dye degradation rate. However, Hu et al. (2017) found that the excessive loading of Fe-doped TiO2 had an adverse effect when observing the degradation of vehicle exhausts as it will lower the specific surface area, total pore volume and the average pore volume of the catalyst which reduced its photocatalytic abilities. They also found that the narrower pores on the AC which attached the Fe-doped TiO2 have the greatest photocatalytic ability. AC produced from biomass source (canola hull) was also capable of improving the photocatalytic activity of TiO2 in the degradation of textile dyes without the need of high pressure and temperature (Mahmoodi et al. 2011). El-Salamony et al. (2017) also demonstrated that after incorporating AC made from sugar cane bagasse into TiO2, the degradation of methylene blue was found to be 96% under visible light irradiation. Another essential requirement of an effective photocatalyst is that the photocatalyst must be easily separable from the treated effluent since supports are normally deployed only to improve the quality of the catalyst. A method of pairing a magnetic AC with iodine and nitrogen co-doped TiO2 was devised by Wang et al. (2016b) so that the coupled catalyst can be separated by a magnetic field from the treated water. The presence of synergistic effects between TiO2 and AC could enhance charge transfer and separation. Xue et al. (2011) found out that the degradation of humic acid was improved when granular AC was used as a TiO2 support. AC was also found to be able to produce pH-resistant photocatalyst in terms of photocatalytic performance as demonstrated by Su et al. (2015). They paired Pt/TiO2 with acid-treated AC and found that after the incorporation of acid-treated AC, the influence of pH on the photocatalytic activities decreased drastically and had a higher hydroxyl and carboxylic acid functional group absorption capacity which improved the photocatalytic ability of the catalyst. While the previous approaches in improving the photocatalytic abilities of TiO2 focused greatly on bandgap modification of TiO2 to extend its ability of visible light absorption, Li et al. (2017b) stated that if the bandgap of the catalyst was reduced, it would also reduce the tendency of the photo-redox reactions to occur. This approach was to introduce an agent which is sensitive to visible light via the creation of a heterojunction TiO2 catalyst with MnTiO3, and the composite is to be dispersed on hollow AC fibres. MnTiO3 was able to produce photo-generated electrons when irradiated with visible light while having a wide bandgap, reduce electron-hole recombination rates due to its higher conduction band edge and increase dye attraction. Similarly, the deployed hollow AC fibres were able to provide enhancements to the catalyst in terms of electron-hole recombination rate reduction, improved adsorption capability and catalyst dispersion. The overall catalyst was hierarchically porous which gave rise to more active sites, improved ease of recovery

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and tunable bandgaps. Another approach was demonstrated by Dong et al. (2012) where Er3+:YAlO3 (luminescence agent)-doped TiO2 was deposited onto the surface of spherical AC supports to promote degradation of methyl orange under visible light. The doped luminescence agent was used to convert visible light into UV light to trigger the photocatalytic activities of TiO2, while the spherical AC supports provide a high adsorption capacity, which enhances charge transfer and charges separation. Facets of TiO2 are also an important factor in photocatalysis as a 001 faceted surface is more photocatalytically reactive than the 101 faceted surface of TiO2. However, the formation of 101 faceted surface is more preferable due to its lower energy as stated by Chen et al. (2017a). As a result, a study was done to induce the formation of more 001 faceted surfaces using NaF on Ag-β-cyclodextrin co-doped TiO2 which produced F-Ag-β-cyclodextrin co-doped TiO2 before being incorporated onto AC. The structure β-cyclodextrin was able to capture hydrophobic molecules, pollutants and photo-generated electrons, while the pollutant adsorption rate of AC was further enhanced to produce an efficient photocatalyst with 001 faceted surfaces as the dominant surface. Vinayagam et al. (2016) demonstrated that AC was also able to improve the photocatalytic abilities of other conventional metal oxides such as ZnO. Similarly, due to the wide bandgap of ZnO, its photocatalytic is only limited to the UV region. Additionally, it was reported that ZnO also contains unwanted activities such as high photo corrosiveness and high electron-hole pair recombination rate. In order to improve the catalyst, a similar strategy for TiO2 catalyst was employed by incorporating AC to ZnO to form an intermediate energy level in the bandgap to allow a better visible light absorption and reduce electron-hole recombination rate due to the electron sink ability of AC. Roonasi and Mazinani (2017) paired AC with barium ferrite which was a much more sophisticated metal oxide photocatalyst with magnetic properties. The barium ferrite was able to absorb a large portion of visible light, had a great number of catalytic sites, inert, and was easily separable due to its magnetic property by nature without the need of incorporating external agent to induce magnetism. Similarly, AC was able to improve the adsorption capacity of the overall catalyst. The Ag-based catalyst can be improved when integrated with carbonaceous materials as reported by Azami et al. (2018). Ag-based catalysts were found to be excellent photocatalyst due to the angles of Ag-O bonds and O-Ag-O bonds in which the catalyst possesses a unique electronic and crystalline structure. Moreover, Ag-based catalysts also exhibit high photocatalytic activities under visible light irradiation due to their surface plasmon resonance effect. However, similar to TiO2, the electron-hole recombination rate of the Ag-based catalyst is significant due to their relatively small bandgap. The stability of Ag-based compounds when irradiated by visible light is also found to be unpromising. With the addition of AC into the catalyst, similar goals were achieved as in previous cases where the electronhole recombination rate was reduced and the specific surface area was increased; both of these properties are beneficial to the photocatalytic abilities of the catalyst. Additionally, the degradation of the catalyst was found to be reduced as AC was able

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to prevent the reduction of Ag+ to metallic Ag0, and the crystallinity of the catalyst increased which led to the enhancement of its photocatalytic abilities. Other than metal oxides, metal sulphide had also been used to synthesize new nanocomposites for the photodegradation of rhodamine B under visible light (Huang et al. 2017). In this study, AC (lotus seedpod) deposited with cadmium sulphide (CdS) demonstrated energy-efficient photocatalytic activity due to the presence of microporous carbonaceous support which enhanced the generation and transfer of photo-generated active species (h+, OH• and O2•
). The electron-hole separation was observed to be more efficient in the composite system compared with treatment using CdS only. However, metal sulphide groups are known to be unstable as they are prone to photoanodic corrosion (Bagheri et al. 2015) and may not be suitable for environmental application due to its toxicity.

3.6

Future Challenges and Conclusions

Agricultural waste-based photocatalysts offer various attractive features such as low-cost, non-toxic, biocompatible and high porosity. Structure of biochar or AC showed outstanding adsorption followed by oxidation capacity for various recalcitrant organic pollutants. However, there are still several research gaps that require more attention such as assessment of the individual roles for biomass-derived carbon coupled with the semiconductor to degrade multicomponent of organic pollutants, the performance of these materials to treat real industrial effluents, regeneration studies and continuous flow studies. A detailed discussion on the porosity and surface area development during activation of biomass-derived carbon either through physical or chemical methods may be the focus in future. Moreover, the fundamental mechanistic approach is needed to understand the role of individual components in the improvement of porosity and surface area. Although a lot of researchers have investigated the photocatalytic behaviour of biomass-derived carbon-modified semiconductors in literature, the detailed mechanisms for the photocatalytic degradation of organic pollutants from wastewater still need to be explored in depth. To date, most of the photocatalytic degradation studies are limited to scale basis. More studies should be extended to full or pilot scale to examine their potential in treating real wastewater. Besides, the current research mainly focuses on the performance of biomassderived carbon-modified semiconductors while ignoring the economic aspect. Further research on the assessment of economic feasibility and environmental impact will be deemed interested. A widespread and great progress in this biomass-derived carbon area can be expected in future despite facing various challenges currently. In short, agricultural waste-based catalysts could offer alternative approach with significant advantages over commercially AC due to lower cost to control water pollution and minimize the solid waste strategy.

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Acknowledgement The authors would like to acknowledge Universiti Tunku Abdul Rahman, Malaysia, for the financial support under UTARRF (IPSR/RMC/UTARRF/2018-C1/P01) through Centre for Photonics and Advanced Materials Research and Centre for Environment and Green Technology (CEGT). The financial support provided by the Fundamental Research Grant Scheme (FRGS/1/2018/TK10/UTAR/02/2) by the Ministry of Education (MOE), Malaysia, is also gratefully acknowledged.

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Chapter 4

Application of Bismuth-Based Photocatalysts in Environmental Protection Ewa Maria Siedlecka

Contents 4.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2 Photocatalytic Oxidation of Pharmaceuticals in Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2.1 Tetracycline . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2.2 Ciprofloxacin and Other Antibiotics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2.3 Carbamazepine . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2.4 Ibuprofen and Diclofenac . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2.5 Other Pharmaceuticals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3 Photocatalytic Oxidation of Industrial Micropollutants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3.1 Bisphenol A . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3.2 Oxidation of Other Industrial Pollutants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.4 Oxidation of the Indoor Air Pollutant NOx . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.5 Photocatalytic Reduction of Pollutants in Water and Air . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.5.1 Reduction of Cr(VI) in Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.5.2 Reduction of CO2 in Air . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.6 Water Splitting . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.7 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

88 90 90 99 100 101 103 105 105 106 108 110 110 111 111 113 114

Abstract Photocatalysis by use of advanced oxidation processes (AOPs) is gaining attention as an effective method of air purification and water treatment. Undoubtedly, photocatalysis can also be applied to produce useful fuels from photocatalytic reduction of CO2 and splitting of water, or it can be utilized as a “green” technology in industrial production. Despite recent research into other photocatalysts (e.g., TiO2 and perovskites), bismuth-based semiconductors such as Bi2O3, BiPO4, (BiO)2CO3, BiOX (where X ¼ Cl, Br, and I), and pentavalent bismuthates (e.g., NaBiO3) are most promising because of their low cost, nontoxicity, and high oxidizing and reduction abilities in solar and visible light. Moreover, the conduction band edge and the valence band edge of Bi-based photocatalysts can be designed by using a suitable strategy for preparation of these materials. The photocatalytic activity of E. M. Siedlecka (*) Department of Chemistry, University of Gdańsk, Gdańsk, Poland e-mail: [email protected] © Springer Nature Switzerland AG 2020 Inamuddin et al. (eds.), Nanophotocatalysis and Environmental Applications, Environmental Chemistry for a Sustainable World 30, https://doi.org/10.1007/978-3-030-12619-3_4

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Bi-based materials can be additionally enhanced by heterostructures, e.g., using carbon or graphene quantum dots, Ag/AgCl, modified TiO2, or Fe3O4. This chapter aims to highlight recent advancements in application of Bi-based photocatalysts and heterostructures in environmental protection. Albeit nonexhaustive, this review explores the progress made in the last 6 years by focusing on solar and visible light–driven degradation processes to eliminate such contaminants as antibiotics, nonsteroidal anti-inflammatory drugs, beta blockers, anticonvulsants, hormones, resorcinol, bisphenol A, and other derivatives of phenol, many of which have been detected in aqueous ecosystems. The application of Bi-based photocatalysts for removing NOx from indoor air using solar and visible light illumination is also presented. Finally, advances in water splitting and CO2 reduction to CO and CH4 with Bi-based photocatalysts are discussed. Keywords Pharmaceuticals · Micropollutants · Industrial pollutants · Bi-based photocatalysts · Modification of photocatalysts · Photooxidation · Photoreduction · Air purification · Water splitting · Water treatment

4.1

Introduction

Environmental pollution (mainly of air and water) and the search for renewable energy sources with applications that are nonhazardous to the environment are the main issues that need to be resolved in the near future. Up till now, the applied wastewater treatment technologies have been focused on biodegradable organic matter and biogenic compound removal, because of the threat of eutrophication. However, humans are currently facing new problems such as the presence of micropollutants (pharmaceuticals, surfactants, phthalates, bisphenol A, etc.) in waters and wastewater, which even in very low concentrations affect living organisms very negatively. Micropollutants are contaminants that are poorly biodegradable and insufficiently removed by conventional wastewater treatment technologies. The continued release of micropollutants in wastewater effluent is believed to cause long-term hazards, but the exact effects are not fully known. Combustion of fossil fuel, which supplies energy in the daily lives of people, is a source of air pollutants, including CO2 or NO. These gases are considered common air pollutants because they can cause a greenhouse effect, photochemical smog, or even pulmonary edema (Wang et al. 2016). Their concentrations in the atmosphere have greatly increased over the past few decades; therefore, development of efficient and economical technologies to eliminate NO and CO2 contamination and application of renewable energy sources are necessary. Photocatalysis, classified as a green technique, is regarded as a promising approach for environmental issues, including removal of water and air pollutants, sourcing of solar energy, and production of fuels. Photocatalytic materials and light sources are being applied to determine the potential applications of photocatalysis. An early photocatalytic material, which is still of interest to researchers, is the TiO2

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Bi2O3 , Bi2S3 M(BiO3)n

Bi2MO6 (M= Mo; Cr, W)

(n=1, M=Li, Na, K, Ag;, Fe n=2, M=Mg, Ba, Zn, Sr, Pb)

Bi containing PCs

BiVO4

(BiO)2CO3

BiPO4

BiOX (X=Cl, Br, I)

Fig. 4.1 Classification of Bi-based photocatalysts (PCs) used in environmental protection

semiconductor. Because of the large bandgap (3.2 eV) of TiO2, only ultraviolet (UV) light possesses enough power to trigger a photocatalytic process in the presence of this semiconductor. Although application of artificial UV light for photocatalytic activation of TiO2 is possible at the laboratory scale, it is a completely uneconomical approach on an industrial scale. Therefore, a hot research topic and a remaining serious challenge is development of photocatalysts that absorb light over a wide range of wavelengths and efficiently separate charge carriers, while simultaneously being photostable. Among different materials, bismuth-containing photocatalysts, with strong visible light absorption ability and tunable bandgap edges, have received tremendous attention. They exhibit good photocatalytic activity and stability during the reaction and usually possess bandgaps less than 3.0 eV. A variety of bismuth-based semiconductors have been tested as photocatalysts, e.g., Bi2O3, Bi2MO6 (where M ¼ Cr, Mo, and W), BiVO4, BiPO4, (BiO)2CO3, BiOX (where X ¼ Cl, Br, and I), and pentavalent bismuthates (NaBiO3, SrBi2O6, and BaBi2O6) (Fig. 4.1).

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With regard to environmental applications, bismuth-based semiconductors have been used for: • Photocatalytic oxidation of organic and reduction of inorganic toxic or refractory pollutants in water into nontoxic products without secondary pollutants • Generation of hydrogen—a clean alternative fuel—by water splitting • NOx oxidation in the gaseous phase under solar light • CO2 reduction into useful organics in mild conditions Photocatalysis using bismuth-based materials has been considered a potential method of organic pollutant removal from wastewater such as pharmaceuticals (Zeng et al. 2017; Ding et al. 2017; Akbarzadeh et al. 2018), dyes (Di et al. 2016b; Liao et al. 2017; Ao et al. 2016), industrial pollutants including phenol (Liu and Wang 2016), isoproturon (Xie et al. 2015), bisphenol A (Di et al. 2016b; Wen et al. 2017), parabens (Xiao et al. 2017), methanol (Zheng et al. 2017), and Cr (IV) (Xu et al. 2017; Ma et al. 2010) under visible or solar light. Antibiotics and painkillers are the most common pharmaceuticals found in aquatic environments all over the world. Moreover, antibiotics are recognized as one of the serious emerging issues in environmental chemistry because their antimicrobial nature prevents effective removal of them in sewage treatment processes (Li and Shi 2016). In this chapter, current Bi-based photocatalysts and their applications in environmental protection are discussed. Table 4.1 summarizes the main features of studies on photocatalyzed degradation of pharmaceuticals in aqueous solutions published between 2010 and 2018.

4.2 4.2.1

Photocatalytic Oxidation of Pharmaceuticals in Water Tetracycline

Tetracycline is an antibiotic commonly used to treat a number of infections. Because it is widely consumed all over the world, its concentrations in surface water and treated wastewater are significant. Carbon quantum dots/BiOI hollow microsphere structures, prepared in the presence of the ionic liquid 1-butyl-3-methylimidazolium iodide at room temperature, were applied for the degradation of tetracycline in an initial concentration of 20 mg L1 under visible light irradiation (λ > 580 nm) and near-infrared light irradiation (λ > 700 nm). The efficiency of industrial micropollutant degradation (bisphenol A and phenol in an initial concentration of 10 mg L1) by the as-prepared photocatalyst was also tested. The introduction of 3% carbon quantum dots to the carbon quantum dots/BiOI material induced strengthening of the light absorption within the full spectrum, enhanced the separation efficiency of the photogenerated electronhole (e/h+) pairs, and increased the specific surface areas of the photocatalyst. After 120 min of visible light irradiation, the rates of photodegradation of tetracycline and bisphenol A by the 3-wt% carbon quantum

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Table 4.1 Studies on pharmaceutical and bisphenol A photocatalytic removal with Bi-based materials Structure of pharmaceutical

Tetracycline

Type of photocatalyst (N-doped carbon quantum dot)/BiOBr ultrathin nanosheets BiOBr 50% NiFe2O4/Bi2O3

Bi/Bi2S3/graphitic carbon nitrate Pillared graphitic carbon nitrate plasmonic semiconductor Bi/α-Bi2O3/g-C3N4

3% carbon quantum dot/BiOI hollow microspheres BiOI 22.12% SrTiO3/BiOI heterojunction photocatalyst Bi24O31Br10 hierarchical nanoflakes

BiYO3 nanorods

BiVO4 rod-like structure

BiVO4--palygorskite

Experimental conditions and removal efficiency Dose of photocatalyst 1 g L1 C0(drug) 20 mg L1 67.1% in 120 min 38.6% in 120 min Dose of photocatalyst 1 g L1 C0(drug) 10 mg L1 90.78% in 90 min Dose of photocatalyst 1 g L1 C0(drug) 10 mg L1 90% in 150 min Dose of photocatalyst 1 g L1 C0(drug) 10 mg L1 90.2% in 180 min TOC removal 55.6% Dose of photocatalyst 0.5 g L1 C0(drug) 10 mg L1 59% in 120 min 50% in 129 min Dose of photocatalyst 1 g L1 C0(drug) 20 mg L1 85% in 90 min Dose of photocatalyst 1.5 g L1 C0(drug) 40 mg L1 94% in 60 min Dose of photocatalyst 0.5 g L1 C0(drug) 10 mg L1 70% in 3 h Dose of photocatalyst 1 g L1 C0(drug) 20 mg L1 t½ 12 min Dose of photocatalyst 1 g L1 C0(drug) 30 mg L1 91% in 240 min

References Di et al. (2016a)

Ren et al. (2014)

Chen et al. (2017)

Chen et al. (2018a)

Di et al. (2016b)

Wen et al. (2017)

Xiao et al. (2013)

Wu et al. (2015)

SánchezMartínez et al. (2015) Shi et al. (2017b)

(continued)

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Table 4.1 (continued) Structure of pharmaceutical

Type of photocatalyst Ag/AgCl/Bi4Ti3O12 nanosheet with exposed {001} facet photocatalyst Bi24O31Br10 High percentage of {001} facets on surface

Ciprofloxacin

2% BiOI/(Fe) with Z-scheme mechanism

PO4-doped Bi2WO6

Gatifloxacin

Danofloxacin mesylate Ciprofloxacin

Experimental conditions and removal efficiency Dose of photocatalyst 1 g L1 C0(drug) 5 mg L1 75% in 120 min Dose of photocatalyst 0.6 g L1 C0(drug) 20 mg L1 94.8% ciprofloxacin in 180 min Dose of photocatalyst 0.2 g L1 C0(drug) 10 mg L1 About 80% in 4.5 h Dose of photocatalyst 1 g L1 C0(drug) 20 mg L1 Gatifloxacin: about 49% in 140 min Danofloxacin mesylate: about 50% in 140 min Ciprofloxacin: about 60% in 140 min Enrofloxacin hydrochloride: about 55% in 140 min

References Shi et al. (2017a)

Zeng et al. (2017)

Islam et al. (2017)

Li et al. (2016a)

Enrofloxacin hydrochloride 5% Bi2O3/TiO2 10% Bi2O3/TiO2

Dose of photocatalyst 0.5 g L1 C0(drug) 25 mg L1 60% in 120 min 92% in 120 min

Sood et al. (2016)

Ag/AgCl/Bi4Ti3O12 nanosheet with exposed {001} facet photocatalyst

Dose of photocatalyst 1 g L1 C0(drug) 5 mg L1 80% in 60 min 75% mineralization in 120 min Dose of photocatalyst 0.5 g L1 C0(drug) 20 mmol L1 57% in 30 min at the dark

Shi et al. (2017a)

Ofloxacin

Carbamazepine Bi3+ self-doped NaBiO3 nanosheets

Ding et al. (2017)

(continued)

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Table 4.1 (continued) Structure of pharmaceutical

Type of photocatalyst

1% graphene quantum dots/BiVO4

BiOBr

Ibuprofen

Fe3O4/Bi2WO6

Graphitic carbon nitrate/ Ag/AgCl/BiVO4

Graphitic carbon nitrate/ Bi2WO6

Experimental conditions and removal efficiency stage 99.8% in 60 min at the light stage Monohydroxy carbamazepine and ketocarbamazepine. were identified by LC–MS and GC–MS as primary intermediates of the carbamazepine oxidation process Dose of photocatalyst 0.2 g L1 C0(drug) 10 mmol L1 95% in 180 min Dose of photocatalyst 0.2–2.0 g L1 C0(drug) 5  103– 1  101 mmol L1 80% after 2 h 63% TOC removal; mineralization of ibuprofen by the photocatalysis process was weak (about 11%); the main process was adsorption Dose of photocatalyst 1.4 g L1 C0(drug) 10 mg L1 pH 4.7 80% in 120 min Dose of photocatalyst 0.25 g L1 C0(drug) 10 mg L1 94.7% in 1 h Dose of photocatalyst 0.2 g L1 C0(drug) 25 μM 96.1%, 75.3%, and 38.2% were achieved after 1 h in Ibuprofen/ deionized water, tap water, and river water, respectively Ibuprofen removal rates in the presence of 5 mM of electrolytes decreased in the order NaCl > Na2SO4 > NaHCO3

References

Tang et al. (2017)

Li et al. (2016b)

Bastami et al. (2017)

Akbarzadeh et al. (2018)

Wang et al. (2017b)

(continued)

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Table 4.1 (continued) Structure of pharmaceutical

Type of photocatalyst Carbon quantum dots/ BiOCOOH

Diclofenac

5 mol% Ag/Bi2WO6/ 2.5 wt% graphene Metoprolol

BiOCl/graphitic carbon nitrate/Cu2O/Fe3O4

Sulfamethoxazole

BiOCl BiOI

17 α-Ethinyl estradiol

Experimental conditions and removal efficiency Dose of photocatalyst 0.6 g L1 C0(drug) 4 mg L1 100% in 60 min 70% TOC removal in 120 min 94% dechlorination in 120 min About 60% deamination in 120 min Dose of photocatalyst 1 g L1 C0(drug) 10 mg L1 100% in 2 h The intermediates at m/z 284 and 300 were identified to be mono- and dihydroxylated products of metoprolol The dominant intermediate at m/z 238 was formed by loss of the ether group and H-abstraction, due to attack of the •OH radical on the alkyl group and O-atom addition Dose of photocatalyst 0.2 g L1 C0(drug) 100 μM 99.5% in 60 min with visible light 41.6% TOC removal in 3h 92.1% in 120 min in natural sunlight •OH radicals attacked the benzene ring, forming mono- or dihydroxylated sulfamethoxazole, or attacked the isoxazole ring, leading to cleavage of the S–N bond Dose of photocatalyst 0.5 g L1 C0(drug) 10 mg L1 100% (BiOI) and 92% (BiOCl) in 30 min with 360 nm UV light

References Chen et al. (2018b)

Yu et al. (2015a, b)

Kumar et al. (2018)

Ahern et al. (2015)

(continued)

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Table 4.1 (continued) Structure of pharmaceutical

Estradiol

Type of photocatalyst C/N-doped β-Bi2O3 nanosheets named DBO-300 (with N 1.883 wt%, C 2.579 wt %) and DBO-350 (with N 0.913 wt%, C 0.512 wt %) Bi2WO6

Norfloxacin

Bi2WO6

3 wt% carbon quantum dots/BiOBr

Bisphenol A

Bi/Bi4O5I2

NaBiO3 in HCl solution

Experimental conditions and removal efficiency Dose of photocatalyst 0.5 g L1 C0(drug) 10 mg L1 95% in 2 min under DBO-300 98.9% in 20 min under DBO-350 Dose of photocatalyst 1 g L1 C0(drug) 20 mg L1 at initial pH 9, 90% in 120 min The concentration increased from 10 to 50 mg/L, the adsorption percentage on the catalyst decreased from 55% to 2%, and the final photocatalytic degradation rate decreased from 95% to 10% Addition of H2O2 (10 mM): 1.95 times the reaction rate without use of H2O2 Dose of photocatalyst 1 g L1 C0(drug) 0.0313 mM, 90% in 20 min C0(drug) 0.1565 mM, 420 nm) irradiation. After 180 min, 94.8% of ciprofloxacin could be photodegraded in the presence of this sample. The study suggested that the adsorption modes between the ciprofloxacin species and the surface of the Bi24O31Br10 were both monodentate coordination and electrostatic interaction (Zeng et al. 2017). Ciprofloxacin was removed from water solutions by a BiOI/iron-based metal– organic framework MIL-88B(Fe) composite, which was successfully synthesized through a simple precipitation method. Degradation experiments using ciprofloxacin and another pollutant, phenol, confirmed that the best photocatalytic activity was exhibited by the BiOI/MIL-88B(Fe) hybrid nanocomposite with 2 w% BiOI. The results showed that the BiOI/iron-based metal–organic framework MIL-88B had improved charge carrier separation and a reduced recombination process via a simple Z-scheme mechanism. Moreover, the photocatalytic degradation rate for ciprofloxacin was significantly higher than that observed for phenol decomposition (Islam et al. 2017). Li et al. (2016b) applied a PO4-doped Bi2WO6 photocatalyst (1 g L1) to remove a group of antibiotics (20 mg L1) including gatifloxacin, danofloxacin mesylate, ciprofloxacin, and enrofloxacin hydrochloride. The photocatalyst was prepared by a urea precipitation method in a hydrothermal process. The degradation rates of the antibiotics with use of the PO4-doped Bi2WO6 sample were all greater than those with use of a pristine Bi2WO6 sample, over the same reaction time of 140 min: 50%, 32%, 55%, and 60% degradation rates were achieved for gatifloxacin, danofloxacin mesylate, enrofloxacin hydrochloride, and ciprofloxacin, respectively. The main activated species that participated in the antibiotic degradation were h+ and O2•. The doping effect of the PO4 group in the Bi2WO6 influenced the energy band structure by lifting up and moving down the conduction band position and the valence band position of the Bi2WO6 sample, respectively. The light absorption

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property and separation efficiency of the charge carriers were also elevated (Li et al. 2016a). Ofloxacin removal was examined in the presence of Bi2O3/TiO2 under visible and solar light irradiation. The p–n heterojunction was composed of crystalline anatase TiO2 and monoclinic α-Bi2O3. There was superior catalytic activity with a 10% Bi2O3/TiO2 heterostructure with a relatively small size and high surface area of the catalyst. Degradation of 92% of ofloxacin in the presence of 10% Bi2O3/TiO2 was achieved in 120 min, and the process was more rapid than that in the presence of commercially available photocatalysts such as TiO2 P25 (mixed phase of anatase/ rutile; Degussa), PC 50, and PC 500 (pure-phase anatase TiO2; Millennium PC). A plausible mechanism for the photocatalytic degradation reaction taking place on the Bi2O3/TiO2 heterojunction showed that the photogenerated electron could easily migrate from TiO2 to Bi2O3 as a result of the interface potential difference. This type of phenomenon causes efficient charge separation, thereby leading to an effective reduction in the recombination of charge carriers (Sood et al. 2016).

4.2.3

Carbamazepine

Carbamazepine is an anticonvulsant used to treat epilepsy, nerve pain, and diabetic neuropathy. Because it is widely consumed, it has been detected in wastewater treated in many countries and water bodies receiving effluent. Photocatalytic decomposition of a carbamazepine and tetracycline hydrochloride aqueous solution of 5 mg L1 was studied by Shi et al. (2017a). A Bi4Ti3O12 nanosheet with an exposed {001} facet photocatalyst decorated with ethanol > methanol. Bi4Ti2.6Cr0.4O12 achieved a hydrogen evolution rate 2.85 times that of a Bi4Ti3O12 nanosheet (Chen et al. 2016a). Bi2O3 clusters immobilized on anatase TiO2 nanostructures by a wet impregnation method were tested for H2 evolution under natural solar light, UV-LED [lightemitting diode] (365  5 nm), and visible LED (420  5 nm) light sources. Catalyst H2 production of 26.02 mmol h1 g1 achieved by 0.4% Bi2O3/TiO2 showed that this heterostructure was remarkably efficient for H2 production under solar light. Moreover, this performance was reproducible over five cycles (Reddy et al. 2017).

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4.7

113

Conclusions

The presence of pharmaceuticals and other micropollutants in municipal wastewater and the consequences of their presence in aquatic environments have received considerable attention from the scientific community. Photocatalytic processes are used in the most advanced technologies, and it is likely that they will be applied in the future on an industrial scale, especially for wastewater containing relatively low concentrations of pollutants or for air purification. However, photocatalytic methods that require ultraviolet light are not being applied, because of their expense. Therefore, visible light–activated Bi-based photocatalysts are a promising alternative to ultraviolet light–activated photocatalysts such as TiO2. This chapter has summarized the progress achieved with novel Bi-based photocatalysts applied to environmental protection including water treatment, air purification, and water splitting. Bi-based photocatalysts with p–n heterojunctions, with Z-scheme mechanisms, modified with carbon quantum dots or Fe2O3 nanoparticles, or doped with N are now well accepted as promising visible light photocatalysts due to their excellent activity and mineralization ability. In some cases, they have shown much higher activity than P25 (TiO2). On the basis of the collected research reports from the period 2012–2018, the following conclusions can be drawn: 1. The concentrations of eliminated pharmaceuticals and other micropollutants have mainly been in the range of 5–30 mg L1. Generally, most investigations have dealt with micropollutant concentrations higher than those found in municipal wastewater. A concentration of 10 mg L1, which has mainly been used by different research groups, allows comparison of activity by differently synthesized photocatalysts. The efficiency of micropollutant oxidation and mineralization of the studied photocatalysts has been high (between 50% and 99%), but transformation products of the contaminants have rarely been detected and identified. 2. The activity of synthesized photocatalysts under visible and solar light has been the most investigated topic, while activity in light-emitting diode light or natural sunlight has been the least studied subject. 3. In many studies, a single pollutant solution in distilled water has been used to evaluate the efficiency of its photocatalytic removal; the influences of other organic and inorganic pollutants that naturally exist in the water on the efficiency and the mechanism of elimination of the studied pollutant have rarely been investigated. 4. Photocatalytic processes using Bi-based catalysts have been successfully applied at the laboratory scale but not investigated in pilot-scale removal of micropollutants from water or air. Acknowledgements The author would like to acknowledge financial support received from the Polish Ministry of 406 Science and Higher Education under grant number DS 530-8626-D596-191F.

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Chapter 5

Phosphors-Based Photocatalysts for Wastewater Treatment Olga Sacco, Vincenzo Vaiano, and Diana Sannino

Contents 5.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2 Phosphor Materials: A Historical Background . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.3 Inorganic Phosphors in Photocatalysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.3.1 Types of Inorganic Phosphor Materials . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.3.2 Down-Conversion Phosphors in Photocatalysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.3.3 Up-Conversion Phosphors in Photocatalysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.3.4 Long-Persistent Phosphors in Photocatalysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.4 Organic Up-Conversion Phosphors in Photocatalysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract Every day a large amount of products has been released by chemical and pharmaceutical industries threatening the environment and human health. Moreover, their removal using conventional oxidation methods is difficult because a lot of pollutants are biorecalcitrant. Photocatalysis, an advanced oxidation technology, appears one of the most viable solutions due to its ability to oxidize a wide range of toxic organic compounds into harmless compounds such as CO2 and H2O by irradiation with UV light. Organic pollutants can be removed from water by a UV-driven photocatalytic process involving nanoparticles with semiconducting properties. However, their use in a photoreaction system suffers of the disadvantage due to the no-uniform photon distribution inside the reactor core. The chapter describes a possible solution for enhancing the photon distribution inside the photoreactors, using inorganic and organic light-emitting particles (phosphors) coupled with photocatalysts to be applied in water and wastewater treatment. The chapter also underlines the difference between inorganic particles having downconversion, up-conversion, and long-afterglow luminescence properties. Additionally, the use of up-conversion organic phosphors has been proposed. Finally, some examples concerning the use of semiconductors coupled with different O. Sacco · V. Vaiano (*) · D. Sannino Department of Industrial Engineering, University of Salerno, Fisciano, SA, Italy e-mail: [email protected] © Springer Nature Switzerland AG 2020 Inamuddin et al. (eds.), Nanophotocatalysis and Environmental Applications, Environmental Chemistry for a Sustainable World 30, https://doi.org/10.1007/978-3-030-12619-3_5

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photoluminescent materials in the removal of pollutants from water and wastewater are presented. Keywords Photocatalysis · Wastewater treatment · Phosphors · Down-conversion phosphors · Up-conversion phosphors · Long-afterglow phosphors · Organic up-conversion phosphors · Coupling with photocatalysts · Doped photocatalysts · UV and visible light

5.1

Introduction

In the last years, due to the growth of population, the use and reuse of water have become a great concern (Jury and Vaux 2005), leading to the necessity of developing cheap and effective technologies for water and wastewater treatment (Maliva and Missimer 2012). The processes currently used for water depollution are mainly based on chemical-physical and biological processes. However, they are not able to remove not biodegradable organic pollutants (Naidoo and Olaniran 2014), classified as bio-recalcitrant (Samer 2015). For this reason, heterogeneous photocatalysis may be a valid alternative technology able to remove pollutants hardly oxidizable by conventional processes (Herrmann et al. 1993). Heterogeneous photocatalysis is based on the use of the energy associated with a light source to activate a catalyst that must be a semiconductor (Ciambelli et al. 2017). In a semiconductor material, called “photocatalyst,” the conduction band is separated by the valence band with an energy gap called “bandgap energy.” When the photocatalyst is irradiated with a light having energy equal to or greater than the bandgap energy, electrons are excited from the valence to the conduction band. Simultaneously, positive holes are generated in the valence band. This phenomenon implies the generation of electron (e
) and hole (h+) pairs (Ibhadon and Fitzpatrick 2013). These species may recombine, releasing heat, or promote reduction and oxidation reactions by the interaction with species adsorbed on the photocatalyst surface (Hoffmann et al. 1995). The electron in the conduction band can interact with electronic acceptors, such as adsorbed O2, producing a superoxide radical anion, whereas the photoinduced holes are captured by electronic donors, such as organic molecules oxidizing them (Al-Ekabi and Serpone 1988). Taking into account the above considerations, a photocatalyst must have appropriate positions of valence and conduction bands in a manner to activate simultaneously both oxidation and reduction reactions (Liu et al. 2013a). The most commonly used photocatalysts to oxidize a great variety of toxic organic compounds to CO2 and H2O are titanium dioxide (TiO2) and zinc oxide (ZnO) (Herrmann 1999; Sacco et al. 2018; Sunandan et al. 2012; Vaiano et al. 2017a). It must be considered that a photocatalytic process is based on the use of a photoreactor, which is a specific device able to realize an optimal interaction between photons, chemical reactants, and the photocatalyst (Spasiano et al. 2015). The photoreactor design is significantly different from classic chemical reactors

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since their geometry must be designed in order to assure that the photons emitted by light sources are efficiently collected (Vaiano et al. 2015b). For this reason, the typical operating parameters, such as pressure, temperature, and fluid dynamics conditions, could be less important in order to achieve an optimal photocatalytic process operation (Van Gerven et al. 2007). Another key aspect related to the design of a photoreactor is its optical-path length that must ensure that the photon flux is uniform everywhere inside the reactor (Mukherjee and Ray 1999; Ray and Beenackers 1998; Spasiano et al. 2015). This last aspect is very important especially if it is considered that, as the light travels through the photoreactor, the photons are scattered and absorbed by the photocatalyst particles and its intensity decays; since photons initiate the photoreaction, this results in a nonuniform reaction rate in the photoreactor (Motegh et al. 2014). Therefore, as pointed above, the key aspect to consider is the photon distribution in the core of the photoreactor that should be uniform to irradiate the photocatalysts in an effective way (Palma et al. 2010; Sacco et al. 2015a). The drawback associated with the light attenuation inside the photoreactors could be overcome with the use of light-emitting phosphorescent particles (generally called “phosphors”) coupled with photocatalysts (Ciambelli et al. 2011; Murcia et al. 2012; Palma et al. 2010; Sacco et al. 2015a, b, 2017; Sannino et al. 2013; Vaiano et al. 2015a, c, 2016, 2017b). In this case, the photoreactivity can be enhanced because suitable phosphors introduced into the system were able to transform external light radiation into emission, able to photoexcite the fraction of photocatalyst particles in the core reactor volume, otherwise screened by the photocatalyst itself (Sacco et al. 2015a, b). In recent years the use of phosphors has attracted tremendous interest owing to their special electronic structure and unique photophysical properties, especially with respect to the generation and amplification of light (Liu et al. 2015a).

5.2

Phosphor Materials: A Historical Background

Natural inorganic phosphors have been known during the tenth century both in China and Japan and in Europe during the end of the middle ages (Lin et al. 2016). Later, during the seventeenth century, the stone of Bologna attracted the interest of Galileo Galilei. The stone of Bologna is a barite mineral emitting yellow to orange light with long persistence when excited by the direct sunlight. In 1671, Kirchner discovered that the luminescence of the stone of Bologna was due to BaS impurity (Smet et al. 2010). During that historical period, phosphors were mainly used for decorative purposes (Feldmann et al. 2003). In the following centuries, many scientists synthesized and investigated phosphor materials (such as SrS- and CaS-based phosphors) (Virk 2015). However, the luminescent properties of ZnS (one of the most important luminescent hosts in the twentieth century) were not

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recognized until 1866, when the “Sidot blend” (i.e., hexagonal ZnS) was discovered and developed by Theodor Sidot (Valeur and Berberan-Santos 2011). It is very important to mention the work made in 1888 by Eilhard Wiedemann who was the first scientist to classify phosphors in different classes according to the type of excitation (Valeur and Berberan-Santos 2011). Together with the development and classification of inorganic phosphors, there was also the discovery of bioluminescence phenomenon generated by organic phosphor materials (Chen et al. 2011; Harvey 1919). Compared to inorganic phosphors, organic ones are interesting because of their high luminescence efficiency.

5.3 5.3.1

Inorganic Phosphors in Photocatalysis Types of Inorganic Phosphor Materials

A phosphor material is a substance that exhibits the phenomenon of luminescence. In general, phosphors are based on lanthanide-doped materials able to emit light at a specified wavelength when excited by an electromagnetic radiation (Werts 2005). In particular, the key to enabling precisely tuned color emission is governed by the 4f-5d transitions of divalent and trivalent lanthanides (Qin et al. 2017). Phosphors include down-conversion phosphors and up-conversion phosphors. Down-conversion phosphors transform higher-energy photons into photons with lower energy, whereas up-conversion phosphors are able to transform lower-energy photons into higher-energy photons. Besides the previous types of phosphor materials, long-afterglow phosphors can emit long-lasting phosphorescence that can allow to the photocatalyst to maintain its photocatalytic activity also after turning off the external irradiation light (Li et al. 2016).

5.3.2

Down-Conversion Phosphors in Photocatalysis

As pointed above, down-conversion phosphors can absorb the high-energy photons and emit the low-energy photons at a longer wavelength. Many transition metal ions (Eu2+, Er3+, Pr3+, Ce3+, Nd3+) show emission at different wavelengths in different host lattices (Ronda 2007). This phenomenon allows changing, within certain limits, the emission wavelength. In this way, the emission spectra (and excitation spectra) can be tuned toward the specifications required (Ronda 2007; Tamrakar et al. 2015, 2016). An example of excitation and emission mechanism for a typical down-conversion phosphor material (Gd2O3: Er3+) is reported in Fig. 5.1. The cooperative energy transfer between Gd3+ and Er3+ generates the down-conversion process (Fig. 5.1). The external radiation at 375 nm provided to the phosphors is absorbed by Gd3+ that is excited to the first state 6P7/2. The energy from 6P7/2 state of Gd3+ to the 4P3/2 state

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Fig. 5.1 Schematic energy-level diagram showing the excitation and emission mechanism of down-conversion phosphors (Gd2O3: Er3+). (Tamrakar et al. 2015); UV: ultraviolet light, vis: visible light

of the Er3+ ion populates the Er3+lower-energy levels by non-radiative decay. These populated energy levels are responsible for radiative decay and generate emission bands in visible regions due to the following transitions: 4F7/2 ! 4I15/2, 2H11/2/4S3/2 ! 4I15/2, and 4F9/2 ! 4I15/2 (Tamrakar et al. 2016). The most commercial phosphors, such as ZnS-based phosphors (ZSP) (Sacco et al. 2015a, b) and BaMgAl10O17: Eu (BaMgP) (Lu et al. 2006), emit blue light when excited by UV light (Vaiano et al. 2017b). The most literature about the photocatalysts/down-conversion phosphors reports that the down-conversion phosphors can absorb the high-energy photon and emit the low-energy photon at a longer wavelength, which can be easily absorbed by given dye and thus effectively excite the dye to generate more electron-hole pairs, resulting in the improvement of the self-sensitized degradation of dye (Liu et al. 2013a, b, 2015b), but in this case the photocatalyst coupled with down-conversion phosphors was not doped, so it was not able to absorb the visible light commonly emitted by the down-conversion phosphors. The way to better use the visible light of downconversion phosphors is to couple the phosphors with the doped photocatalyst. In order to extend the absorption ability of photocatalyst to the visible-light region, many strategies, such as doping with heteroatoms, have been developed (Vaiano et al. 2017b). The most commonly studied doped photocatalyst is N-doped TiO2 (N-TiO2) that is able to be excited by the visible light coming from down-conversion phosphor, thanks to its bandgap energy of about 2.5 eV (Sacco et al. 2012). By coupling down-conversion phosphors with N-TiO2, it is possible to enhance the photon transfer inside the reactor, according to the following simplified mechanism (Fig. 5.2): down-conversion phosphors can catch the UV fraction of solar light

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Fig. 5.2 Schematic energy transfer mechanism accounting for enhanced photocatalytic processes in doped semiconductor catalysts (N-doped TiO2) coupling with down-conversion phosphors. (Sacco et al. 2015a, b); N-TiO2 N-doped TiO2, DC-phosphors down-conversion phosphors

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Fig. 5.3 Behavior of atrazine experimental concentration as a function of UV irradiation time. (Sacco et al. 2015a, b); N-TiO2 : N-doped TiO2

and emit their specific visible light in the core of photoreactor, transferring additional photons to the photocatalyst in its close proximity (Sacco et al. 2015a, b). This system was proven effective in the removal of several organic compounds from water and wastewater (Sacco et al. 2015a, b; Vaiano et al. 2017b). As an example, Fig. 5.3 shows the experimental behavior of atrazine relative concentration as a function of irradiation time for N-TiO2 supported on ZnS-based

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Fig. 5.4 Effect of N-doped TiO2/down-conversion phosphors on methylene blue, phenol, and terephthalic acid degradation after 60 min of UV irradiation. (Vaiano et al. 2017b)

down-conversion phosphors (Sacco et al. 2015a, b). The experimental results showed an enhancement of photocatalytic performances in comparison to pure N-TiO2 and bare down-conversion phosphors. The photocatalytic performances using N-TiO2/down-conversion phosphors were also enhanced in the degradation of methylene blue, phenol, and terephthalic azid (Fig. 5.4) (Vaiano et al. 2017b).

5.3.3

Up-Conversion Phosphors in Photocatalysis

An attractive issue in the field of photocatalysis is to increase the photocatalysts’ absorption properties from UV to visible light (Tong et al. 2012). To overcome this limitation, up-conversion phosphors have been considered as one of the most promising solutions for overcoming the drawback related to the fact that TiO2 photocatalyst absorbs only 5% of the solar light impinging on the Earth’s surface (Mekprasart and Pecharapa 2011). Typically, the up-conversion phosphors are composed of a host lattice and an activator. In addition, sometimes a sensitizer is also required (Vuojola and Soukka 2014). Usually, the host material is a combination of optically inert cations and anion, whereas the activators are generally optically active cations. Figures 5.5, 5.6 and 5.7 schematize the three simplest forms of up-conversion emission occurring in inorganic up-conversion phosphors (Li et al. 2010; Shi et al. 2011; Zhang et al. 2010).

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Fig. 5.5 Diagram of excited-state absorption mechanism. (Wang and Liu 2009); ω: energy, E0 : ground state, E1 : intermediate state, E2 : excited state

Fig. 5.6 Diagram of photon avalanche mechanism. (Chen and Zhao 2012); ω: energy, E0 : ground state, E1 : intermediate state, E2 : intermediate state, Ex : excited state, a : energy transfer

Fig. 5.7 Diagram of energy transfer up-conversion mechanism. (Li et al. 2009); S: sensitizer, A: activator

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Three basic mechanisms can be recognized: excited-state absorption, photon avalanche, and energy transfer up-conversion (Demas 2011). The excited-state absorption mechanism (Fig. 5.5) involves sequential absorption of two photons by an ion or electron using an intermediate excited state. When ions or electrons at the ground state (named E0) absorb photons, they are first excited to the intermediate state (E1). After the sequential absorption of the second photon, the electrons or ions can jump to another excited state (E2) with the emission of photons at higher energy (ω* > ω1, ω2) when they come back to the ground state (Chan 2015). The excitation should be realized by tunable solid-state lasers (Liu et al. 2017). Photon avalanche mechanism is characterized by a more complicated process since it is generally characterized by three nonlinear behaviors that are transmission, emission, and rise time on the pump power intensity with the general presence of a critical pump threshold (Gonzalez-Bejar and Perez-Prieto 2015). An electron or ion is excited, thanks to the absorption of a suitable excitation radiation (named ω1) (Fig. 5.6). The excitation radiation is typically a little higher than excited state (E2)(a), and it is not resonant with the absorption transition from ground state (E0) to the intermediate states. It returns to the excited state (E2) by means of cross-relaxation phenomena. A transfer of energy happens between the excited state (E2) and the ground-state (E0) electron, inducing the formation of two electrons in the intermediate states (E1) a1 and a2 (energy transfer I), respectively. One of the two electrons is able to absorb the excitation radiation, and, consequently, it is excited to Ex state, in which it interacts with ground-state (E0) electrons and three excited-state (E1) electrons are formed as a consequence of energy transfer II. At this situation, the absorption transition from intermediate state (E1) to excited state (Ex) is resonant with the excitation radiation. The repetition of the previously described steps induces the increase of the number of electrons present in the excited state (Ex). When the electrons return to the ground state (E0), there is the emission of photons with high energy. In conclusion, photon avalanche process is based on the involvement of resonant excited-state absorption, cross relaxation, and population of the reservoir level, leading to up-conversion emission. It is worthwhile to note that, despite the photon avalanche process is very efficient, it presents some disadvantages, such as the limited maximum output caused by the scarce ground-state E0 absorption and the needing of high pump powers necessary to reach the threshold condition (Liu et al. 2015a). Energy transfer up-conversion is the most efficient up-conversion process occurring in rare earth-doped nanomaterials because it can be realized by traditional light sources (such as lamps) (Auzel 2004). In the case of sensitizer and activator which are excited by energies almost equal and the distance between sensitizer and activator is close enough, the energy is transferred from sensitizer to activator, inducing the excitation of activator from its ground state to an excited state before that the sensitizer emits photons (Auzel 2004; Wang and Liu 2009) (Fig. 5.7). Up-conversion phosphors with energy transfer up-conversion mechanism can be realized using trivalent rare earth ions (such as Er3+, Tm3+, Ho3+, Tb3+, Pr3+, Nd3+, and Eu3+) in different host materials (Chen et al. 2015). For example, the NaYF4: Yb3+ Er3+ system can emit red, blue, and green light through the process schematized in Fig. 5.7. The Yb3+ ion, together with its excited

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Fig. 5.8 Schematic energy transfer mechanism accounting for enhanced photocatalytic processes in semiconductor catalysts coupling with up-conversion phosphors. (Liu et al. 2015a); UC phosphors: up-conversion phosphors, NIR: near-infrared radiation, UV: ultraviolet radiation, vis: visible radiation

state 2F5/2, possesses an energy similar to 4I11/2 (Er3+), and therefore it can act as a sensitizer for Er3+ (activator) and be able to transfer its energy to an unexcited of Er3+ ion by means of the following energy transfer process: 2F5/2 (Yb3+) þ 4 I15/2 (Er3+) ! 2F7/2 (Yb3+) þ 4I11/2 (Er3+) (Gainer et al. 2011). Through further phonon-assisted and cross relaxation and processes, a red (~654 nm) and green (~526 nm and ~533 nm) emission can be achieved (Auzel 2004). In order to extend the limited optical absorption edge of TiO2, up-conversion phosphors able to transform near-infrared radiation (NIR) photons to ultraviolet radiation (UV) and visible radiation (vis) photons (Fig. 5.8) have been presented as a promising solution for overcoming the problems related to the large-scale application of photocatalysis for water depollution (Gao et al. 2017; Huang et al. 2013; Li et al. 2010, 2011; Tang et al. 2013; Wang et al. 2014; Wu et al. 2014; Zhang et al. 2012). Although numerous studies have been reported on the coupling of up-conversion phosphors with photocatalysts, the use of these types of phosphor materials requires a high-incident light intensity (10 to 104 W/cm2) (Li et al. 2010; Shi et al. 2012; van der Ende et al. 2009). Moreover, inorganic up-conversion phosphor materials, typically, can be excited effectively only by electromagnetic radiation at two defined wavelengths (800 and 980 nm) (Ye et al. 2017). Therefore, it must be considered that the use of up-conversion phosphors coupled with semiconductors must require laser irradiation or concentrated sunlight to effectively activate the up-conversion phosphors, clearly hindering the practical implementation of up-conversion phosphor-based photocatalytic system aimed to wastewater treatment (Liu et al. 2015a).

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Long-Persistent Phosphors in Photocatalysis

In addition to down-conversion phosphors and up-conversion phosphors, in order to further improve the energy efficiency of the process, photocatalysts were recently coupled with long-persistent phosphors (long-afterglow phosphors), such as CaAl2O4: Eu2+,Nd3+ (Aitasalo et al. 2003), Sr2.90Eu0.03Dy0.07Al4Si11, SrAl2O4: Eu2+,Dy3+ (Matsuzawa et al. 1996), and BaAl2O4:Eu2+,Dy3+ (Lei et al. 2010), whose emission can excite the photocatalyst and maintain its activity even after turning off the light sources external to the photoreactor. This happens because long-afterglow phosphors store energy from the external light source and emit it slowly, also when the irradiation source is cut off (Li et al. 2009). As an example, in order to explain the emission mechanism of SrAl2O4: Eu2+, Dy3+ possessing long-persistent emission (Fig. 5.9), Matsuzawa et al. (1996) proposed that, when the Eu2+ ion is excited by incident photons, a hole escapes to the valence band, promoting the generation of Eu+ ion. The generated hole is then captured by a Dy3+, thus generating a Dy4+ ion. Moreover, the release of thermal energy causes the trapped hole to be released into the valence band again. From there the hole moves back to a Eu+ ion, allowing europium ions to return to the Eu2+ ground state with the simultaneous emission of a photon (Matsuzawa et al. 1996). The scientific literature reports that the use of long-afterglow phosphors as support for photocatalysts allowed to maintain the photocatalytic activity even in the darkness. More in detail, long-afterglow phosphors are able to be charged with either natural or artificial light sources (den Eeckhout et al. 2010) and, by means of the light emission in dark conditions, provide the necessary photons for the

Fig. 5.9 Schematic energy-level diagram showing the excitation and emission mechanism of longafterglow phosphors (SrAl2O4: Eu2+, Dy3+). (Matsuzawa et al. 1996)

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Fig. 5.10 Photocatalytic degradation of (a) atrazine and (b) oxytetracycline under the visible light irradiation using long-afterglow phosphors (Strontium aluminate based) as support for Fe-N-doped TiO2 photocatalyst. (Sacco et al. 2015a, b); LAPS : long-afterglow phosphors, Fe-NTiO2 : ironnitrogen-doped TiO2, Fe-NTiO2/LAPS : iron-nitrogen-doped TiO2/long-afterglow phosphors

excitation of the photocatalyst. It is worthwhile to note that the capability of the photocatalytic process to work also in dark conditions can allow a significant reduction of the electric energy consumption related to the photocatalytic reactor. In 2003, Ko (Ko and Ko 2003) studied the use of long-afterglow phosphors as a light source for TiO2-based photocatalysts, showing the efficiency in the ammonia degradation even if the catalyst was maintained in dark conditions. Starting from literature papers dealing with the use of long-afterglow phosphors in photocatalytic reactions, in 2004, Zhang et al. (2004) proposed the concept of energy storage photocatalysis materials. In 2012, Li et al. (2012b) developed a doped TiO2/long-afterglow phosphor system for photocatalytic degradation of acetaldehyde in gaseous phase (Li et al. 2011a, b, 2012a). Only recently, doped TiO2/longafterglow phosphor system was used for the photocatalytic treatment of water and wastewater. As an example, in 2015, Sacco et al. (2015a, b) developed an energy storage photocatalysis material consisting of a visible light-active Fe-N-doped TiO2 functionalized with long-afterglow phosphors with green emission and applied this material in the photocatalytic removal of atrazine and oxytetracycline (Sacco et al. 2015a, b). The obtained results demonstrated that, when Fe-N-doped TiO2 was coupled with long-afterglow phosphors, the photocatalytic activity proceeds also after turning off the external light (Fig. 5.10). Another example is the study carried out by Alberti et al. (2017) who developed a N-P-doped TiO2 supported on (3ZnO: Ga2 O3:2GeO2): Cr3+ persistent luminescence material having emissions at 350 nm and 700 nm under simulated solar light irradiation. In particular, the total degradation of ofloxacin was achieved by irradiating the system for 10 min alternately to 10 min of darkness for a total time of 50 min (total irradiation time: 30 min) (Alberti et al. 2017).

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Organic Up-Conversion Phosphors in Photocatalysis

Among different approaches to enhance the up-conversion properties of phosphor materials, organic up-conversion phosphors (organic up-conversion phosphors) could be a valid alternative to up-conversion phosphors. Organic up-conversion phosphors are based on the triplet-triplet annihilation mechanism and present a quantum yield (3–40%) at low excitation intensities (1–10 mW/cm2), significantly higher than that one obtained with inorganic up-conversion phosphors (Kim and Kim 2012; Zhao et al. 2011). The TTA mechanism is obtained, thanks to the simultaneous energy transfer processes between two organic chromophores. The first chromophore, named sensitizer (such as Pd complex octaethylporphyrin), is able to absorb low-energy photons and transfer the excitation energy to the second chromophore, named acceptor (such as 9,10-diphenylanthracene) by means of intermolecular triplet-triplet energy transfer (TTET) (Singh-Rachford and Castellano 2010). A schematic picture of the triplet-triplet annihilation mechanism for organic up-conversion phosphors is reported in Fig. 5.11. Considered the organic up-conversion phosphor properties, high quantum yield at low excitation intensities seems to be a promising solution for overcoming the problems related to the large-scale application of photocatalysis for water pollution. In particular, organic up-conversion phosphors together with a suitable photocatalyst could be able to catch the radiation coming from external light sources and emit their specific radiation in the core of photoreactor. Until now there are few papers reporting the coupling of a photocatalyst with organic up-conversion phosphors, and they too concern mainly with photocatalytic water splitting (Monguzzi et al. 2017). Recently, the application of organic up-conversion phosphors in photocatalytic wastewater treatment was the object of

Fig. 5.11 Schematic diagram of the energy level of the sensitized triplet-triplet annihilation process between Pd complex octaethylporphyrin (PdOEP) and 9,10-diphenylanthracene (DPA); ISC: intersystem crossing, TTET triplet-triplet energy transfer, TTA: triplet-triplet annihilation (SinghRachford and Castellano 2010)

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Fig. 5.12 Schematic energy transfer mechanism accounting for enhanced photocatalytic processes in semiconductor catalysts (N-TiO2) coupling with organic up-conversion phosphors. (Vaiano et al. 2015a); OUPs: organic up-conversion phosphors, N-TiO2 : N-doped TiO2, E : energy

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Fig. 5.13 Effect of N-TiO2/organic up-conversion phosphors on the degradation of organic dyes after 60 min under visible light irradiation emitted by white LEDs (nominal power: 6 W). (Vaiano et al. 2015a); N-TiO2 : N-doped TiO2, OUP: organic up-conversion phosphors

Vaiano et al. that selected organic up-conversion phosphors (prepared according to the method reported by (Liu et al. 2012)) emitting light at specific wavelength (430 nm) able to photoexcite a visible light-active N-doped TiO2 (with energy bandgap of 2.5 eV) (Fig. 5.12) (Sacco et al. 2012). The used organic up-conversion phosphors are excited by green light (544 nm). This system (N-TiO2/organic up-conversion phosphors) was proven effective in the removal of several organic dyes (Fig. 5.13) and antibiotics (like spiramycin) (Fig. 5.14) from water and wastewater (Sacco et al. 2017; Vaiano et al. 2015a)

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1 0.9

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Fig. 5.14 Total organic carbon profile as a function of irradiation time for N-TiO2 and N-TiO2/ organic up-conversion phosphor photocatalysts under visible light irradiation emitted by white LEDs (nominal power: 6 W). (Sacco et al. 2017); N-TiO2 : N-doped TiO2

showing in all cases an enhancement of photocatalytic performances in comparison to pure N-TiO2 nanoparticles. Figure 5.14 shows the photocatalytic performances as a function of irradiation time for N-TiO2/organic up-conversion phosphors (Sacco et al. 2017) in terms of spiramycin mineralization. In particular, 58% of total organic carbon was removed after 180 min of visible light irradiation compared to pure N-TiO2, which after 300 min of irradiation showed a TOC removal of only about 31%.

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Chapter 6

Nanocarbons-Supported and PolymersSupported Titanium Dioxide Nanostructures as Efficient Photocatalysts for Remediation of Contaminated Wastewater and Hydrogen Production Kakarla Raghava Reddy, M. S. Jyothi, A. V. Raghu, V. Sadhu, S. Naveen, and Tejraj M. Aminabhavi

Contents 6.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.1.1 Heterogeneous Semiconductor Photocatalysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.1.2 Potential TiO2-Based Photocatalysts . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.1.3 Limitations of the Fine Powder Form of TiO2-Based Photocatalysts . . . . . . . . . . . . 6.2 TiO2 Photocatalysts with Polymer-Based Hybrid Photocatalysts for Wastewater Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.2.1 Need for Immobilization of TiO2-Based Photocatalysts . . . . . . . . . . . . . . . . . . . . . . . . . 6.3 TiO2 Photocatalysts Supported with Nanocarbons for Wastewater Treatment . . . . . . . . . . 6.3.1 TiO2-Functionalized Nanocarbon-Based Photocatalysts . . . . . . . . . . . . . . . . . . . . . . . . . 6.4 Conclusions and Future Outlook . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract Organic contaminants (textile dyes, pesticides) in industrial wastewater have adverse effects on the environment and human health. Such environmental

K. R. Reddy (*) School of Chemical & Biomolecular Engineering, The University of Sydney, Sydney, NSW, Australia M. S. Jyothi · A. V. Raghu · S. Naveen Department of Basic Sciences, School of Engineering and Technology, JAIN Deemed-to-be University, Bangalore, Karnataka, India V. Sadhu School of Physical Sciences, Banasthali Vidyapith, Banasthali, Rajasthan, India T. M. Aminabhavi Soniya College of Pharmacy, Dharwad, Karnataka, India © Springer Nature Switzerland AG 2020 Inamuddin et al. (eds.), Nanophotocatalysis and Environmental Applications, Environmental Chemistry for a Sustainable World 30, https://doi.org/10.1007/978-3-030-12619-3_6

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pollutants are resistant in the environment and are difficult to completely remove through treatment techniques. Therefore, titanium dioxide (TiO2) nanostructurebased photocatalytic processes have received much attention due to their environmentally green nature with high efficiency for complete photodegradation of organic pollutants to produce safe and clean water. In this chapter, zero-dimensional to three-dimensional TiO2 nanostructures functionalized with various polymeric and nanocarbon hybrid materials are discussed as low-cost, nontoxic, and highly efficient photocatalytic materials for photodegradation of chemical pollutants, in comparison with pristine TiO2, through expansion of the visible light photoresponse and regulation of the bandgap properties of TiO2. Various chemical synthesis methods, surface modifications with various polymers and nanostructured carbons, compositions, morphological structures, growth mechanisms, physicochemical properties, electronic and optical characteristics, and photocatalytic mechanisms (e.g., reactive oxygen species generation) of various heterostructured TiO2-based photocatalysts are discussed in terms of their prospects and future challenges in the fields of photocatalytic environmental remediation and hydrogen generation. Keywords Photocatalysis · TiO2 · Nanocarbons · Graphene · Reduced graphene oxide · Carbon nanotubes · Conjugated polymers · Hybrid photocatalysts · Chemical synthesis · Surface modification · Morphology control · Heterostructures · Bandgap properties · Photocatalytic mechanism · Organic chemical pollutants · Water treatment · Organic dye degradation · Environmental remediation · Hydrogen evolution

Abbreviations λ λmax 1D 2,4-D 2-CP A AC ALD APS B BPA C C0 CB CEPDA

Wavelength (nm) Specific wavelength maximum (nm) One-dimensional 2,4-Dichlorophenoxyacetic acid 2-Chlorophenol Absorbance Activated carbon Atomic layer deposition Ammonium persulfate Path length of sample (m) Bisphenol A Concentration (mol/m3) Initial concentration (mol/m3) Conduction band Electrophoretic deposition–anodization

6 Nanocarbons-Supported and Polymers-Supported Titanium Dioxide. . .

CFL CNT COD CVD DSC Ε e
Eg FE-SEM FTIR G GO Η h+ HOMO hv K LED LUMO MB MO MWCNT P(3HB-co-3HHx) PAA PANI PC PE PET PMMA PP PPF PPy PS PSP4VP PTh PVA PVAc PVC PVDF R rGO RhB SC SDS

Compact fluorescent lamp Carbon nanotube Chemical oxygen demand Chemical vapor deposition Digital scanning calorimeter Molar absorptivity (m2/mol) Electron Bandgap energy Field emission scanning electron microscopy Fourier transform infrared Graphene Graphene oxide Degree of photocatalytic degradation Hole Highest occupied molecular orbital Photon energy Rate constant (min
1) Light-emitting diode Lowest unoccupied molecular orbital Methylene blue Methyl orange Multiwalled carbon nanotube Poly(3-hydroxybutyrate-co-3-hydroxyhexanoate) Poly(acrylic acid) Polyaniline Polycarbonate Polythene Poly(ethylene terephthalate) Poly(methyl methacrylate) Polypropylene Polypropylene fabric Polypyrrole Polystyrene Poly(styrene)-co-poly(4-vinylpyridine) Polythiophene Poly(vinyl alcohol) Polyvinyl acetate Polyvinyl chloride Poly(vinylidene difluoride) Relative concentration Reduced graphene oxide Rhodamine B Semiconductor Sodium dodecyl sulfate

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SWCNT T T Tc TCP TEM Tg TGA UV UV-Vis VB XRD

6.1

Single-walled carbon nanotube Time (min) Transmittance Crystallization temperature (K) Transformer-coupled plasma Transmission electron microscopy Glass transition temperature ( C) Thermogravimetric analyzer Ultraviolet Ultraviolet–visible Valence band X-ray diffraction

Introduction

Efficient utilization of solar energy is one of the major goals of modern science and engineering. The most outstanding utilization of solar energy is through employment of photocatalysts. Photocatalysis can be defined as acceleration of photoreaction in the presence of a catalyst. As noted by Teoh et al. (2012), a photocatalyst “modifies the speed of reaction without being involved in the modification.” The concept of photocatalysis came to light after the discovery of water splitting into hydrogen and oxygen, using TiO2 as a photocatalyst, by Fujishima and Honda in 1972 (Fujishima and Honda 1972). Semiconductor (SC) metal oxides have been commonly used photocatalysts because of their favorable light absorption abilities, electronic structures, charge transport features, and excited-state lifetimes. Photocatalytic chemistry involving SC materials has attracted esoteric interest in both academic and technological research (Linsebigler et al. 1995; Mills and Le Hunte 1997). It has various applications to deal with important environmental issues such as water treatment (removal of heavy metal ions, water disinfection, etc.) and energy-based applications (photovoltaics, hydrogen generation for fuel generation, etc.) (Ibhadon and Fitzpatrick 2013). Semiconductors are the most popular photocatalysts in use because of their special light absorption properties, electronic structures, charge transport characteristics, and excited-state lifetimes. Semiconductor photocatalysis has been applied for (a) destruction of microorganisms such as bacteria and viruses, (b) inactivation of cancer cells, (c) water splitting for hydrogen production, (d) removal of pollutants in the air, and (e) odor control and fixation of nitrogen (Hoffmann et al. 1995; Nakata and Fujishima 2012). There are two types of catalysis: homogeneous and heterogeneous catalysis. In homogeneous catalysis, the reactants and the catalyst exist in the same phase. The most commonly used catalysts are ozone and photo-Fenton systems (Fe+ and Fe+/ H2O2). In heterogeneous catalysis, the reactants and the catalyst exist in different phases. This is the most intensively studied area of research because of its potential

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use in environmental and energy-related areas, wherein light-absorbing SC catalysts are used (Hoffmann et al. 1995). An ideal photocatalyst should be stable, inert, nontoxic, inexpensive, and sustainable toward a broad range of the solar spectrum. It should give high power conversion efficiency and quantum yield, with good adsorption in the solar spectrum (Bhatkhande et al. 2002).

6.1.1

Heterogeneous Semiconductor Photocatalysis

In the heterogeneous photocatalysis process, the photoinduced reactions occur at the surface of the catalyst where the interface of the catalyst and substrate is subjected to light (photons) of appropriate energy (Fox and Dulay 1993). The principles of heterogeneous SC catalysis involve the steps described in the following sections (Li 2013). 1. Photon absorption and separation of bands in Semiconductor: When a photon with energy (hν) equal to or higher than the bandgap energy (Eg) of the SC is absorbed, an electron is promoted from the valence band (VB) to the conduction band (CB) of the SC. This results in generation of holes (h+) in the VB and electrons (e
) in the CB. 2. Electron–Hole recombination In the absence of suitable separation forces, the e
and h+ can recombine within a few nanoseconds on the surface or in the bulk of the particle to dissipate the input energy as heat. 3. Electron–Hole utilization The excited state of the SC may result in recombination or can react with the electron donors and electron acceptors adsorbed on the surface of the SC, causing subsequent anodic and cathodic redox reactions. The probability and the rate of charge transport in any semiconductor-mediated photoreaction depends on the respective band edge positions of the SC with that of redox potential levels of adsorbate species. The reduction and oxidation ability of the e
and h+ are decided by the energy levels placed at the bottom and top of the CB and VB, respectively. The appropriate potential state of the acceptor species is thermodynamically required to be lower than the CB potential of the SC. The potential level of the donor needs to be above the VB position of the SC to donate an electron to the vacant hole. In heterogeneous photocatalysis, breakdown or conversion of complex pollutants into simple nonhazardous substances takes place without any significant destruction of the redox catalytic activity of the SC. Usually the SC photocatalyst reacts with the water molecule rather than the pollutant directly. Oxidation of water by h+ produces a hydroxyl radical—a powerful oxidant, which can act on the pollutant (organic). At the CB, reduction of adsorbed oxygen to oxygen radicals occurs, thus preventing recombination of e
and h+, or e
can also be used to reduce heavy metal ions in

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Fig. 6.1 Semiconductor excitation by bandgap illumination, leading to the creation of an electron (e
) in the conduction band (CB) and a hole (h+) in the valence band (VB), and utilization of e
and h+ for pollutant degradation. hv photon energy. (Reproduced from Ibhadon and Fitzpatrick (2013))

water. Removal or reduction of heavy metal ions from water is an upcoming field in heterogeneous SC photocatalysis (Pirkanniemi and Sillanpää 2002). An advantage of heterogeneous SC photocatalysis is that it provides a simpler, low-cost, effective process for treatment of contaminated water (heavy metals, organic and inorganic pollutants, etc.), and air purification. The process of heterogeneous photocatalysis is depicted in Fig. 6.1.

6.1.2

Potential TiO2-Based Photocatalysts

An ideal semiconductor should be inexpensive, photostable, nonhazardous for humans and the environment, easy to produce, and capable of catalyzing a reaction efficiently. ZnO is unstable in water because it rapidly dissolves in water to form Zn (OH)2 on the ZnO surface, thus reducing the catalytic properties (Gupta and Tripathi 2011). Among various semiconductors (such as SnO2, TiO2, ZnO, and CdS), TiO2 has been widely applied in industry and has been the focus of a range of works in the photocatalysis area. This material has a bandgap energy of 3.2 eV, without any modification, and absorbs near-ultraviolet (near-UV) light (380 nm). The VBs are approximately þ3.1 V, while the CBs are
0.1 V (Robertson 1996). As can be seen by the oxidizing and reducing values, TiO2 can be widely used in redox reactions (Robertson 1996). TiO2-based photocatalysts have received increasing interest in many technological applications—for example, photocatalysts, photovoltaic materials, gas sensors, optical coatings, structural ceramics, electrical circuit varistors, biocompatible materials for bone implants, and spacer materials for magnetic spin valve systems. TiO2 is

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widely used because of its availability, low price, photocatalytic activity, and photostability (Nakata et al. 2012). Water splitting using TiO2 is an important application, which was discovery by Fujishima and Honda in 1972 (Bai et al. 2012). By production of hydrogen, electricity can be generated in a fuel cell without having CO2 as a by-product. Additionally, TiO2 can decompose organic materials; thus, it can be used in wastewater and air pollution treatment.

6.1.3

Limitations of the Fine Powder Form of TiO2-Based Photocatalysts

A major drawback of TiO2-mediated photocatalysts is rapid electron–hole recombination (Choudhury et al. 2013) with a wide bandgap, which narrows its application for photoactivity in solar spectra (Mou et al. 2014). The dimensionality is another characteristic of TiO2 that needs to be improved to provide better chemical and physical performance. To date, several techniques have been developed to find suitable one-dimensional (1D) TiO2 nanomaterials for use as nanorods, nanowires, and nanotubes (Liu et al. 2012).

6.1.3.1

Comparison of Synthesis Methods

Numerous synthesis methods have been used to produce nanostructured photocatalysts. The commonly used methods include the sol–gel method, hydrothermal processing, electrochemical processing, the microwave method, and sonication processes. The sol–gel method is one of the most common methods used to produce functional nanohybrid photocatalysts. It allows compositional and microstructural tailoring through control of the precursor chemistry, such as concentrations, and processing conditions (Wang and Ying 1999). This approach makes it possible to control a number of determining parameters of the final product, such as homogeneity, purity, nanostructure, porosity, and surface area (Yang et al. 2013). The sol–gel method is simple and effective for deriving a unique metastable structure at low reaction temperatures, making it a safe and attractive option for synthesis of nanocomposites. Hydrothermal processing is widely used in the production of advanced ceramic powders. In hydrothermal treatment the grain size, particle morphology, crystalline phase, and surface chemistry can be controlled via processing variables such as the sol composition, pH, temperature, pressure, and aging time (Wang and Ying 1999). It requires moderate temperatures and pressures. The necessity for additives and surface treatments render it a more difficult process than the sol–gel method (Byrappa and Yoshimura 2001). Hydrothermal treatment also offers less flexibility of structural control than the sol–gel method (Orikasa et al. 2007).

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Electrochemical methods, including electrospinning, are versatile for generating ultrathin fibers made of various materials under the influence of an electrostatic field (Liu et al. 2006). However, electrospun nanofibers suffer from clogging and low efficiency, thus making this an inefficient process for production of photocatalysts (Zhou et al. 2009). However, if nanofibers were required, electrospinning would provide the most attractive synthesis route. The microwave-assisted method is essentially a process involving energy in the form of high-frequency electromagnetic waves, typically between 900 and 2450 MHz (Chen and Mao 2007). At lower microwave frequencies, conductive currents flowing within the material due to movement of ionic molecules can transfer energy from the microwave field to the material, while at higher frequencies, the energy absorption is predominantly due to molecules with a permanent dipole, which tend to reorientate under the influence of a microwave electrical field (Chen and Mao 2007). Microwave radiation can potentially shorten the synthesis time; however, as this field is not widely used, there are many uncertainties when it comes to producing new materials, especially those on polymer substrates (Tierney and Lidström 2005). Sonication is the utilization of ultrasound to disrupt the molecules in a particle, causing them to increase their distance (Santos et al. 2009) and eventually break apart, forming smaller particles. However, sonication processes are hindered by low efficiency for large-scale materials; therefore, potential scaling up is not possible with this process (Gogate et al. 2002). Therefore, for facile fabrication of TiO2 nanoparticles, the sol–gel method is most suitable. It is a relatively safe procedure and, more importantly, it is possible to easily control the morphology and structure using different concentration and processing conditions. A combination of methods has been adopted to improve TiO2 properties. For instance, a hydrothermal sol–gel approach has been applied to improve the phase state, electrical properties, optical properties, and photocatalytic performance of TiO2 (Liu et al. 2012).To make TiO2 able to work in sunlight, structural defects or dopants have been added to the TiO2 structure, such as the nonmetals C, N, S, or the metals Cu and Ag (Liu et al. 2012). These modifications also incorporate a subband in the TiO2 bandgap (Mou et al. 2014). Apart from the aforementioned methods, the photocatalytic efficiency of TiO2 can be enhanced by coating or immobilizing TiO2 on various inorganic or polymer substrates.

6.1.3.2

Improvements in TiO2 Performance by Structural Change, Doping, and Hybridization

TiO2 has three crystallographic forms: brookite, anatase, and rutile. The anatase structure has been intensively investigated because of its greater photocatalysis activity. Among the anatase facets {001}, {100}, {101}, and {010}, the major attention is on the synthesis of {001}. An Ag/TiO2 nanofiber heterostructure was developed by a single-step electrospinning process. The compound showed high photocatalytic performance

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under solar light, due to electronic excitation of Ag nanoparticles and transfer of electrons to the TiO2 CBs. Further studies were performed to investigate the potential application of doped TiO2 heterostructures. The addition of Cu also changed the structure of TiO2 and improved its performance. Cu-doped TiO2 nanoparticles were prepared by a sol–gel method. Cu changes the structure of the TiO2 by increasing formation of the brookite phase along with the anatase phase. This new material has been used as a photocatalytic agent in decomposition of gas-phase alcohols, acid orange 88, and methylene blue. The Cu d state and oxygen defects can decrease the bandgap of TiO2 through electron transition to these levels. The absorption in the visible range is due to the single electron in the d orbital. Another important improvement in the structure was TiO2 nanosheets with 95% of the facet exposed (Grandcolas and Ye 2012). The new material has the highest facet exposed. Photocatalytic activity under TiO2 cuboids with 53% of facet exposed was achieved (Ballav and Biswas 2003). It showed promising activity for hydrogen evolution and CO2 reduction. This research provided evidence that study of the exposed facet can be feasible way to achieve better materials for solar conversion applications (Grandcolas and Ye 2012).

6.2

TiO2 Photocatalysts with Polymer-Based Hybrid Photocatalysts for Wastewater Treatment

6.2.1

Need for Immobilization of TiO2-Based Photocatalysts

Many efforts have been made to coat TiO2 on several substrates, with a view to some advantages. Immobilization of TiO2 provides ease of posttreatment recovery, which reduces operational costs at scale-up levels. This method provides relatively high quantum efficiency as compared with the powder form (Byrne et al. 1998). In the case of water treatment, it also minimizes loss of the catalyst and provides longer contact time between the photocatalyst and the pollutants that are to be degraded. Despite these advantages, immobilization also has some disadvantages, which include a reduced surface area for the photocatalytic reaction, and it needs welldefined techniques and equipment to establish a well-coated or immobilized material. Overall, the advantages of immobilization overweigh the disadvantages and have drawn interest from researchers to continue their efforts in immobilizing TiO2 on competent substrates to obtain maximum efficiency.

6.2.1.1

Features of a Stable Substrate, and Available Substrates

The overall efficiency of an immobilized catalyst depends on the features of the catalyst and the support/substrate. There must be enough compatibility between TiO2 and the substrate for stable immobilization (Zhiyong et al. 2008). A prepared

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composite must be stable under different reaction conditions and longer operation times (Shan et al. 2010). On the whole, the immobilization method should not affect the photocatalytic activity. Various substrates are available to immobilize photocatalysts, such as glass spheres, glass plates, wool, fabrics, polythene (PE) sheets, plastics (Kwon et al. 2004), zeolites, steel materials, anodized iron (Fabiyi and Skelton 2000), and inorganic carbon materials (Yuranova et al. 2004). The extensive list of available substrates indicates that the only reason for any researcher to not try a material as a support could be poor adsorption or poor compatibility with the photocatalyst. Among all support materials, use of polymers for support has been a widely used method because of several advantages. Many polymers are thermoplastics and are readily available. These thermoplastic materials possess thermosoftening properties, which facilitate easy coating on TiO2 by simple thermal treatment (Shan et al. 2010). Most of the polymers are mechanically stable and chemically inert, also possessing long durability (Fabiyi and Skelton 2000; Magalhães et al. 2011). The availability of polymers is vast; most of them do not undergo oxidation voluntarily and hence they exhibit UV resistance (Fabiyi and Skelton 2000). The hydrophobic nature of the polymers allows for preconcentration of the organic pollutants on their surface and increases the adsorption efficiency (Magalhães et al. 2011). These characteristics of polymers have allowed many researchers to make best use of them as support materials for photocatalysis. In 1995, Tennakone et al. used a polymer support to anchor titania on it for the first time (Tennakone et al. 1995). They used PE films as a support. Many types of support have since been tested by numerous researchers, and some of them are mentioned here: polyaniline (PANI) (Li et al. 2008), polyvinyl acetate (PVAc) (Brezová et al. 1994), polyvinyl chloride (PVC) (Cho and Choi 2001), polystyrene (PS) beads (Fabiyi and Skelton 2000; Magalhães and Lago 2009), polyvinyl chloride (PVC) (Abd El-Rehim et al. 2012), polypropylene fabric (PPF) (Han and Bai 2010), polyethylene terephthalate (PET) bottles (Fostier et al. 2008), poly(styrene)-co-poly (4-vinylpyridine) (PSP4VP) (Murugan and Rangasamy 2011), poly(methyl methacrylate) (PMMA) and polycarbonate (PC) (Langlet et al. 2002), cellulose fibers, polypropylene (PP) granules, and so on.

6.2.1.2

Comparison of Polymeric Supports for Wastewater Treatment

With encapsulation of metal oxides with conjugated polymers, composites possess better properties, such as better thermal, electrical, mechanical, and rheological properties. This is due to combination of the properties of the different materials (Reddy et al. 2010). As TiO2 requires a support for effective recovery after each loading, immobilizing it into films provides an attractive option for multiple cyclic loadings. A more attractive alternative is to enhance the photocatalytic effect by using a conducting polymer, where the advantages of polymer processing and stability are coupled with the efficiency of TiO2. Conducting polymers—such as PANI, polythiophene (PTh), polypyrrole (PPy), and their derivatives—have been

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Table 6.1 Physicochemical characteristics of organic conjugated polymers Properties Thermal stability ( C) Tg ( C)

PANI 400 (Chandrakanthi and Careem 2000) 130 (Cardoso et al. 2007)

Conductivity (s/cm) Cost of monomer ($/kg)a Synthesis

10 (Kumar and Sharma 1998) 197 (Sigma-Aldrich 2013a) Easy

Toxicity of monomer (mg/kg)b

464 (ScienceLab.com 2005a)

PTh 380 (Mohammad et al. 1995) 231 (Kelkar and Chourasia 2012) 200 (Kumar and Sharma 1998) 226 (Sigma-Aldrich 2013b) Difficult (Liu and Liu 2009) 420 (ScienceLab.com 2005b)

PPy 300 (Basavaraja et al. 2009) 134 (Basavaraja et al. 2009) 600 (Kumar and Sharma 1998) 936 (Sigma-Aldrich 2013c) Easy (Ballav and Biswas 2003) 98 (ScienceLab.com 2012)

PANI polyaniline, PPy polypyrrole, PTh polythiophene, Tg glass transition temperature Costs are expressed in 2013 US dollar values b Acute toxicity in mice a

reported as promising sensitizers to extend the spectral response of TiO2 to visible light effectively in TiO2 photocatalysts. The effective charge separation of photogenerated carriers is due to the heterojunction built between TiO2 and the conducting polymer (Liao et al. 2011). PANI possesses very attractive properties as a photosensitizer, as it is a conducting polymer with low cost, high environmental stability, and good electrical conductivity (Dey et al. 2004). It is also easy to synthesize and can be easily processed with a simple acid-doping/base-dedoping mechanism (Jamal et al. 2013). The overall properties of PANI, PTh, and PPy are summarized in Table 6.1. According to the properties listed in Table 6.1, the three conducting polymers PANI, PTh, and PPy exhibit good thermal stability up to 300  C; thus, photocatalytic dye degradation treatment will never go beyond the point where the polymers start degrading. As is clear from the table, PANI possesses the best stability, making it the most environmentally suitable polymer to use for photocatalytic degradation of pollutants. The glass transition temperature (Tg) tells us how brittle and flexible the polymers are. PANI and PPy are less flexible than PTh. PANI and PPy can form powders, while PTh can be processed as thin films. However, PANI and PPy can be immobilized into films if coated on a surface. Although PTh would be more useful in a cyclic water treatment application, PTh would pose problems during synthesis and is less environmentally stable, which makes it less suitable as a polymer support. The conductivity of PANI is significantly lower than those of PTh and PPy: 20 times lower than that of PTh and 60 times lower than that of PPy. However, the conductivity is also dependent on the synthesis method, the morphology, and the size of the polymer particle. Specifically, in comparison with bulk polymers, nanoscale polymers show better conductivity due to their larger surface area. Kolouch et al. (2006) reported a study on the relationship between conductivity and

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photocatalytic activity. Their results showed that there was a general tendency for a material with higher conductivity to exhibit higher photocatalytic activity; however, this tendency does not always apply, as there is no direct relationship between conductivity and photocatalytic activity (Kolouch et al. 2006). A comparison between the three monomers shows that PANI is the least expensive alternative, making it a more attractive option to synthesize and potentially scale up. Furthermore, synthesis of PANI and PPy is also relatively simple as compared with that of PTh. Thiophene monomers have a high oxidation potential, which requires more energy or a catalyst for polymerization, unlike aniline and pyrrole monomers, which are able to polymerize in a chemical oxidative reaction at room temperature (Ballav and Biswas 2003; Liu and Liu 2009).The toxicity of aniline, thiophene, and pyrrole monomers has been analyzed, which showed that thiophene and pyrrole were more toxic than aniline (by 44 mg/kg and by 366 mg/kg, respectively, in a mouse specimen). The toxicity of thiophene and pyrrole pose safety issues during synthesis of the polymer, leading to the need for a more stringent and controlled manner of production. The toxicity of aniline is due to its NH2 group, whereas the toxicity of thiophene is due to its SH group. Lai et al. (2010) compared the performance of PPy and PANI anode composites for use in lithium ion batteries. PPy and PANI were synthesized using the same method; however, PPy exhibited poorer electrochemical performance than PANI, which was due to the better charge retention and recyclability of PANI as compared with PPy (Lai et al. 2010). The UV–visible light (UV-Vis) absorbance spectra of PANI (Abdulla and Abbo 2012), PTh (Senthilkumar et al. 2011), and PPy (Chougulea et al. 2011) demonstrate that they all show good light absorption of visible light in the region of 400–700 nm. However a major difference between PANI, PTh, and PPy is their absorption efficiencies at 500 nm, which is higher in the case of PTh. That of PPy drops drastically, in comparison with PANI, which increases its absorption at lower light energies. For extension of a photocatalytic response toward the visible light range, a polymer that has absorption greater than 400 nm is favorable, as PANI exhibits strong absorption in the lower energy levels. The absorption range of PANI is more suited to TiO2, as a composite of the two combines their absorption ranges, extending its photocatalytic efficiency. Surface hybridization of PANI/TiO2 composite particles induces visible light photon absorption (Xu et al. 2011). The mechanism of charge separation is shown in Fig. 6.2, where Fig. 6.2a shows the photocatalysis process under UV light and Fig. 6.2b shows the process under visible light. As shown in Fig. 6.3a, PANI/TiO2 composite catalysts induce generation of electron–hole pairs under UV light irradiation, yielding radical and hole oxidants (Zhang et al. 2008). The CB position of TiO2 is lower than that of the lowest unoccupied molecular orbital (LUMO) of PANI; therefore, TiO2 particles can trap photogenerated electrons in PANI/TiO2 composite particles. The PANI layer absorbs visible light, which induces excited-state electrons, in which these photoexcited electrons are injected into the CB of TiO2, transferring it to the active site of the surface of the photocatalyst. These electrons react with water and oxygen to generate hydroxyl

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Fig. 6.2 Photocatalysis process of polyaniline (PANI)/TiO2 composites under ultraviolet (UV) light (a) and under visible light (b). CB conduction band, HOMO highest occupied molecular orbital, LUMO lowest unoccupied molecular orbital, VB valence band. (Reproduced from Xu et al. (2011))

Fig. 6.3 Various forms of carbon nanomaterials. (Reproduced from Yan et al. (2016), published by the Royal Society of Chemistry)

radicals and superoxide radicals, which have very strong oxidizing effects (Xu et al. 2011). PANI=TiO2 þ hυ ! PANIþ =TiO2 þ ecb

ð6:1Þ

ecb þ O2 ! O2∙


ð6:2Þ

PANIþ =TiO2 ! Photocatalyst þ hυþ VB

þ þ þ hυVB þ H2 O⇆H þ OH ! H þ OH ∙

ð6:3Þ ð6:4Þ

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Dye þ OH ∙ ! Degradation Products

ð6:5Þ

Equations (6.1), (6.2), (6.3), (6.4), and (6.5) show the mechanism of the PANI/ TiO2 composite for harvesting light in the visible region. A positively charged hole forms by migration of electrons from the TiO2 VB to PANI, triggering production of very reactive superoxide and hydroxyl radicals (Nabid et al. 2013). With use of this semiconductor, titania-mediated photocatalysis remains the most promising option for solar energy utilization–related applications. Recent studies focusing on polymer-supported photocatalysis are summarized in Table 6.2. APS ammonium persulfate, CFL compact fluorescent lamp, LED light-emitting diode, MB methylene blue, MO methyl orange, P(3HB-co-3HHx) poly (3-hydroxybutyrate-co-3-hydroxyhexanoate), PAA poly(acrylic acid), PANI polyaniline, PMMA poly(methyl methacrylate), PPy polypyrrole, PVA poly (vinyl alcohol), PVDF poly(vinylidene difluoride), RhB rhodamine B, UV ultraviolet

6.3 6.3.1

TiO2 Photocatalysts Supported with Nanocarbons for Wastewater Treatment TiO2-Functionalized Nanocarbon-Based Photocatalysts

In the quest for improved efficiency of photocatalysts, carbonaceous materials offer numerous advantages such as chemical stability, tunable structural and electrical properties, large surface areas, and high mobility of charge carriers (Nouri et al. 2016; Wei et al. 2013). Carbonaceous materials such as fullerenes, activated carbon (AC), carbon nanotubes (CNTs), and graphene have been used for a long time. The adsorption properties of carbon-based materials are surprising and can perceptibly increase the adsorption of pollutants on the catalyst surface. An increase in the concentration of pollutants on the surface of the catalyst itself is a key contribution to enhancement of photocatalysis (Perera et al. 2012; Kamat 2011). Hybrids of catalysts and carbonaceous nanomaterials enhance photocatalytic efficiency in the following ways: they offer higher adsorption of pollutants, as well as visible light, with facile charge separation and transportation (Zhang et al. 2010a). Several reports are available that support the concept that light absorption of TiO2 can be improved by carbonaceous materials in the whole visible range, in comparison with bare TiO2, due to Ti–O–C bond formation between the carbonaceous materials and TiO2 (Carp et al. 2004; Zhang et al. 2010b; Lin et al. 2011). The electron–hole charge separation in TiO2 is efficiently enhanced by carbon materials, since they act as reservoirs to conduct away the electrons (Shi 2013; Park et al. 2016). The following section discusses the pros and cons of carbon material (carbondoped, AC/CNT/fullerene/graphene composite)–based photocatalysts. Recently reported carbon material–supported photocatalysts are listed in Table 6.3.

Tetracycline

Sphere coating

Degussa P25 TiO2

Poly (N-isopropylacrylamide) and calcium alginate Poly (N-isopropylacrylamide) and calcium alginate PVDF grafted with methacrylic acid side chains Multilayers of poly (diallyldimethylammonium chloride) and PAA Multilayers of poly (diallyldimethylammonium chloride) and PAA Polysulfone

RhB

Coprecipitation approach

CdS

Polymelamine– formaldehyde PMMA

MB

MB

Chromium

Layer-by-layer self-assembly

Plasma-enhanced chemical vapor deposition Mixing of TiO2 in casting solution

TiO2

TiO2

Sunlight

UV light

UV light

(continued)

Jyothi et al. (2017a)

Starr et al. (2016)

Mungondori et al. (2016) Starr et al. (2016)

Sunlight

Pb2+ and Fe3+

Dry–wet phase inversion

N-doped TiO2

Ida et al. (2016)

Formation of beads

Fe3O4 nanoparticles

Five black-light fluorescent lamps (6 W)

MO

Formation of beads

Degussa P25 TiO2 MO

Steplin Paul Selvin et al. (2017) Li et al. (2017b) Hartley et al. (2017)

References Li et al. (2017a)

Ida et al. (2016)

300-W xenon lamp

Low-pressure mercury lamp (6 W), UV, and CFL (70 W)

Light sources 32-W high-pressure mercury lamp

(XeLED-Ni1UV-R4–365E27-SS) irradiance of 26 mW/cm2 at 365 nm Five black-light fluorescent lamps (6 W)

RhB

Cysteinecapped ZnO

Target contaminants MO

P(3HB-co-3HHx) fiber

Immobilization methods Hydrothermal treatment of electrospun PMMA nanofibers containing titanium n-butoxide precursor Electrospinning

Photocatalysts TiO2

Substrates PMMA

Table 6.2 Recent studies on polymer-supported photocatalysts

6 Nanocarbons-Supported and Polymers-Supported Titanium Dioxide. . . 153

Photocatalysts TiO2

TiO2

TiO2

TiO2

TiO2

TiO2

TiO2

Black anatase TiO2

TiO2

TiO2

TiO2

Copper-doped titania

Substrates PPy

PANI

Microcrystalline cellulose

Chitosan

Poly(3-hexylthiophene)

Eggshell membrane

PANI/cork

PANI

PVA

Aminated polysulfone

Sulfonated polysulfone

Polythiophene nanorods

Table 6.2 (continued)

Modified sol–gel technique

Solution casting

Solution casting

Modified hydrothermal method and successive in situ polymerization of aniline Dispersion via sonication

Surface modification of eggshell membrane by polyethyleneimine Impregnation method

Sonication and evaporation

In situ chemical oxidative polymerization of aniline with APS oxidant in the presence of hydrothermally synthesized TiO2 nanoparticles Preradiation grafting–embedding method Encapsulation method

Immobilization methods TiO2 used as a co-catalyst

RhB

Chromium

Visible light

Sunlight

Solar simulator resulting in 72 mW cm
2 Sunlight

Triton X-100 Chromium

Visible light

Orange dye MO

Sunlight

RhB

Mercury lamp

Xenon arc lamp

UV light

Light sources Fluorescent bulbs 70 W m
2

500-W tungsten halogen lamp Sunlight

MO

Azo and anthraquinone dyes

MB

Target contaminants Methyl 4-chlorobenzoate RhB, MB, and phenol

Hegedűs et al. (2017) Jyothi et al. (2016) Jyothi et al. (2017b) Chandra et al. (2017)

Li et al. (2017c) Dhanya and Aparna (2016) Zhang et al. (2015) Li et al. (2017d) Sboui et al. (2017) Kavil et al. (2017)

References Petroff Ii et al. (2017) Reddy et al. (2016)

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Table 6.3 Recent studies on carbon material–supported photocatalysts Supports Graphene

Nanomaterials ZnS

Carbon fiber

m-BiVO4 hollow spheres

Graphitic carbon nitride Graphitic carbon nitride Carbon spheres

Iron oxides

β-FeOOH

Immobilization methods Electron TCP-enhanced CVD Hydrothermal

Incipient wetness impregnation method Ultrasonic treatment

Targets Hydrogen evolution MB

Light sources 300-W highpressure mercury lamp 500-W xenon arc lamp

4-Nitrophenol

300-W xenon lamp

RhB

300-W halogen tungsten bulb 300-W xenon lamp

CdS

Modified hydrothermal synthesis

Hydrogen evolution

Carbon

ZrO2

One-pot synthesis

Orange 5 G

14-W visible light lamp

Nitrogendoped rGO

Silver chloride– enwrapped silver Ce–TiO2 and Fe–TiO2

In situ oxidation reaction

2-CP, BPA, phenol, and 2,4-D

150-W xenon arc lamp

In situ sonochemical method In situ hydrothermal method

Crystal violet dye

UV light

p-Cresol and Cr(VI)

250-W lamps

Microwaveassisted hydrothermal method Hydrophilic surfactant–stabilized in situ method One-step solvothermal method –

MB

40-W UV lamps

NOx

UV-A light (~10 W/m2)

NO

UV lamps (8 W and 6 W) UV radiation (4  15 W) and sunlight 20-W LED

G

rGO

WO3, Fe2O3, WO3

rGO

Cu-doped ZnO

G and GO

TiO2

GO

Bi metal sphere

rGO

TiO2–Pt

rGO

NiO

Low-temperature solution method

TiO2

Solvent evaporation method

Amaranth, sunset yellow, tartrazine Nitroaromatics

MB

Mercury lamp (100-W)–UV light

References Chang et al. (2017) Ahmed et al. (2017) Lin et al. (2017) Zheng et al. (2016) Wang et al. (2017a) BailónGarcía et al. (2017) Wang et al. (2017b) Shende et al. (2018) Kumar et al. (2017) Hsieh and Ting (2018) Trapalis et al. (2016) Wang et al. (2017c) Rosu et al. (2017) Al-Nafiey et al. (2017) Song et al. (2016) (continued)

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Table 6.3 (continued) Supports Graphitic carbon nitride CNTs

Nanomaterials Ni complex

Ni–TiO2

Immobilization methods –

Targets Hydrogen evolution

Light sources Visible light

Incipient wetness impregnation method Hydrothermal method

Hydrogen evolution

Liquid phase plasma

MO

Visible light

GO

Bismuth oxyiodide

rGO

TiO2

Microwave synthesis

MB

Sunlight

Activated carbon

Barium ferrite

Sol–gel combustion method

Basic yellow 28

Solar light

rGO

Ag/Ag2S

Ciproflaxin

Visible light

rGO

TiO2

Dispersion on surface –

MO

UV light

References Kalousek et al. (2017) Chung et al. (2017) Vinoth et al. (2017) Kumar et al. (2015) Roonasi and Mazinani (2017) Huo et al. (2018) Cao et al. (2015)

2,4-D 2,4-dichlorophenoxyacetic acid, 2-CP 2-chlorophenol, BPA bisphenol A, CNT carbon nanotube, CVD chemical vapor deposition, G graphene, GO graphene oxide, LED light-emitting diode, MB methylene blue, MO methyl orange, NOx nitrogen oxide, rGO reduced graphene oxide, RhB rhodamine B, TCP transformer-coupled plasma, UV ultraviolet

6.3.1.1

Potential Photocatalytic Improvements with Carbon Nanostructures for Wastewater Treatment

Different forms of carbonaceous materials are shown in Fig. 6.3. Composites of carbon materials with TiO2 come in different forms, such as carbon-doped TiO2, fullerene/TiO2 composites, AC/TiO2 composites, CNT/TiO2 composites, and graphene/TiO2 composites. Combinations of all forms of carbon with TiO2 have attracted significant interest due to their special properties. Carbon-doped TiO2 has received considerable attention because carbon possesses an excellent storage capacity for electrons (Zhang et al. 2013a). Usually, carbon replaces either oxygen or titanium in the crystal lattice of TiO2, making it a visible light–sensitive photocatalyst (Ooyama et al. 2011). Also, when carbon precipitates on the TiO2 surface, it acts as a photosensitizer to transport photogenerated electrons to the CB of TiO2. The major overlap between the near-VB edge doping state and the O2p state makes carbon doping highly preferable to any other type of element doping (Carp et al. 2004; Ni et al. 2007; Han et al. 2009). Ab initio methods rely on density functional theory parameters, proving that substitutional carbon dopants incorporated into TiO2 drastically affect the electronic

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structure of the material. With substitutional carbon, the VB edge shifts to higher energy, in comparison with bare TiO2, which narrows down the bandgap and shows a red shift in the absorption spectra. This is due to penetration of carbon 2p states into the upper VB of TiO2 possessing O2p states (Hao and James 2005). However, formation of oxygen vacancies and substitution of carbon in the vacancies is always observed in oxygen-poor conditions, and carbon substitution of Ti favors oxygenrich conditions (Di Valentin et al. 2005). This greatly depends on the doping method. Usually, carbon doping is achieved by thermal assembly of titania in the presence of a carbon source. Carbon was doped with TiO2 using a hydrothermal method where amorphous titania and glucose were autoclaved at a lower temperature to get carbon-doped titania particles (Ren et al. 2007). The carbon-doped titania catalyst possessed a larger surface area and showed better photocatalytic activity than commercial P25. Different carbon precursors, treatment with a gaseous carbon source, pyrolysis of Ti metals, thermal oxidation of TiC, etc., have been tried by various researchers to enhance the efficiency (Kiran and Sampath 2013; Lin et al. 2013; Li et al. 2012; Zhang et al. 2013b; Hiroshi et al. 2003). However, carbon doping to get titania without using an external carbon precursor. A typical sol–gel method in which titanium butoxide and an HClO4 solution were employed to directly prepare C– TiO2 has been reported (Park et al. 2009). Recently, carbonate-doped titania was prepared by a conventional sol–gel method, along with a subsequent carbonization process in a hypoxic atmosphere (Liu et al. 2017). In this method, formation of the VB tail states and bandgap narrowing were achieved, since the carbonate dopant modified the electronic band structure of the TiO2 by strong electron-withdrawing bidentate carboxylate linkage, as shown in Fig. 6.4. Codoping of TiO2—along with some other elements such as N, S, and B—has been tried by many researchers to improve photocatalytic activity (Tachikawa et al. 2004; Chen et al. 2007; Sakthivel et al. 2004).The catalytic activity was enhanced for carbon-doped TiO2 and for TiO2 codoped with carbon and tungsten, prepared through sol–gel processes in the presence of melamine borate. However, there have been reports describing a decline in the catalytic efficiency of C-TiO2 when it is codoped with other elements (Tachikawa et al. 2004; Sun et al. 2006). The next type of carbon material with the largest internal surface area and porosity (with a wide range of pres from micro to macro) is activated carbon (AC) (Serp et al. 2003; Zhu and Zou 2009). The electron transfer process during photocatalysis is promoted by the high adsorption efficiency of AC (Tryba et al. 2003). Home-prepared AC and Degussa P25 were mixed to get a composite to photo-oxidize 2-propanol in a gas–solid phase (Takeda et al. 1995). The sol–gel method was used for degradation of tetracycline (Martins et al. 2017). There were improved structural parameters such as a smaller crystal size and a predominant anatase phase, with the lowest bandgap energy observed with this. A composite supercapacitor containing para wood–derived activated carbon and TiO2 was developed under different carbonizing temperatures in combination with a sol–gel technique (Nitnithiphrut et al. 2017). The electrochemical charge/discharge of a capacitor depends on the scan rate of the applied potential, and the capacitance is up to 173.69 F/g at a scan rate of 1 mV/s.

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Fig. 6.4 Proposed band structure of CT-300 and visible light photocatalytic process. (Reprinted from Liu et al. (2017), with permission from Elsevier)

A TiO2/AC composite photocatalyst was prepared by a temperature impregnation method and evaluated for oxidation of the pharmaceuticals amoxicillin, ampicillin, diclofenac, and paracetamol (Gar Alalm et al. 2016). Likewise, various synthesis routes have been followed to prepare AC–TiO2, such as the precipitation, hydrothermal, microwave-assisted, sol–gel, dip-coating, and chemical vapor deposition routes (Sharma and Lee 2017; Singh et al. 2016; Liu et al. 2007). Since the chemical bond formed between the support and TiO2 greatly influence the bandgap of TiO2, the selection of an appropriate procedure to ensure absorption of light of a higher wavelength is crucial (Herrmann et al. 1999). Other codopants such as nitrogen (Cojocaru et al. 2009) and iodine (Wang et al. 2016) have also been found to improve the photocatalytic activity of AC–TiO2. A delocalized conjugated system formed by a close shell configuration with 30 bonding molecular orbits is fullerene (C60). Fullerene is known not only for its strong absorption of UV light but also for its moderate absorption of visible light (Xie et al. 1992; Sibley et al. 1992). It is an excellent proton acceptor and efficiently separates photogenerated charge carriers (Apostolopoulou et al. 2009; Krätschmer et al. 1990). Correspondingly, it behaves as a sensitizer on coupling with semiconductor photocatalysts (Orfanopoulos and Kambourakis 1994; Kamat et al. 2004). Functionalization of these fullerenes can further increase the adsorption of dye

6 Nanocarbons-Supported and Polymers-Supported Titanium Dioxide. . . (a) e-h pair generation

(b) resonant e transfer

159

(c) e-h recombination

LUMO+1 resonant transfer

electron LUMO CBM

Ferml level

UV HOMO HOMO-1 hole

VMB

TiO2

–

C60

TiO2

e-h pair

+

C60

TiO2

C60

e-h recombination e transfer

Fig. 6.5 Elemental steps in electron–hole (e
/h+) pair generation, resonant electron (e) transfer, and e
/h+ recombination. Temporal shifts in the energy levels of both C60 and TiO2 are also indicated. (Reprinted from Ozawa et al. (2016), with permission from Elsevier)

molecules; hence, extensive research activities have been devoted to this (Makarova 2001; Guldi and Prato 2000; Accorsi and Armaroli 2010; Terazima et al. 1991). A fullerene/TiO2 composite catalyst was synthesized by growing fullerol (C60–OH) via atomic layer deposition (ALD) (Justh et al. 2017). This enhanced the photocatalytic degradation of MO dye. For ALD, nucleation sites were created by functionalizing bare fullerene with H2SO4/HNO3 treatment, which resulted in C60–OH, then TiO2 was grown by the ALD technique. Metal–fullerene/TiO2 composites increase photocatalytic degradation of methylene blue because of the increased photoabsorption effect of fullerene and the cooperative effect of the metal introduced as a dopant (Meng et al. 2012). A Pt-treated C60–TiO2 composite was investigated for degradation of MO under visible light. The increased efficiency of the Pt-treated composite was attributed to the same cause mentioned above. A schematic representation of electron–hole formation, transformation, and recombination in a fullerene/TiO2 composite catalyst is shown in Fig. 6.5. Another group of carbonaceous materials that have attracted many researchers’ interest is CNTs. The 1D structure—corresponding to outstanding structural, thermal, and electronic properties—offers significant advantages (Serp et al. 2003; Iijima 1991; Xie et al. 2005). CNTs are broadly classified into two types: single-walled CNTs (SWCNTs) and multiwalled CNTs (MWCNTs). A sheet with a single layer of carbon atoms rolled into a tube forms SWCNTs, and carbon atoms forming multiple concentric sheets form MWCNTs (Yao et al. 2008; Ueda et al. 2009). As in other carbon materials, the synthesis routes of CNTs greatly affect the physicochemical properties and morphology of the catalysts. Composites prepared using a sol–gel method have better performance than those prepared hydrothermally (Gao et al. 2009; Aryal et al. 2008). Simple mixing of CNTs with TiO2, followed by illumination with UV radiation, has been reported to give better results in visible light (Wang et al. 2009).

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Inefficient dispersion of TiO2 on CNTs is the main challenge involved in increasing photocatalytic activity. A new method has been proposed to improve TiO2 performance by depositing it onto a hydrophilic CNT surface. The modified surface was achieved by treatment with acid and L-lysine. Improvements in photodegradation of methyl orange under UV and visible light were attributed to the larger surface area achieved, which facilitated the adsorption capacity of reactive and absorbance of photos. Other reasons were the great dispersion of TiO2 on the surface of the CNTs and the efficient contact, which led to dense heterojunctions at the interface between the TiO2 and the CNTs. In comparison with CNTs and active carbon, graphene oxide has better conductivity, as well as a larger surface area. These characteristics can enhance TiO2 photocatalytic performance (Li et al. 2011). A scalable method was used to synthesize surface-fluorinated TiO2 nanosheets (F–TiO2) hybridized with GO (Dai et al. 2014). This novel material was prepared by a hydrothermal treatment and colloidal self-assembly approach. The F–TiO2 nanosheets were anchored on the surface of the GO sheets. The photocatalytic activities were analyzed under UV light by photodegradation of methylene blue. The results showed that the GO/F–TiO2 hybrid catalyst with facet (001) had better photocatalytic performance than commercial P25 and pure TiO2. The enhancement of the photocatalytic activity was mainly due to the reinforcement of electron–hole separation, which could be attributed to the presence of GO. One of the reasons was that GO acted as a recombination center rather than providing an electron pathway. An optimum ratio value of 3% hybrid photocatalyst was found for GO/TiO2 (Matsuoka et al. 2007). Another photocatalyst, a TiO2–GO hybrid, was synthesized using two steps: in situ growth of a uniform TiC layer on graphene oxide sheets, using a molten salt method, following by oxidation conversion of TiC to an anatase TiO2 layer in airflow. The formation of the Ti–O–C bond introduced a mid-bandgap near the TiO2 VB and allowed absorption of light in a longer-wavelength region. This improvement resulted in better performance than P25 in degradation of methylene blue under visible light. Finally, it was explained that to achieve a reasonable absorption capability and photodegradation activity, the ratio of the GO/TiO2 hybrid catalyst should be adjusted (Li 2011). Work using reduced graphene carbon combined with ZnO or TiO2 has provided satisfactory results. ZnO-reduced graphene oxide was prepared by UV-assisted photocatalytic reduction of GO by ZnO nanoparticles. The new material was tested to reduce Cr(VI). It showed a higher photocatalytic performance than pure ZnO. The photocatalytic performance of ZnO–rGO was found to be dependent on the proportion of rGO in the composite and the ZnO–rGO composite. The better performance was due to an increase in the light absorption intensity and reduction of photoelectron–hole pair recombination in ZnO. There was an improvement in the synthesis of rGO–TiO2 with the development of a new technique called the electrophoretic deposition–anodization (CEPDA) method. This has an advantage in fabrication of composite electrodes because it requires just a single step, which provides anodic growth of TiO2 nanotubes and electrophoresis-driven attachment of rGO at the same time. It achieves a balance between the oxygen functional group and electron conductivity, which allows better

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performance of TiO2. Graphite has greater conductivity than GO; however, the absence of oxygen makes interaction with TiO2 difficult. Therefore, the right conditions found using the CEPDA technique represent a potential mechanism to improve the photoelectrochemical performance of TiO2 (Yun et al. 2012). Fullerene–, CNT–, and graphene–TiO2 hybrid photocatalysts were successfully synthesized and a comparative study on their photocatalytic performance was performed. The three synthesized catalytic materials were used to oxidize benzyl alcohol; however, it was found that the three performed similarly to each other, with approximately 40% of the initial benzyl alcohol being left over (Yang et al. 2013). It was found that the smallest amount of graphene (0.1%) was able to oxidize more benzyl alcohol, as compared with CNTs and fullerene (0.5% and 1.0%, respectively). In this study, the graphene composite was able to oxidize 5% more benzyl alcohol (Yang et al. 2013). As shown in one study, graphene has a significant effect in raising the photocatalytic performance of TiO2. Graphene has exceptional properties such as excellent electron mobility, a large surface area (approximately 2600 m2/g), and optical transparency (Yang et al. 2013). A comparison of the photocatalytic performance of graphene–TiO2 and CNT–TiO2 hybrid catalysts, synthesized using a similar method, showed that the graphene composites had a prominent advantage over the CNT composites in controlling the morphology of the nanocomposite. Experiments done by Zhang et al. (2011) suggested that preparation methods play a very important role, as these control the structure of the catalyst. One of the key factors that affect the activity of a nanomaterial is its morphology (Zhang et al. 2011). Therefore, when the fullerene-, CNT-, and graphene-based nanomaterials mentioned above are compared, graphene is the most suitable one for extending the photoresponse of TiO2, owing to its large surface area. Furthermore, it has been reported that because of its simpler structural control, use of graphene promoted better interfacial contact between TiO2 and the graphene surface, exhibiting much more active visible light photocatalytic activity than CNT–TiO2 composite catalysts. Graphene is also less expensive than CNTs, which makes it more attractive, and it is also simpler to synthesize (Solid State Technology 2011).

6.4

Conclusions and Future Outlook

Nanostructure semiconductor–mediated photocatalysis is an effective and viable method for photodegradation of organic contaminants present in air and water. Innumerable efforts have been made to make titania as highly visible active, yet more research efforts are needed to support TiO2 photocatalysts with a polymer matrix or carbonaceous materials for highly efficient photocatalytic applications. Emphasis must be given to developing such effective prototypes.

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Chapter 7

Investigation in Sono-photocatalysis Process Using Doped Catalyst and Ferrite Nanoparticles for Wastewater Treatment Sankar Chakma, G. Kumaravel Dinesh, Satadru Chakraborty, and Vijayanand S. Moholkar

Contents 7.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.2 Dependency of Catalytic Activity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.2.1 Size-Dependent Catalytic Activity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.2.2 Shape-Dependent Catalytic Effect . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.2.3 Interparticle Distance-Dependent Catalytic Effect . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.2.4 Support Interaction and Charge Transfer-Dependent Reactivity . . . . . . . . . . . . . . . . . 7.3 Type of Nanoparticles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.3.1 Non-metallic Nanoparticles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.3.2 Metallic Nanoparticles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.3.3 Semiconductor Nanoparticles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.3.4 Ceramic Nanoparticles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.3.5 Polymer Nanoparticles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.3.6 Lipid-Based Nanoparticles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.4 Types of Nanoparticles Based on Structure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.5 Synthesis and Applications . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.5.1 Discussions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.6 Synergetic Effect . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.7 Conclusion and Overview . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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S. Chakma (*) · G. K. Dinesh · S. Chakraborty Department of Chemical Engineering, Indian Institute of Science Education and Research Bhopal, Bhopal, Madhya Pradesh, India e-mail: [email protected] V. S. Moholkar (*) Department of Chemical Engineering, Indian Institute of Technology Guwahati, Guwahati, Assam, India e-mail: [email protected] © Springer Nature Switzerland AG 2020 Inamuddin et al. (eds.), Nanophotocatalysis and Environmental Applications, Environmental Chemistry for a Sustainable World 30, https://doi.org/10.1007/978-3-030-12619-3_7

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Abstract Industrialization and urbanization affect the environment directly, and water is one of the primary natural resources which are affected significantly. With rapid development of science of nanotechnology, the use of nanomaterials in environmental applications, especially water treatment, has attracted the scientific community in the last decades. Nanomaterials have unique properties, for example, surface-to-volume ratio, quantum effect, low band-gap energy, etc., which give extra features in catalytic performance. This chapter gives a brief introduction of nanomaterials including their classification, shape and structure, type of nanomaterials and their applications in degradation of recalcitrant organic contaminants. Moreover, an attempt was made to emphasize the role of catalyst surface in degradation mechanism in the presence of transient metal ions or other elements and an external oxidant such as H2O2. Additionally, we have also discussed process intensification using sono-hybrid advanced oxidation processes of sono-photocatalysis and heterogeneous Fentonlike reaction for wastewater treatment. Some of our investigations revealed that nanophotocatalyst such as ZrFe2O5 possesses dual characteristic and it contains α-Fe2O3 phase which acts as a centre of recombination for holes and electrons resulting to low photoactivity. However, this phase promotes Fenton-like reaction in presence of H2O2 leading to higher degradation. Therefore, the dual activities of photo and Fenton, ZrFe2O5, were found to be better catalyst for hybrid advanced oxidation processes than other conventional photocatalysts. On the other hand, the doping of transition metal ions into nanophotocatalyst helps to generate more •OH radicals which attack the organic molecules adsorbed on the catalyst surface and enhanced the degradation efficiency. In sono-hybrid advanced oxidation processes, such photocatalysts exhibit negative synergy as the intense shock waves generated due to the transient collapse of cavitation bubbles influence the desorption of organic molecules from the solid surfaces. As a result, low degradation efficiency was seen due to reduction of interaction probability between radicals and organic molecules. Keywords Photocatalysis · Sonocatalysis · Advanced oxidation process · Nanoparticles · Ferrite nanoparticle · Doped catalyst · Degradation · Water treatment · Ultrasound · Cavitation

7.1

Introduction

In the past decades, the environmental issues have become foremost issues due to growing industrialization and urbanization, especially in the developing countries. Among various ecology- and environment-related issues, water scarcity is a primary issue as the clean water is an indispensable resource for the society as well as the variety of industries including process industries, agriculture industries, and food and beverage industries. Estimation says that approximately 50% of the world total population will be suffering from the shortage of water by 2025. At present, approximately 70% of industrial wastewater is being released to the environment without treatment, whereas only 20% of the global water is being treated properly for

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future usages (UN 2016; Zhang et al. 2016). Hence, development of an advanced, low-cost and efficient technology for production of clean water is very important. Recently, a number of nanomaterials have attracted the scientific community for water treatment. Many researchers have already witnessed the potential of the nanomaterials and dedicated to understanding the science behind the catalytic activity of nanomaterials. Nanomaterials vary in dimension from nanometre-scale range (less than 100 nm) and possess excellent properties which differ from their bulk materials (Xu et al. 2012; Weinberg et al. 2011; Laurent et al. 2010). The principal activity in unravelling the nanocatalyst is its synthesis procedure, size and shapes depending upon the area of applications. Investigations are in progress to understand the behaviour of the catalysts as well as their synthesis with costeffective techniques. Moreover, the economy of the catalytic process and production industry also depends on the development of efficient catalysts as 90% of production and process industries are dependent on catalysts (Cuenya 2010). In the catalytic processes, catalyst plays a critical role in breaking the complex and hazardous substances into less toxic by-products during treatment. In most of the commercial and environmental applications, nanomaterials are being used as a catalyst to enhance the chemical reaction kinetics. Therefore, the nanocatalysis as one of the scientific disciplines will have a substantial impact in the future on its applications in the field of catalysis. One of such applications is the treatment of wastewater through advanced oxidation processes. The water treatment and discharge practices need high-efficiency performance for their unit operations; and only the nanocatalyst can offer leapfrogging opportunity in developing sustainable water treatment method by providing high-performance, affordable water and solutions for wastewater treatment (Qu et al. 2012). The conventional techniques for wastewater treatment like adsorption, coagulation, sedimentation, etc. convert the pollutants from one phase to another phase, whereas the catalyst and its hybrid methods have the capability for complete destruction of the complex or recalcitrant compounds including intermediates formed during degradation process (Natarajan et al. 2011). In wastewater treatment process, the catalyst performance can be modified by the selective design, stability and characterization of nanomaterials. The surface phenomena and quantum confinement of the catalyst particles strongly dominate the catalyst properties at nano-level – which basically allows enhancing the physical properties of the material by controlling its electronic energy level through quantum size and surface effects (Mohamed et al. 2012). It has been also found that the catalytic performance is not only strongly materials-dependent but also depends on the sizes, shapes and structures of the materials used in the treatment processes (Tiwari et al. 2012). The size and shape of the catalyst influence the physiochemical properties of the catalyst such as optical properties, adsorption affinity, etc. when they are used in nanoscale. Several efforts have been made in understanding the features of the catalysts and their performances by tuning the sizes and the shapes. It has been also reported that the size and shape of the catalyst envisage the availability of active sites on the catalyst surface which helps for radical formation (Ouyang et al. 2013). Additionally, the morphology of the catalyst also influences charge transfer phenomena or interfacial strain (Mistry et al. 2014a). Investigations have reported that the other factors such as geometry, composition,

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oxidation state and chemical and physical environment also influence the selectivity and catalytic properties of nanomaterials (Mostafa et al. 2010). The desired size and shape of the catalyst can be achieved just by altering chemical and physical conditions during the growth or the synthesis process. The seed-mediated growth forms the major key step process in shape-selective catalyst. Generally, the nucleation process is fast and requires a supersaturated condition with a high reaction rate. Behafarid and Cuenya (2012, 2013) concluded that the initial size selection is not stable during the nucleation and growth following the Ostwald ripening and/or diffusion-coalescence pathways. The other factors like shape and size, density, pretreatment, dopant and support material also influence the stability of the catalyst (Campbell and Sellers 2013). It should also be noted that all these factors are directly or indirectly related and influence the efficacy of catalyst performance. This chapter describes the classification of nanocatalysts, surface mechanism, catalyst activities and their applications for treatment of wastewater.

7.2 7.2.1

Dependency of Catalytic Activity Size-Dependent Catalytic Activity

The studies showed that the catalytic activity enhancement is due to the influence of size and support and interaction among them while forming doping or composite catalyst (Cao et al. 2016). The high activity of the catalyst can be understood at the atomic level by using several models that explain how the enhanced activity depends on the size. The size-dependent activity could be due to the formation of clusters that results in (i) non-metallic behaviour in metallic clusters, (ii) higher density of the catalyst by accumulation of atoms, (iii) change in electronic charge and (iv) increase in a number of active sites (Naitabdi et al. 2006; Ono et al. 2006). Mistry et al. (2014b) reported the catalytic activity of micelle-synthesized Au nanoparticles in the range of approximately 1–8 nm depending upon their size. It has been also reported that by decreasing the size of the nanoparticle from 8 nm to 1 nm, the current density increases significantly. This is attributed to the rise in the number of low-coordinated sites causing an increase in the evolution of H2 and decrease in faradaic selectivity towards CO due to a reduction in the size of the nanoparticles (Mistry et al. 2014b). Recent advances in the study of cluster nanoparticles desirably showed the interest in photocatalysis reaction. Vaneski et al. (2011) have examined the catalyst smaller than 1 nm with comparatively low band-gap alignment and observed that the size of interfaces increases catalytic activity by approximately 2–3 times in various reactions. Schweinberger et al. (2013) demonstrated that the cluster area of 1 nm2 size range showed a strong affinity with ionization potential and cluster properties, like chemical and catalytic activity. The electronic state below 0.04 e nm2 lowers the edge of the conduction band and favours the charge density in the cluster. The change in size varies the band-gap due to the attribution of the electron states (quantum size effects). Thus, the positive binding energy shifts with a decrease in the size of the particle. Increasing the surface-to-volume ratio with decreasing size

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favours more low-coordinated atoms available for interaction and facilitates the dissociation of reactants stabilizing intermediate species (Reske et al. 2014). Furthermore, the effect of catalyst size can also be described with the distinct or support interactions (support charge transfer) – which could be electronic or structural in nature (strain effects).

7.2.2

Shape-Dependent Catalytic Effect

The shape, structure and morphology of nanoparticles play an important role in catalytic performance. Almost, all the physical properties of material directly depend on synthesis method to yield the desired shape. This necessarily indicates that the catalytic reactions occurring at the edge interfaces around the particles are significant due to an increase of active sites with the change of shape of the nanoparticles. Yun et al. (2011) reported the effects on shape-dependent properties of TiO2 nanocrystals (viz. nanospheres and nanorods) for H2 evolution from aqueous ethanol solution under ultraviolet light irradiation, and it has been found that Pt/TiO2 nanorods generate 1.6-folds more H2 gas than spheres. In another study by Camposeco et al. (2016) on the degradation of organic pollutants using TiO2 nanomaterials with different structures and shapes, it was observed that the photocatalytic activity was strongly influenced by the TiO2 nanostructures, viz. nanotubes, nanofibres, nanowires and nanoparticles. The shape, the crystal phase and the cavities of materials can be tuned by thermal treatment as well. Multilayered nanotubular structure of CuO incorporated TiO2 with inner and outer diameters of the nanotubes of 5–7 nm and 10–12 nm, respectively, reported to be highly efficient with better thermal stability than CuO-TiO2 P25 Degussa (Xu et al. 2011). (Kumar et al. 2017) have investigated different morphologies of TiO2 (viz. nanosquares, nanorods and nanotubes) and reported that nanotube predominantly showed higher rate of H2 production (4.6 mmol h
1g
1 cat) due to one-dimensional (1D) nanotubular shape compared to nanorod (2.9 mmol h
1g
1 cat) and nanosquare (2.1 mmol h
1g
1 cat). (Klubnuan et al. 2016) studied the different morphologies of ZnO platelets and ZnO mesh-like lamelle which directly influence the photocatalytic activity. They have reported that by changing the morphology, the crystalline size can be changed and thus the band-gap energy (Eg ¼ 3.222–3.238 eV for platelets and Eg ¼ 3.216–3.226 eV for mesh lamelle).

7.2.3

Interparticle Distance-Dependent Catalytic Effect

The stability and reusability of the catalyst under different reaction conditions can be improved by controlling the size and distribution of the nanoparticles. The distance between the two adjacent particles known as interparticle distance influences the catalyst activity. In fact, the mass transport rate depends on the interparticle distance during the oxidation or reduction reactions. Yang et al. (2013) have studied

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Pt-nanoparticle arrays for oxygen reduction reaction, and it was found that the activity was increased by decreasing the interparticle distance, whereas the formation of H2O2 was increased with increasing interparticle distance. This is due to the particle size-dependent adsorption and interparticle spacing-dependent mass transport (Antolini 2016). (Speder et al. 2014) also observed a similar phenomenon and explained the catalyst activity on the basis of interparticle distance. They have concluded that for a large interparticle distance, the electrical double-layer (EDL) overlap is negligible but it increases as the interparticle distance starts decreasing as shown in Fig. 7.1. It was also found that the activity of the catalyst can be retained up

Fig. 7.1 Interparticle distance effect of nanoparticles and nanoclusters. (Reprinted from Speder et al. (2014) with permission of Royal Society of Chemistry)

Electrical layer Nanoparticle Interparticle distance at the edge

Support

Overlapping of Electrical layers Nanoparticle

Interparticle distance

Support

Overlapping of Electrical layers increases Interparticle distance

Support

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to 80% by optimizing the nano-spatial distribution particle near-maximum interparticle spacing.

7.2.4

Support Interaction and Charge Transfer-Dependent Reactivity

Support interactions can play a significant role in the catalytic performance by affecting (i) the electronic structure via charge transfer, (ii) the crystalline and electronic structure change by the strain from nanoparticle/support interface and (iii) interfacial sites at the support and nanoparticle interface. One of such studies has been reported by Tian et al. (2014). In this study, self-supported nanoporous cobalt phosphide nanowire arrays have been synthesized through a top tactic conversion reaction via low-temperature phosphatization of Co(OH)F/CC precursor. The selfsupported arrays showed excellent catalytic activity in the generation of 3D hydrogen with better durability or stability under neutral and basic conditions. For supported catalyst, the factor that determines the thermal stability and shape is the binding or adhesion energy between the support and the nanoparticles (Kim et al. 2014). The adhesion energy between the catalyst and support is nothing but a long-range van der Waals force between them. These forces are explained by the periodic density functional theory suggesting a strong binding energy between support and catalyst which weakens the adhesion forces of other particles, thus increasing stability. For instance, when MoS2 supported catalyst was employed in H2 evolution reaction, the strong adhesions with the support weaken the hydrogen binding with the catalyst (Tsai et al. 2014). The major advantage of supported catalysts is that they can be recycled and reused in many gas- and solution-phase reactions by recovering through hightemperature regeneration or any other processes. Additionally, the support provides a large surface area and enhances the catalytic performance by preventing the aggregation of the catalyst particles. A similar type of synthetic strategy can also be extended to a metal-based catalyst such as Pd and Ru, for synthesizing robust catalysts (Shang et al. 2014). The smaller nanoparticles in the range of 0–3 nm accommodate in the support and induce interfacial strain through the catalyst, while larger particle results in distortions at the upper atomic layers in contact with the support and remains strainfree. Nanoparticles with a particle size range 10–15 nm release the interfacial strain through lattice dislocations resulting in disturbed surface energy anisotropy (σ100/ σ111) or shape change (Ha et al. 2014; Gao et al. 2015; Ahmadi et al. 2016).

7.3

Type of Nanoparticles

The nanoparticles can be classified into a variety of categories based on their morphology, shape and size and physiochemical properties. Some of the classifications are listed below.

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Non-metallic Nanoparticles

Graphene, graphitic carbon nitride, carbon nanotubes and fullerenes are some of the major metal-free catalysts. Fullerenes contain allotropic forms of carbon nanomaterial with globular hollow cage. These fullerenes have attracted attention due to its potential properties such as high strength, electrical conductivity, structure, electron affinity and versatility (Astefanei et al. 2015). They constitute pentagonal or hexagonal sp2 hybridized carbon with a diameter ranging from 7.114 to 7.648 nm. Another metal-free nanoparticle is carbon nanotube (CNT) with different forms (viz. elongated or tubular structure (He et al. 2017)) – which are used as conductive materials like metallic or semiconducting material depending upon their diameter (Arjmand et al. 2016; Johannsen et al. 2016). Carbon nanotubes structurally resemble like graphite when they are rolled into sheets in the form of singlewalled nanotubes (SWNTs), double-walled nanotubes (DWNTs) or multi-walled carbon nanotubes (MWCNTs). Having unique structural, electronic, optoelectronic and semiconductor as well as mechanical, chemical and physical properties, carbon nanotubes are widely used in wastewater treatment for removal of heavy metals and dyes (Sadegh et al. 2015). Recently, graphitic carbon nitride (C3N4) with relatively low band-gap energy (Eg ¼ 2.7 eV) and high valence and conduction band positions has drawn attention towards a photocatalytic process for water splitting and hydrogen storage as well as other environmental applications (Liu et al. 2015).

7.3.2

Metallic Nanoparticles

Metal nanoparticles are used as commercial catalysts for application in various process industries like ammonia synthesis, hydrocarbon reforming, oxidation, hydrogenation reactions, photocatalytic water splitting and wastewater treatment. Metal nanoparticles are generally synthesized from metal precursors. They possess a unique characteristic that helps them in active interactions with electromagnetic radiation (photons) (Linic et al. 2015). These metal nanoparticles (e.g. Cu, Ag, Ti, Fe, Zn, Bi and Au) have advanced optical properties with the broad absorption band of visible solar spectrum that enables them for many applications. The morphology of these nanoparticles can be controlled during the synthesis as they play a key role in defining its characteristics. Composites based on metal-organic frameworks (MOFs) as emerging porous material groups have shown to possess unique functional properties (Falcaro et al. 2016). An emerging approach named pseudomorphic replication relies on the preparation of core-shell particles having the ‘functional’ metal nanoparticles as core, and the shell is a feedstock material (Khaletskaya et al. 2013). These metal nanoparticle composites are experimented to highlight their unique characteristics with enhanced performance in the areas of gas adsorption, catalysis, sensing, microelectronics, sequestration, delivery and biomedical applications, fuel production, separation, etc.

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Semiconductor Nanoparticles

Semiconductor nanoparticles are widely used in chemical and process industries due to their intrinsic electronic structure which consists of a filled valence band (VB) and an empty conduction band (CB). Generally, the band-gap energy of these nanoparticles varies in the range of 1.1–5.8 eV. Most of the cases, semiconductors used in photocatalytic reaction are in the form of oxides (e.g. TiO2, ZrO2, ZnO, WO3, MoO3, SnO2, α-Fe2O3, etc.) or in the form of sulphides (e.g. ZnS, CdS, CdSe, WS2, MoS2, etc.). Some of the typical semiconductors used in photocatalytic reaction have a band-gap energy of TiO2, 3.87 eV; ZnO, 4.35 eV; SnO2, 4.3 eV; WO3, 5.24 eV; α-Fe2O3, 5.88 eV; etc. (Wu et al. 2015). When photon energy (hv) matches or exceeds the band-gap energy (Eg), the semiconductor excites an electron (e
) from its valence band (VB) to conduction band (CB) and forms a positive hole (h+) in the valence band. These charge carriers form the fundamental mechanism for the photocatalytic degradation process. The hole (h+) oxidizes organic compounds through the formation of hydroxyl radicals (•OH), while the electrons help to produce superoxide radicals (O2• ) through redox reactions. However, these charge carriers are unstable in an excited state and can easily be recombined resulting in low photocatalytic efficiency (Dinesh et al. 2016a). Therefore, synthesizing a stable, highly photoactive, inexpensive and nontoxic photocatalyst remains a challenge. The band-gap vs. oxidation and reduction potentials is shown in Fig. 7.2.

Eg(Bandgap) Vs Oxidation and Reduction Potential -2.5

-2.0

-3.5

-1.0

ZnS CdS

Si

TiO2 -4.5

0.0

-5.5

1.0

1.1 3.6 3.2

WO3 Fe O 2 3

2.0

-7.5

3.0

-8.5

4.0

O2/O2 . H2/H2O

2.4

2.8

-6.5

ZnO

2.2

3.2

O2/H2O OH./H2O

Fig. 7.2 Band-gap energy vs. oxidation-reduction potential of different semiconductors. (Reprinted from Wu et al. (2015) with permission of Royal Society of Chemistry)

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Ceramic Nanoparticles

Ceramic nanoparticles are inorganic non-metallic solids synthesized at different sintering temperatures followed by successive cooling. The major advantage of these nanoparticles is they are highly stable under extreme conditions of pH and temperature (Roy et al. 2003). This type of nanoparticles is mostly in the amorphous, polycrystalline, dense, porous or hollow form in nature. Therefore, they have attracted the immense attention of the scientific community due to their versatile applications such as catalysis, photo-sensing, photocatalysis and imaging applications (Thomas et al. 2015). Most commonly, silica-based nanoparticles synthesized using cationic surfactants as templates are being used widely for preparation of mesoporous silica as they provide a high surface area, tunable pore size, large pore volume and rich morphology (Han et al. 2011). According to the applications, these parameters can be modified as per the desirability. These nanoparticles are found with different mesostructures (e.g. disordered, wormhole-like, hexagonal, cubic and lamellar mesophases), morphologies (e.g. spheres, hollow spheres, fibres, tubules, gyroids, helical fibres, crystals and many hierarchical structures) and dimensions (nanometre to centimetre) (Wu et al. 2013).

7.3.5

Polymer Nanoparticles

Polymer nanoparticles can be defined as nanosized colloidal organic compoundbased materials which are produced from polymeric materials. Recently, polymeric nanoparticles have gained attraction in the area of polymer science and technology due to their massive advancement in several applications. They come mostly in nanosphere or nanocapsule shape (Mansha et al. 2017). The former is matrix particles whose overall mass is generally solid and the external molecules are adsorbed at the outer boundary of the spherical surface. In the latter case, the solid mass is encapsulated within the particle entirely (Rao and Geckeler 2011). The polymer nanoparticles are readily functionalized and thus find numerous applications in the processes (Abd Ellah and Abouelmagd 2017; Abouelmagd et al. 2016). One of the many potential applications of polymer nanoparticles that have been widely investigated is in the domain of drug delivery: polyhydroxyalkanoates (PHAs) are one class of the PNPs which have been developed as controlled drug release vectors owing to their biocompatibility and biodegradability (Xiong et al. 2010). Polylactide-co-glycolide nanoparticles have been used to prepare peptidebased nanomedicine and even nanoparticle-based gene delivery system mediated by ultrasound (Figueiredo and Esenaliev 2012). Photoacoustic imaging is essential to biology and medicine, and the semiconducting polymeric nanoparticles provide a better photoacoustic imaging. A recent study has reported that modifying the chemical structure of semiconducting polymer nanoparticles with the introduction

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of an electron-deficient structure leads to a considerable photoacoustic amplification of the polymeric nanoparticles (Xie et al. 2017).

7.3.6

Lipid-Based Nanoparticles

Lipid-based nanoparticles have attracted the attention of the scientific community for intravenous applications as they are claimed to be an alternative carrier system for particles. This type of nanoparticles is generally prepared from solid lipids. At the initial stage, liquid-phase lipids are investigated for medical application. But due to many associated disadvantages, they have been immediately replaced by the solid lipids which eventually led to the usage of solid lipid nanoparticles (Ekambaram et al. 2012). These nanoparticles contain lipid moieties and found high efficiency in many biomedical applications. Generally, a lipid nanoparticle is characteristically spherical with a diameter ranging from 10 to 1000 nm. Like polymeric nanoparticles, lipid nanoparticles possess a solid core made of lipid, and a matrix contains soluble lipophilic molecules. Surfactants or emulsifiers stabilized the outer core of these nanoparticles (Rawat et al. 2011). Lipid nanotechnology is a unique field, which focuses on the designing and synthesis of lipid nanoparticles for various applications such as drug carriers, drug delivery and ribonucleic acid (RNA) release in cancer therapy (Mashaghi et al. 2013; Puri et al. 2009; Gujrati et al. 2014). In addition to various pharmaceutical applications, solid lipid nanoparticles are also being investigated for their various applications in treating fungal diseases in plants. In one of the recent studies, it has been observed that encapsulation of fungicides in these nanoparticles help to decrease the cytotoxicity. Thus, the release of the fungicides can also be controlled (Campos et al. 2015).

7.4

Types of Nanoparticles Based on Structure

On the basis of structure, nanoparticles can also be categorized as spherical, tubular (nanorod or nanofiber), nanosheet and interconnected nanocube (Hussein and Shaheed 2015). A schematic diagram of such nanoparticles is shown in Fig. 7.3. A spherical-type catalyst with zero dimensionality possesses large specific surface area and provides high decomposition kinetics of recalcitrant pollutants (Liu et al. 2011). Smaller band-gap one-dimensional fibres or tubes make the charge diffusion faster. Additionally, they also exhibit less recombination and light-scattering properties. The fabrication of self-standing fibres or tubes in the form of arrays is easy in the one-dimensional (1D) nanoparticle, which further can be formed two-dimensional (2D) with smooth surfaces just by assembling or through hierarchical growth resembling the arrays. The two-dimensional (2D) nanoparticle has high adhesion property, whereas three-dimensional (3D) nanoparticle has interconnecting structure and can be easily tuned by doping to facilitate high carrier

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Spherical

Zero-Dimensional

Tubular (nanorod or nanofiber)

One-Dimensional

Nanosheet

Interconnected nanocube

Two-Dimensional

Three-Dimensional

Fig. 7.3 Dimensional variation in nanoparticles. (Reprinted from Hussein and Shaheed (2015) with permission of Omics International)

mobility (Alenezi et al. 2014). The electronic structure is responsible for the optical response of any nanomaterial, and hence the electronic properties of a catalyst are strongly related to its chemical nature of the bonds between the atoms or ions, atomic arrangement and confinement of carriers for nanometre-sized materials. In order to alter the optical properties of a catalyst, a metal or non-metal component is introduced in their electronic structure known as doping. In fact, the nanocatalyst can be turned into one-dimensional (1D), two-dimensional (2D) and three-dimensional (3D) forms just by doping or increasing impurities as shown in Fig. 7.4 (Pan et al. 2017). Doped metal oxide catalysts are widely used in numerous industrial applications whose catalytic activity is improved by substituting cations of the host oxide by another cation which modifies the chemical bonding at the surface of the oxide. Dopants are not limited to only within the metals, but non-metallic substances are also being used for doping of catalysts such as doping of sulphur on graphene used

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Ultrasound Reaction Molecule aggregation Formation of 0D

lf Se

n tio isa n a org

Clusters Aging Nanorods

Formation of 1D

Nucleation Nanofibres

2D

Nanosheets 3D Structure

Fig. 7.4 Controllable synthesis of various forms of nanoparticle under ultrasound irradiation. (Reprinted from Pan et al. (2017) with permission of Royal Society of Chemistry)

as an oxygen reductant (Yang et al. 2012). Nitrogen doping in nanosheets of nanoporous Co3O4 helps to increase the electronic conductivity due to the availability of more active sites. These types of nanosheets are also imparted with oxygen vacancies by N2 plasma and demonstrated to have excellent electrocatalytic activity and efficiency in oxygen evolution (Xu et al. 2017). In another study on gas sensing, it has been reported that ZnO nanostructures doped with bismuth (Bi) and Tin (Sn) oxide can enhance the hydrogen and carbon monoxide responses (Postica et al. 2017), while VOx clusters supported on ceria (Ce) can improve the oxidizing efficiency of methanol than either the individual oxides (Weckhuysen and Keller 2003).

7.5

Synthesis and Applications

Several methods or techniques have been developed for synthesizing of nanoparticles. Some of the most commonly used conventional techniques for catalyst synthesis are sol-gel method, hydrothermal and solvothermal method, micelle and inverse micelle method, direct oxidation, chemical vapour deposition, mechanochemical method, co-precipitation method and emulsion method. One of the most

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recent and advanced techniques for the synthesis of nanoparticles is sonochemical synthesis (Chakma and Moholkar 2015a). Sonochemical synthesis relies on transient cavitation which provides a unique interaction between energy and matter with hot spots inside the microbubbles arising due to pressure variation in the medium. At the moment of transient collapse of cavitation bubbles, extremely high temperature and pressure (approximately 5000 K and 1000 bar) are produced with heating and cooling rates of more than 1010 K s-1 (Bang and Suslick 2010). One interesting fact of sonochemical synthesis is that there is no direct interaction between the chemical species and ultrasound at a molecular level. In addition to sonochemical synthesis, ultrasonic spray pyrolysis is also being employed to synthesize materials. Ultrasonic spray pyrolysis involves the phenomenon of nebulization which generates mist due to the passage of ultrasound through a liquid medium and its impingement on the liquid-gas interface. Sonochemical synthesis provides several advantages over the other conventional methods, namely, one-pot synthesis, economic energy requirements, use of nonhazardous chemicals, mild reaction conditions, etc. This technique has gained popularity due to the above-mentioned advantages. Numerous catalysts have been prepared using the sonochemical route and used for the environmental application. Some of these are MnO2 microspheres and nanostructured Fe-SiO2, Cu2O, CdWO4, MnFe2O4, Fe-TiO2 and ZVI-Bi2O3 (Dinesh et al. 2016a; Zhang et al. 2010; Suslick et al. 1995; Bhosale and Bhanage 2016; Hosseinpour-Mashkani and Sobhani-Nasab 2016; Goswami et al. 2013; Dinesh et al. 2016b). The application of ultrasound can also be used for doping and ferrite synthesis (Chakma and Moholkar 2015b). In this chapter, we have focused on the sonochemical synthesis of pure and doped catalyst and magnetic ferrite nanoparticles and their applications for wastewater treatment. For synthesizing of Fe(0)-doped TiO2 catalyst (Dinesh et al. 2015; Reddy et al. 2016), green tea extract (polyphenols) was used as reducing agent for Fe. The major advantage of this extract is it acts as a stabilizer, is nontoxic, and has the capability to reduce metals easily. The desired amount of fresh green tea leaves (2 g) was brewed in 100 mL of Millipore water and allowed for 10 min at 80oC. The solution was cooled down and filtered. Then FeSO4 (13.9 g) was added to it and sonicated for 0.5 h at 20 kHz and a power of 500 W. A black-coloured Fe precipitate was collected and filtered followed by drying in a hot air oven at 105oC for 24 h. Then 1 mL of titanium tetraisopropoxide was added dropwise to a mixture of water (20 mL) and ethanol (1 mL) with the desired amount of Fe powder (viz. 0.1, 0.2, 0.3 and 0.4 g). The whole solution was kept under ultrasonic irradiation for 0.5 h, and the obtained solid powder was collected and dried at 105oC for 24 h. In another study of Fe-doped ZnO, a similar method has been adopted for the preparation except for the addition of Fe in the ZnO mixture (Chakma and Moholkar 2015b; Dinesh et al. 2016c). Magnetic nanoparticles also attracted the attention due to their potential in degradation and ease of separation. In a study, magnetic nanoparticles (viz. Zr-ferrite) have been synthesized through sonochemical route using zirconium oxychloride and iron (II) acetate with a molar ratio of 1:2 at controlled pH (Chakma

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and Moholkar 2015a; Das et al. 2018). The solution was allowed for 2 h under ultrasonic irradiation of 20 kHz and 40% amplitude with the 5s ON-OFF mode at a temperature of 30oC. The synthesized nanoparticles have been filtered and calcined at different temperatures and characterized using powder X-ray diffractometer (XRD), scanning electron microscope (SEM), transmission electron microscope (TEM), diffuse reflectance spectrometer (DRS), energy-dispersive X-ray (EDX), vibrating sample magnetometer (VSM), etc. Finally, the catalytic activity of the synthesized nanoparticles has been determined using advanced oxidation processes (AOPs) for degradation of recalcitrant organic contaminants in wastewater. In this study, we have investigated decolourization/degradation of different textile dyes (azo and non-azo), viz. Acid Red B, Methylene Blue and Bismarck Brown G, using sono-photocatalysis hybrid advanced oxidation processes (Chakma and Moholkar 2015b; Dinesh et al. 2015). Also, the synergistic effect has been studied by analysing the results obtained in sono-hybrid advanced oxidation processes and conventional methods.

7.5.1

Discussions

The synthesized nanoparticles have been used in photocatalysis process for decolourization/degradation of recalcitrant organic textile dyes. From the experimental results, the following salient features have been observed. 1. The decolourization results of Acid Red B (ARB) and Methylene Blue (MB) using Zr-ferrite and TiO2 have shown interesting results (Chakma and Moholkar 2015a). It was found that Zr-ferrite showed a lesser decolourization photocatalytic activity than TiO2 even though the band-gap energy of Zr-ferrite is lesser than TiO2. This phenomenon is ascribed by the fact that the Zr-ferrite nanoparticles have haematite phase (α-Fe2O3) on its surface. This α-Fe2O3 phase acts as a recombination centre for the electron-hole pairs (e––h+), thereby reducing the photoactivity. However, when the Zr-ferrite catalyst was used in the presence of an oxidant such as H2O2, dye decolourization was increased by 88% in 60 minutes which can be explained by the Fenton-like reaction occurring due to the in situ generation of hydroxyl radicals as shown in Fig. 7.5. These radicals are capable of dye degradation. Thus the recombination of radicals was compensated by the occurrence of the heterogeneous Fenton-like reaction. It was also conjectured that adsorption is a contributing mechanism in the dye decolourization as the radicals generated by the photochemical effect mainly degrade the adsorbed dye molecules. Moreover, the sonochemically synthesized Zr-ferrite has a capacity of high adsorption which endorses the Fenton and photoactivity. Hence, Zr-ferrite is found to be an efficient photocatalyst for hybrid advanced oxidation processes compared to TiO2. Also, the degradation results of textile dyes explored that the Zr-ferrite has the higher contribution of decolourization through Fenton-like reaction than photocatalysis reaction.

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Fig. 7.5 Decolourization and degradation mechanisms of textile dyes with TiO2 and Zr-ferrite (ZrFe2O5) catalysts under ultrasound irradiation. (Reprinted from Chakma and Moholkar (2015a) with permission of Royal Society of Chemistry)

Additionally, the most effective protocol for degradation is found to be sonophoto-Fenton process where both Fenton and photoactivities of Zr-ferrite are utilized efficiently. 2. In another study, Bi2O3 was successfully synthesized and utilized for treating synthetic wastewater (Dinesh et al. 2016b). In this study, an investigation of dye degradation using metal ions doped with Bi2O3 (i.e. Fe-Bi2O3) was performed, and it revealed that doping enhances the degradation efficiency significantly. The experimental results showed that 99% degradation of dye could be achieved when Fe-doped Bi2O3 catalyst was used in the presence of H2O2. Basically, the doping of metal ions increases the photocatalytic activity by increasing the light absorption range and enables the generation of more electron-hole pairs which contribute to higher degradation percentage through the generation of free radicals. In one of the studies, photocatalysis process has been investigated using zero valent iron-doped TiO2 for the decolourization of Bismarck Brown G dye (Dinesh et al. 2015). The activity of Fe-doped TiO2 was studied under different experimental protocols sonolysis, photocatalysis and sono-photocatalysis. The maximum decolourization percentage of approximately 99% was achieved with sono-photocatalysis process when Fe-doped TiO2 was used. This was attributed to the higher adsorption of ultraviolet light by TiO2 due to doping of Fe0 that leads to enhanced photocatalytic activity. Also, doping causes narrow band-gap in the catalyst which enhances the electron-hole pair separation. As a result, the rate of kinetic increases and enhances the dye decomposition. Moreover, the presence of Fe on the catalyst surface accelerates the Fenton-like reaction through the in situ generation of H2O2 from the transient collapse of cavitation bubbles and produces more amounts of free radicals in the reaction solution, thus improving degradation efficiency.

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Another similar study on pure and Fe-doped ZnO as photocatalysts was investigated for cationic and anionic dye degradation to assess the effect of doping on the catalysts (Chakma and Moholkar 2015b). The dye degradation experiments have been performed using four different experimental protocols (viz. sonolysis, sonocatalysis, photocatalysis and sono-photocatalysis). The maximum degradation was seen with the hybrid advanced oxidation process of sono-photocatalysis, i.e. combination of sonolysis and photocatalysis. The degradation efficiency was significantly high with Fe-doped ZnO catalyst. In this case, Fe-ion helps to enhance absorption range through a redshift of band-gap of the semiconductor. Also, Fe-ion functions as a photo-generated hole and trap and inhibits the recombination of electron-hole (e– – h+) as illustrated by the following equations (Chakma and Moholkar 2015b): ZnO + hv ! e
+ h+ 2þ Fe3þ þ e
cb ! Fe Fe2þ þ O2, ðadsÞ ! Fe3þ þ O2•
4þ Fe3þ þ hþ vb ! Fe 4+
Fe + OH ! Fe3+ + •OH

(Charge pair generation) (Electron trap) (Electron release) (Hole trap) (Hole release)

In the above reaction mechanism, Fe3+ ion acts as electron holes and trap and forms Fe2+ and Fe4+. These ions are comparatively unstable and react with O2 and
OH ions adsorbed onto the surface of the catalyst and produce •OH and O2•
radicals. These radicals react with organic molecules present in the bulk solution, thus enhancing the rate of degradation. However, sometimes the excess amount of Fe-ions (greater than 2 wt.%) shows adverse effects on photocatalytic activity as the availability of traps is decreased due to the following reaction, Fe4+ + e
! Fe3+, and Fe3+ ions act as a recombination centre for electrons (e-) and holes (h+) as 3þ follows: Fe3+ + e
! Fe2+ and Fe2þ þ hþ vb ! Fe . This causes a reduction in the photocatalytic activity and thus reduces the rate of degradation. Since in our investigation the concentration of Fe-ions was less the optimum value, the adverse effect of Fe-ion was insignificant. It was also seen that addition of ZnO either in pure or doped form gave a marked increase in the extent of degradation, but the investigation also found a negative synergy among the individual processes of sonolysis and photocatalysis. This is attributed to the possibility of desorption of dye molecules from the photocatalyst surface in case of the ultrasound-assisted process – which reduced the chances of dye-radical interaction as shown in Fig. 7.6. It was also revealed that in the absence of a light source, the photocatalytic activity of ZnO and Fe-doped ZnO remained inactive and sonoluminescence wasn’t able to activate the catalysts.

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Fig. 7.6 Synergetic effect in sono-photocatalysis process for degradation of dyes. (Reprinted from Chakma and Moholkar (2015b) with permission of Elsevier)

7.6

Synergetic Effect

The synergistic effect on the degradation of recalcitrant pollutants was estimated by comparing the degradation percentage obtained after treatment using the individual technique (such as sonolysis, sonocatalysis, photocatalysis) and hybrid technique (i.e. sono-photocatalysis). The synergy index or the percentage of synergy is basically to evaluate the interaction of the individual techniques, viz. sonolysis, sonocatalysis and photocatalysis. The equations used for calculation of the synergistic effects are as follows (Eq. 7.2):
KðUSþUVþCatalystÞ
KðUSÞ þ KðUVþCatalystÞ
Synergy Index ¼  100 KðUSÞ þ KðUVþCatalystÞ

ð7:1Þ

Synergy Effect ð%Þ       Degradation obtained Degradation Degradation
þ with hybrid-AOPs with Sonolysis with individual AOP     ¼ Degradation Degradation þ with Sonolysis with individual AOP ð7:2Þ where K is the rate of reaction for degradation of organic molecules in the presence and absence of ultrasound. In all the studies, a positive synergetic effect has been observed, while sonolysis and photocatalysis have been combined. The individual techniques of sonolysis and photolysis can degrade Acid Red B approx. 33% and 8%, respectively, while Acid Red B degradation was achieved approximately 93%

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in hybrid advanced oxidation processes of sono-photocatalysis (Chakma and Moholkar 2015b). The first-order rate constant for sono-photocatalysis was 3  10
3 s-1 – which was almost twofold higher than the individual methods of sonolysis (1.20  10
4 s-1) and photocatalysis (1.87  10
5 s-1). Also, it was observed that decolourization percentages in the presence of ultrasound were 46% and 44% with the application of individual catalyst Fe and Bi2O3, respectively (Dinesh et al. 2016b). But the combination of these two catalysts (i.e. Fe-Bi2O3) showed a constructive synergy with enhancement of decolourization up to 77%.

7.7

Conclusion and Overview

The major contribution of ultrasound and cavitation effect in most of the sono-hybrid processes investigated by us seems to be physical, and the sonochemical effect (i.e. generation of highly oxidizing radicals) is not much significant for decolourization/degradation. However, the sono-hybrid advanced oxidation processes showed positive synergy due to proper dispersion and utilization of •OH radicals generated through heterogeneous Fenton-like reaction, photocatalysis process or cavitation. It was also found that the physical effects of cavitation have an adverse effect on heterogeneous coupling advanced oxidation process such as sonocatalysis and sono-photocatalysis. This is due to the shock waves generated during the transient collapse of microbubbles, which causes desorption of organic molecules from the surface of the catalyst. Hence, even after the positive synergy caused by the physical effect of ultrasound, the shock waves are found to be discrete and highly energetic phenomena that impact the degradation process adversely, though the shock waves have been reported to be beneficial in many catalytic reactions due to the cleaning of catalyst surface and prevent the catalyst poisoning or de-agglomeration of catalyst nanoparticles providing a large number of available active sites on catalyst surface, thus increasing the chemical reaction at the surfaces.

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Chapter 8

Magnetic-Based Photocatalyst for Antibacterial Application and Catalytic Performance Sze-Mun Lam, Jin-Chung Sin, and Abdul Rahman Mohamed

Contents 8.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.2 Magnetic-Based Photocatalysts in Inactivation of the Microorganism . . . . . . . . . . . . . . . . . . 8.3 Factors Affecting the Photocatalytic Bacterial Inactivation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.3.1 Effect of Magnetic-Based Photocatalyst Concentration and Light Intensity . . . . 8.3.2 Nature of Microorganism . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.3.3 Solution pH of Magnetic-Based Photocatalyst Suspension . . . . . . . . . . . . . . . . . . . . . . 8.3.4 Initial Bacterial Concentration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.3.5 Physiological State of Bacteria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.4 Proposed Mechanism for Bacteria Disinfection by the Magnetic-Based Photocatalyst . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.5 Using Magnetic-Based Catalyst in Photocatalytic Abatement of Organics . . . . . . . . . . . . . . 8.6 Photocatalysis for the Simultaneous Treatment of Bacteria and Organics . . . . . . . . . . . . . . . 8.7 Conclusion and Future Prospects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract Photocatalysis using magnetic-based photocatalyst in water and wastewater treatment offers a green and effective technique for the disinfection of harmful microorganisms along with its unwanted chemical pollutants. Introduction of magnetic materials to the catalytic material composites allows for the convenient magnetic separation, hence providing more economical, effective and environmentally friendly water decontamination processes. In this work, we disclosed a brief review on the effect of various magnetic-based photocatalyst nanomaterials on the S.-M. Lam (*) Department of Environmental Engineering, Faculty of Engineering and Green Technology, Universiti Tunku Abdul Rahman, Kampar, Perak, Malaysia e-mail: [email protected] J.-C. Sin Department of Petrochemical Engineering, Faculty of Engineering and Green Technology, Universiti Tunku Abdul Rahman, Kampar, Perak, Malaysia A. R. Mohamed School of Chemical Engineering, Universiti Sains Malayisia, Nibong Tebal, Pulau Pinang, Malaysia © Springer Nature Switzerland AG 2020 Inamuddin et al. (eds.), Nanophotocatalysis and Environmental Applications, Environmental Chemistry for a Sustainable World 30, https://doi.org/10.1007/978-3-030-12619-3_8

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application of the photocatalytic disinfection and degradation processes. The influencing factors including photocatalyst concentration and light intensity, nature of microorganism, solution pH, initial bacterial concentration and physiological state of bacteria of such processes were presented along with the disinfection mechanisms. The mechanism of magnetic-based photocatalyst was mainly ascribed to the surface generation of reactive oxygen species as well as free metal ion formation. Additionally, the potential utilization of the magnetic-based photocatalyst as visible light nanomaterials was discussed, and their magnetic recoveries were reviewed. It was worth noting that the combined disinfection and decontamination processes will greatly improve the use of magnetic-based photocatalysts as potential alternative to conventional methods of water purification. Keywords Magnetic · Photocatalyst · Nanomaterial · Antibacterial · Organic pollutant · Photocatalysis · Visible light · Mechanism · Recycling · Wastewater treatment

8.1

Introduction

With the advent of rapid industrialization and world population growth, environmental pollution exacerbates and produces large numbers of organic pollutants, and this phenomenon was a severe environmental issue. A myriad of toxic chemicals such as pesticides, drugs, dyes, etc. have been identified in the surface and groundwater, being their removal mostly studied in the last several decades (Rajeshwar et al. 2008; Lam et al. 2012; Sin et al. 2012; Qi et al. 2017). Additionally, there were also concerns regarding the presence of harmful microorganisms that can cause epidemic outbreaks (Vortmann et al. 2015). Therefore, it was highly necessary to develop low-cost, environmentally friendly and highly effective techniques for microbial-contaminated wastewater treatment. A wide variety of treatment technologies including adsorption, coagulationflocculation, membrane processes, electrochemical techniques and biological process with varying levels of success has garnered a dramatic progress in the scientific community (Carp et al. 2004; Busca et al. 2008; Alhaji et al. 2016; Zhang et al. 2016; Garcia-Segura et al. 2018). Of major interest, semiconductor oxide-based photocatalysis, typically a free radical or light-mediated disinfection and destructive technology, has currently appeared to be the most promising approach in the wastewater detoxification processes. The photocatalysis principle was according to the light absorption by semiconductor photocatalyst to initiate the electrons (e–) from the valence band to the conduction band, leaving behind the holes (h+) in the valence band. The separated e–
h+ subsequently underwent redox reactions to produce reactive species including hole (h+), hydroxyl (•OH), superanion oxide (O2•
) radicals and hydrogen peroxide (H2O2) to oxidize or abate organic pollutants and microorganisms (Gaya and Abdullah 2008; Qi et al. 2017).

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A broad spectrum of semiconductors ranging from large to small band gaps has been extensively investigated for this purpose, for example, TiO2, ZnO, Nb2O5, WO3 and CdS (Carp et al. 2004; Gaya and Abdullah 2008; Lam et al. 2014). In particular, TiO2 has been used most widely owing to its superior merits including high oxidation efficiency, nontoxicity, availability and low cost. Nevertheless, some shortcomings have prevented the practical application of TiO2-based photocatalysis technologies in large-scale water treatment plants. Concerns limiting its practicality included the high e–
h+ recombination, low solar sensitivity and problematic separation of the TiO2 nanoparticles from large volumes of treated water. Therefore, various types of modifications had been developed to improve the photocatalytic efficiency (Lam et al. 2017; Reddy et al. 2017). One promising approach to surmount the issue of material recovery stage was the use of a magnetic-based photocatalyst in the wastewater treatment processes. As compared to the traditional methods such as filtration, sedimentation, centrifugation or membrane separation, the magnetic separation can provide several merits used for the material recovery. This included ease of operation, high selectivity, energy saving and effective material recovery from the treated water (Gomez-Pastora et al. 2014). Most importantly, it was a non-invasive technology since the magnetic field can pass through some transparent materials with no physical interaction on the magnetic-based photocatalysts, avoiding the pollution of the treated water. This magnetic separation was also insensitive to influencing parameters including temperature, solution pH and inorganic ion, permitting the process in a wide spectrum of working situations (Pamme 2006; Luo and Nguyen 2017). Lastly, when magneticbased photocatalysts were used, the particles separation can be simply carried out using the external magnetic field application, consequently enabling the continuous runs of the wastewater treatment (Ge et al. 2017). In summary, despite these promising advantages, there were some vital technical points that require to be addressed in order to allow the utilization of magnetic-based photocatalysts at the practical stage. Consequently, it was essential to review the structural and photocatalytic attributes of these new photocatalysts, evaluate recent advances in the area and analyse their possible applications in the microbial disinfection and wastewater treatment processes.

8.2

Magnetic-Based Photocatalysts in Inactivation of the Microorganism

Many scientific publications employing the novel magnetic-based photocatalysts for bacterial inactivation in the water and wastewater treatment have been reviewed (McEvoy and Zhang 2014; Reddy et al. 2017; Qi et al. 2017; Ganguly et al. 2018). In most cases, magnetic-based photocatalysts can be classified into three principal groups. The first group of the magnetic-based photocatalyst particles was based on a core-shell structure, comprising of a composite constructed by both magnetic and

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photocatalyst materials. The shell was commonly constructed by photocatalysts including TiO2, Ag3VO4, AgBr, etc. (Jing et al. 2016; Huang et al. 2017). The inner core was typically comprised of magnetic iron and cobalt oxide elements including magnetite (Fe3O4) and cobalt ferrite (CoFe2O4) and offered the magnetic properties of the composites. In literature, within the magnetic-based photocatalysts, the coating material including the mesoporous SiO2 has also been found for such configurations (Chanhom et al. 2017). Moreover, the addition of doping noble metals such as Ag, Au, etc. has been observed (Bokare et al. 2014; Li et al. 2014). The second group was related to the construction of magnetic binary oxides including TiO2 or Fe2O3 (Kanchanatip et al. 2014; Naeimi et al. 2015). The last group was associated with the ferrite oxide nanomaterials such as Mn1–xNixFe2O4 (Jesudoss et al. 2016). These nanomaterials have been evaluated as magnetic-based photocatalysts showing good bacterial disinfection activities under visible light irradiation. With regard to the microorganisms that can be inactivated by magnetic-based photocatalysts, Table 8.1 depicts some typical examples of the articles relating to the inactivation of various microorganisms by magnetic-based photocatalysts. The information reported included the preparation technique fabricated on its formulation, the properties of the photocatalyst, the operating conditions and the important findings observed by their applications. As can be seen, bacteria were the most widely employed microorganism in the photocatalytic disinfection studies.

8.3

Factors Affecting the Photocatalytic Bacterial Inactivation

In the photocatalytic inactivation of bacteria, the following are the physiochemical factors which affect the process: magnetic-based photocatalyst concentration, light irradiation time and its intensity, initial bacterial concentration, physiological state of bacteria, solution pH, oxygen concentration, etc. These parameters will be considered one after the other as they influenced the inactivation processes of the bacteria in the wastewaters.

8.3.1

Effect of Magnetic-Based Photocatalyst Concentration and Light Intensity

Researchers have revealed an increase in antimicrobial activity by increasing the photocatalyst concentration (Shang et al. 2013). This concentration governed by the microorganism used, type of light source and the reactor geometry is utilized in different studies. Nevertheless, further increasing the photocatalyst concentration, it was difficult to maintain homogeneous the suspension due to the agglomeration of

γ-Fe2O3@SiO2@TiO2– Ag CoFe2O4/Ag/Ag3VO4

Au-Fe3O4@mTiO2

Fe3O4@PTAA-LDHs

E. coli

E. coli

E. coli

E. coli

S. aureus

S. aureus

E. coli

TiO2/silica/Fe2O3

Silica-coated Fe3O4

Hydrothermalsol-gel Hydrothermal

Ag/AgCl@Fe2O3

E. coli

S. aureus

Solvothermal

Graphene oxide/gC3N4

E. coli

Thermal decompositionmicroemulsion

Co-precipitation

Sol-gelhydrothermal Solvothermalultrasonication

Ultrasonication

Solvothermal

Ag/AgBr@ CoFe2O4

E. coli

Method Hydrothermal

Photocatalyst Ag@TiO2-Fe2O3

Microorganism E. coli

Size ~108.5 nm, sat. mag. 7 emu/g

Size 20 nm, sat. mag. 9.2 emu/g, band gap 2.35 eV

Size 230 nm

Size 0.5–1 μm, sat. mag. 1.8 emu/g Surf. area 66.1 m2/g

Size 2–50 nm, sat. mag. 5.8 emu/g Size ~300 nm

Band gap 2.51 eV

Size ~12 nm, sat. mag. 29.4 emu/g

Characterization Size 96 nm, sat. mag. 9.2 emu/g

Table 8.1 Photocatalytic inactivation of microorganisms via magnetic-based photocatalysts

UV light

[E. coli] ¼ 1.5 x 105 cfu/mL, [photocatalyst] ¼ 0.02 g, UV light

[Photocatalyst] ¼ 10 mg, UV light [E. coli] ¼ 1 x 107 cfu/mL, [photocatalyst] ¼ 0.8 mg/mL, solar light

[E. coli] ¼ 1/(1  103), vis light

vis light

Working condition [E. coli] ¼ 1 x 107 cfu/mL, [photocatalyst] ¼ 50 μg/L, UV light [E. coli] ¼ 1 x 107 cfu/mL, [photocatalyst] ¼ 5 mg, vis light [E. coli] ¼ 100 μg/mL, [photocatalyst] ¼ 100 μL, vis light [E. coli] ¼ 1/(2  104), vis light 100% (30 min) 97% (24 hrs) 100% (12 min) 89.3% (60 min) 99.99% (140 min) 99.99% (140 min) 100% (105 min) 100% (120 min) NA

(continued)

Chanhom et al. (2017)

Naeimi et al. (2015)

Xu et al. (2013) Cui et al. (2013) Jing et al. (2016) Li et al. (2014) Shang et al. (2013)

Sun et al. (2017)

Huang et al. (2017)

100% (20 min) 97.9% (120 min)

References Bokare et al. (2014)

Performance 95% (120 min)

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NA not available

E. tarda

Microorganism P. damselae subsp. piscicida A. hydrophila

Photocatalyst Au-TiO2/Fe3O4

Table 8.1 (continued)

Method Co-precipitationsol-gel

Characterization Size 14–40 nm, sat. mag. 2.7 emu/g, band gap 2.35 eV

Working condition [Photocatalyst] ¼ 2.7 mg, vis light ~50% (90 min) ~38% (60 min)

Performance ~100% (90 min)

References Kanchanatip et al. (2014)

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particles, which decreased the number of active sites. At the same time, this phenomenon restricted the light penetration into the suspension and thus impaired the photocatalyst activities. To investigate the influence of the catalyst concentration on the bacterial inactivation, Shang et al. (2013) have conducted photocatalytic disinfection performance of poly-(3-thiopheneacetic acid)-coated Fe3O4@LDHs (Fe3O4@PTAA-LDHs) against E. coli and S. aureus under solar light. Their results showed that the antibacterial rate improved with increasing of the concentration of the catalyst. Both the antibacterial rates attained 99.99% when the Fe3O4@PTAALDHs concentration was 0.8 mg/mL. However, there were constant antibacterial rates above 0.8 mg/mL, which can be associated to the saturating catalyst concentration. On the other hand, irradiation time and light intensity also acted as an important role as it induced charge carrier separation on the photocatalyst nanoparticles, resulting in cell damage. It was reported that increasing the irradiation time and light intensity affected the photocatalytic bacterial inactivation. Shen et al. (2015) studied the effect of visible light irradiation time and light intensity on the photocatalytic inactivation of three types of microorganisms, S. aureus, E. coli and C. albicans, via graphene oxide-ZnFe2O4 composites. In their reports, the inhibition zone diameter of the sample reduced clearly without the presence of light (0 lx, 24 h), which indicated that the irradiation was an important criterion. When prolonging irradiation time (150 lx, 24 h) to (150 lx, 36 h), the antibacterial activity of composites increased slightly, illustrating the oxidation of reactive oxidative species on the bacteria. On the other hand, when increasing light intensity (150 lx, 24 h) to (300 lx, 24 h), the antibacterial activity of composites has not dramatically changed. They attributed this reason to the amount of reactive oxidative species generated by graphene oxide-ZnFe2O4 composite having no impact on composite antibacterial activity as the oxidation of reactive oxidative species on the bacteria may be a chain reaction. The composites also revealed to display different antibacterial activities on three bacteria, and the inhibition zone diameter of the sample on C. albicans was the greatest of three bacteria. They reported that the cell wall structure of C. albicans can be destroyed easily. Compared to that of S. aureus, the E. coli has the lipid bilayer, but the thickness of its cell wall was thinner than that of S. aureus and the crosslinking degree of its peptidoglycan with net structure was poor. Hence, the cell wall structure of E. coli was easily destroyed by the composites.

8.3.2

Nature of Microorganism

The structure and physiology of the microorganisms are key issues in governing the efficacy of disinfection by magnetic-based photocatalyst-related photocatalytic process. It was well known that bacteria have features in common such as the cytoplasm, cell membrane and cell wall. The cell wall surrounded the bacterial inner membrane which comprised of a uniform peptidoglycan layer composing of sugars and amino acids (Sirelkhatim et al. 2015). Destruction of bacteria was highly reliant

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on the bacterial strains. The outer cell wall plays a key role in the tolerance or susceptibility of the bacteria in the presence of magnetic-based photocatalyst particles. For instance, Naeimi et al. (2015) stated that the higher deactivation efficiency was obtained for gram-negative bacteria (E. coli) as compared to that of grampositive (S. aureus) using Fe3O4 encapsulated-silica sulfonic acid nanoparticles. They explained the differences in the cell wall structure inherent in the gramnegative and gram-positive bacteria. The gram-positive and gram-negative bacteria have the same internal but very different external structures. The gram-positive bacterium has a thick peptidoglycan layer that contained teichoic and lipoteichoic acids. In contrast, the gram-negative bacterium has a thin peptidoglycan layer and an outer membrane that contained lipopolysaccharide, phospholipids and proteins. They concluded that the bacteria photoinactivation rate was determined not only by cell wall thickness but also by the morphology of cell envelope and resistance of the outer membrane to the reactive oxidative species generated on the catalyst surface.

8.3.3

Solution pH of Magnetic-Based Photocatalyst Suspension

The solution pH affected the amphoteric behaviour of magnetic-based photocatalysts and nature of microorganisms found in the medium. In literature, there were only a few studies on the effect of solution pH on the antibacterial activity of magnetic-based photocatalyst particles. Ma et al. (2015) investigated the survival of E. coli under different solution pH at 5, 7 and 9 using Fe3O4-TiO2 nanosheets in the presence of solar light. It was found that the removal efficiency of bacteria declined with increasing the pH values. The reason given by them was the E. coli can be simply adsorbed at low solution pH and the negatively charged bacterial membrane can be responsible for this favoured reaction.

8.3.4

Initial Bacterial Concentration

The initial bacterial population is a crucial factor in determining the photocatalytic water disinfection efficiency. For instance, the photocatalytic inactivation at higher initial E. coli concentrations confirmed that a long time was needed for bacterial inactivation. This was because under the circumstances of dead cells and the excreted intracellular compounds, they might compete for the reactive oxidative species or screen the light penetration shielding the bacteria that are left in the suspension. In literature, there was no study on the effect of initial bacterial concentration on the antibacterial activity of semiconductor nanoparticles. Hence, the

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researchers on this area have urged to study the relationship on the effect of initial bacterial concentration in the photocatalytic water disinfection.

8.3.5

Physiological State of Bacteria

Several researchers have studied on the effect of the growth phase of E. coli on UV disinfection (Lewis and Burt-Maxcy 1984; Kadavy et al. 2000). Their findings revealed that the bacteria in the stationary state were inactivated in a lesser extent than that of an exponential growth state. The reason given for this was that the stationary state response to the environmental changes involved the synthesis of a set of proteins conferring the E. coli resistance to UV light oxidation, heat shock, acidity, hyperosmolarity and nutrient scarcity (Murno et al. 1995; Child et al. 2002). Therefore, bacterial inactivation not only is subjected to the bacteria type but also on their growth phases.

8.4

Proposed Mechanism for Bacteria Disinfection by the Magnetic-Based Photocatalyst

The photocatalytic formation of reactive oxidative species has been the main contributor to the bacterial inactivation of many metal oxides. Several works demonstrated the reactive oxidative species production as the principal mechanism responsible for magnetic-based photocatalyst bacterial inactivation activity (Shang et al. 2013; Li et al. 2014; Ma et al. 2015; Padhi et al. 2017; Xia et al. 2018). Ma et al. (2015) examined the dominance of active species during Fe2O3-TiO2 nanosheets (Fe2O3-TNS) photocatalytic disinfection by employing scavengers including Cr (VI) for the electron, isopropanol for •OH radicals, sodium oxalate for hole and Fe (II) for H2O2 (Fig. 8.1a). Their results indicated that the disinfection process led to reactive oxidative species release on the Fe2O3-TNS surfaces under solar light and the reactive oxidative species release led to the lethal bacterial injury. They explained the reactive oxidative species production, in particular, hole and H2O2 on the Fe2O3-TNS surface, via a correlation between photon reactions and the bacterial inactivation activity (Fig. 8.1b). Shang et al. (2013) in their works – poly-(3-thiophene acetic acid)-coated Fe3O4@LDHs with higher disinfection against E. coli in the photocatalytic process – reported that the antibacterial activity can be attributed to the significant disorder in the membrane permeability by •OH radicals. In their reports, the generation of •OH radicals has been proved using the fluorescence method. They also observed the damage of bacteria by •OH radicals using transmission electron microscopy analysis. Their findings showed that the cell was no longer intact and the outer cell membrane was destructed after photocatalytic disinfection, causing interior component leakages. Similar damage reaction was also

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Fig. 8.1 (a) Solar light photocatalytic antibacterial activity against 107 CFU/mL E. coli in the presence of 100 μg/mL Fe2O3-TiO2 nanosheets using different scavengers and (b) photocatalytic disinfection mechanism of E. coli over Fe2O3-TiO2 nanosheets. (Adapted from Ma et al. (2015))

reported by them in the S. aureus suspension. However, the exact antibacterial mechanism of magnetic-based photocatalysts was still not clear and may contain the reactive oxidative species generation, the release of metal ions and damage to cell membranes (Cui et al. 2013; Kanchanatip et al. 2014; Shen et al. 2015; Pant et al. 2017). Shen et al. (2015) studied visible light photocatalytic inactivation of three types of microorganisms, S. aureus, E. coli and C. albicans, via graphene oxideZnFe2O4 composites. They proposed that enhanced graphene oxide-ZnFe2O4 composite antibacterial activity was due to two antibacterial mechanisms: (i) reactive oxidative species antibacterial (the reactive oxidative species as strong oxidizing agents are toxic to cells as they can destruct the cellular constituents such as DNA, lipid and proteins) and (ii) nano-antibacterial (the metal ion can join with bacterial cell-membrane protein (such as the S
S bond), altering the enzyme’s normal operation and preventing the cell’s normal metabolism). They added that the composites can also fragment bacteria slowly by hampering the energy and substance transfers between the outer environment and bacteria. In another experiment by Cui et al. (2013) on the physicochemical characterization and antibacterial tests on γ-Fe2O3@SiO2@TiO2-Ag nanocomposites, it was suggested inactivation by nanocomposites against E. coli was attributed mainly to the Ag+ ion dissolution and the photocatalytic behaviours of the TiO2-Ag. Three mechanisms have been postulated to elucidate the bacterial inactivation activities of Ag: (i) Ag nanoparticles can adhere to the bacterial cell membrane, resulting in the structural damages or malfunction. The membrane contained several sulphurbearing proteins, which could be the favourable sites for Ag particle adhesion owing to the sulphur and Ag affinity. (ii) Ag is a strong nucleic acid binder, and it can produce complexes with DNA and RNA. Such complexation can result in the DNA condensation and replication loss ability. Respiratory enzyme activity can also be malfunctioned because of Ag binding with thiol groups in the proteins, and (iii) free reactive oxidative species formation on Ag particles has been found to lead to antibacterial activities.

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Fig. 8.2 (a) Field emission scanning electron microscopy image of bacterial inactivation results on E. coli for TiO2, pure Ag3PO4 and Ag3PO4/TiO2/Fe3O4 samples, field emission scanning electron microscopy images of E. coli (b) before and (c) after being killed on Ag3PO4/TiO2/Fe3O4 under visible light irradiation and (d) enlarged view of a damaged E. coli cell. (Adapted from Xu et al. (2014))

The study of Xu et al. (2014) also showed that the differences in zone of inhibition diameters suggested that the Ag3PO4/TiO2/Fe3O4 nanocomposite had superior antibacterial effects against E. coli as compared to that of pure Ag3PO4 (Fig. 8.2a). They proposed that the antibacterial test of nanocomposite against E. coli was attributed mainly to the released Ag+ ions. They went further to analyse the bacterial cells before and after the antibacterial treatment using the field emission scanning electron microscopy analysis. As displayed in Fig. 8.2b, the intact E. coli cells demonstrated a well-defined cell wall. After the E. coli damage on the nanocomposite by visible light irradiation, clear destruction was found with the outer membrane being collapsed, indicated as circles in Fig. 8.2c. The Ag+ ions were believed to interact with the main components of the bacterial cell containing the cytoplasmic DNA, bacterial proteins, plasma membrane, peptidoglycan cell wall and led to the bacterial membrane ruptured. The magnified image of E. coli (Fig. 8.2d) confirmed that the outer membrane of the bacterial cell was damaged as well as the

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Fig. 8.3 (a) Zone of inhibition formed by Mn1–xNixFe2O4 nanoparticles against the bacterial strains. (b) Inhibitory effect of the nanoparticles and the standard streptomycin towards bacteria. Schematic illustration of (c) light irradiated on the nanoparticles and (d) inhibition activity of bacterial growth by the nanoparticles. (Adapted from Jesudoss et al. (2016))

ends of the cell after the antibacterial treatment. Similarly, according to Kanchanatip et al. (2014), the visible light photocatalytic bactericidal activities of Ag-TiO2/Fe3O4 against three target microorganisms including Aeromonas hydrophila, Edwardsiella tarda and Photobacterium damselae subsp. piscicida were due to the solubility of the Ag+ ions in the microorganism-containing medium. The reactive oxidative species has then damaged the cell membrane when the cell came into contact with the catalyst surface. The studies by Jesudoss et al. (2016) showed that there was a significant zone of inhibition reduction on both gram-positive bacteria (S. aureus and B. subtilis) and gram-negative bacteria (P. aeruginosa and E. coli) at Mn1–xNixFe2O4 concentration of 200 μg/mL (Fig. 8.3a). Under the similar experimental condition, these zones of inhibition have almost closer activity values as compared to that of standard streptomycin (Fig. 8.3b). They deduced that the possible mechanism for the bacterial inactivation activity of the spinel magnetic nanoparticles can be elucidated in two probable ways as indicated in Fig. 8.3c, d. One possible mechanism was that the different oxidation cations including Mn2+, Ni2+ and Fe3+ ions from the nanoparticles will enter the bacterial cell and partook in the interaction with the

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bacteria negative part and caused the bacterial cell death. The other mechanism was that the kind of electrostatic interaction generated the reactive oxidative species production; in particular, H2O2 being a strong oxidizing agent readily destructed the bacteria. In summary, researchers have reported their distinctive mechanisms for the antimicrobial activity of magnetic-based photocatalyst particles in the literature. Some researchers also analysed morphological bacterial changes were induced by magnetic-based photocatalysts via transmission electron microscopy or field emission scanning electron microscopy analysis to verify the multiple mechanisms. Though the magnetic-based photocatalysts have demonstrated their superior performances, the exact mechanism is still requiring deep explanations on the spectrum of antibacterial activity. The antibacterial mechanisms that should be put forward in the literature are as follows: direct contact of photocatalysts with cell walls, leading to the leakage of bacterial cell integrities such as protein, DNA and potassium ion (Wang et al. 2016; Lam et al. 2018), addition characterization techniques of death cell (Jing et al. 2016) and quantitative detection of metal ions during the disinfection.

8.5

Using Magnetic-Based Catalyst in Photocatalytic Abatement of Organics

Over the past century, different applications for advanced science and technology relating to industrial and agricultural activities have introduced large quantities of dyes, phenols, solvents, pesticides and other organic pollutants in the environment. Most of these organic pollutants have been identified in surface waters, and thus their removal was a subject of broad interest in the past decades. The photocatalytic reaction has become a promising degradation method to convert the organic pollutants into innocuous products such as CO2 and H2O to eliminate the environmental pollution. Semiconductors with magnetic properties have been reported as significant photocatalytic materials in various organic pollutants’ degradation as they facilitated the recycling process because they can be separated easily by external magnetic field after photocatalytic reactions (Pang et al. 2016; Mousavi et al. 2018). Bishnoi et al. (2018) carried out the photocatalytic degradation of aqueous methylene blue using the inedible Cynometra ramiflora fruit extract-mediated green synthesis of magnetic Fe2O3 nanoparticles. Their results showed that the degradation of methylene blue was accelerated with respect to irradiation time in the presence of the synthesized Fe2O3 nanoparticles. The methylene blue showed a characteristic peak at 663 nm, and it became weaker as the irradiation time increased and eventually disappeared after 110 min (Fig. 8.4a). They also added that the vial that contained methylene blue and Fe2O3 nanoparticles completely lost its colour when exposed to sunlight irradiation, revealing the excellent photocatalytic activity of their synthesized nanoparticles. Their synthesized Fe2O3 nanoparticles also can be easily magnetically separated without any significant loss of the photocatalytic activity

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Fig. 8.4 (a) UV-vis spectra of methylene blue degradation over magnetic Fe2O3 nanoparticles under sunlight irradiation. (b) Reusability of magnetic Fe2O3 nanoparticles for methylene blue degradation. (Adapted from Bishnoi et al. (2018))

even after five consecutive cycles of photodegradation (Fig. 8.4b). This demonstrated great potential for practical applications in environmental cleanup and solar energy conversion. The photocatalytic degradation of tetracycline in simulated wastewater was investigated in the presence of FeNi3/SiO2/CuS magnetic nanocomposites (Nasseh et al. 2018). The findings in their report showed that various parameters (solution pH, catalyst dosage and initial substrate concentration) exerted their individual influence on the degradation efficiency of tetracycline. Under the determined optimal conditions (pH ¼ 9, catalyst dosage ¼ 0.005 g/L, tetracycline concentration ¼ 10 mg/L), the degradation efficiency of their synthesized nanocomposites reached almost 100%, and the mineralization efficiency was reported to be 64.96%. Furthermore, the tetracycline degradation showed a decrease of only 15% after recycled for five cycles. They went further to conclude that the FeNi3/SiO2/CuS photocatalysis had an appropriate efficiency for degrading tetracycline from aqueous environments and it was economical regarding its recyclability. Tong et al. (2016) investigated the visible light photocatalytic reaction to degrade methyl orange using hydrothermally synthesized magnetic BiFeO3 disks. It was suggested that the photodegradation of methyl orange was initiated by •OH radicals as evidenced by the terephthalic acid fluorescence probing technique. Under their experimental conditions, the photocatalytic degradation efficiency of methyl orange reached 97% after 180 min irradiation. Methyl orange and rhodamine B dyes were also subjected to visible light in the presence of magnetic-based BiFeO3/CuWO4 heterojunctions, and the degradation of dyes was analysed using UV-vis spectrophotometer (Ramezanalizadeh and Manteghi 2017). It was reported that the methyl orange and rhodamine B dyes were photocatalytically degraded about 85 and 87%, respectively. A detailed photocatalytic enhancement mechanism over BiFeO3/ CuWO4 heterojunctions has been suggested (Fig. 8.5). They went further to attribute that this improvement was associated with high separation and easy transfer of photogenerated charge carriers at the heterojunction interface. Furthermore, the BiFeO3/CuWO4 heterojunctions were also regarded as prominent and durable

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Fig. 8.5 Proposed photocatalytic enhancement mechanism of methyl orange and rhodamine B degradation over BiFeO3/CuWO4 heterojunction. (Adapted from Ramezanalizadeh and Manteghi (2017))

photocatalysts, which are capable of repeated use without substantial loss of catalytic ability. Using TiO2-coated magnetic activated carbon as catalysts, Ao et al. (2008) examined the photocatalytic degradation of phenol aqueous solution under UV-A irradiation. In their study, the magnetic photocatalysts were developed by depositing the anatase TiO2 onto the surface of magnetic activated carbon which was prepared via the adsorption of magnetic Fe2O3 nanoparticles onto the activated carbon. Vibrating sample magnetometer analysis indicated that the developed photocatalysts were superparamagnetic and can be separated by an external magnetic force due to the low remnant magnetization value. Under the experimental conditions, their developed photocatalysts showed excellent degradation activities as compared to commercially available TiO2-P25. A very small amount of intermediates such as hydroquinone, catechol, benzoquinone and resorcinol were reported during the phenol degradation, which was due to the strong adsorption activity of activated carbon to the intermediates. They also added that the developed photocatalysts can be reused without any mass waste and the degradation efficiency of phenol was still higher than 80% after six repetitive uses.

8.6

Photocatalysis for the Simultaneous Treatment of Bacteria and Organics

An urgent demand for the development of a better method for offering clean water has stimulated tremendous research interest in the photocatalytic simultaneous treatment of bacteria and organic pollutants (Adán et al. 2017; Zhang et al. 2018; Eswar et al. 2018). Zhang et al. (2018) tested the photocatalytic removal of organic

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matters and bacteria simultaneously from real wastewater treatment plant over Er-Al co-doped ZnO photoanode. Under visible light irradiation, the chemical oxygen demand and dissolved organic carbon removal efficiencies were reported to be 69.7% and 53.0%, respectively. A nearly total disinfection of bacteria cell and Escherichia coli cell was also observed in their photocatalytic testing. They also added that the removal of organic matters and bacteria was effectively converted to electricity as proven by their high energy conversion indicator of 36.5%. Bacterial inactivation and organic pollutants degradation have also been investigated using the TiO2 supported on porous stainless steel photocatalytic membranes (Adán et al. 2017). Their results showed that the photocatalytic activity of the membrane had the potential to simultaneously disinfect the water and remove the organic pollutants. They also found out that the photocatalytic functionality had biofouling control through a substantial inactivation of bacterial cells. Although there have been reports on the photocatalytic simultaneous treatment of bacteria and organic pollutants, the research towards the application of magnetic photocatalysts in simultaneous bacterial inactivation and organic pollutants degradation is rarely reported, and a number of challenges remained to be explored.

8.7

Conclusion and Future Prospects

This present article focused to elucidate and investigate research works that surmounted the possible utilization of magnetic-based photocatalysts for the photocatalytic disinfection and degradation activities. The utilization of magneticbased photocatalysts for water decontamination has a significant potential due to their excellent catalytic potentials and their facile recoveries with a magnetic separation stage, enabling the continuous use of photocatalyst. As can be gleaned from the reports, there were a number of influencing factors with regard to the effects of photocatalyst concentration and light intensity, nature of microorganism, solution pH, initial bacterial concentration and physiological state of bacteria on the magnetic-based photocatalyst inactivation. Nevertheless, it may be difficult to draw conclusions on which could give the best findings as the interlaboratory variations in methodologies utilized and the bacterial species and concentration as well as other factors. Extensive discussion was provided to action mechanisms which served as the key subject in the photocatalytic disinfection of bacteria. The induction of surface generation reactive oxidative species and release of metal ions which adhered on the cell membrane can lead to cell death and have been deliberated as a main activity of the magnetic-based photocatalysts. The novel magnetic-based photocatalysts based on different nanocomposite materials have shown potential photocatalytic activities in the presence of visible light irradiation. Special attention was paid to the magnetic recovery, and reuse of magnetic-based photocatalysts as a good material recovery should be ensured with higher removal efficiency and rapid removal rates in the presence of visible light irradiation. Additionally, a brief

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discussion was delivered to one important application as a simultaneous treatment of bacteria and organic pollutants on the water decontamination. Despite the substantial progress of the magnetic-based photocatalyst has been achieved, there were some hurdles still requiring to be further examined prior to their complete establishment. The first challenge was lack of comprehensive understanding of the visible light photocatalytic disinfection mechanism by magnetic-based photocatalyst materials. Fundamentally understanding the mechanism was important as the visible light activity and stability of magnetic-based photocatalysts relied on further investigation of the mechanism and the interactions of the catalyst active-site structures with microorganisms. The utilization of new and novel material characterizations and computational techniques was necessitated for further research works. The practices used were not only envisioned at the design and tailor-made photocatalyst model development at molecular and electronic levels but can also be adequately scrutinized with the virtual importance of some synthetic processes as well as under multi-interaction conditions. A direct and systematic comparison of magnetic-based photocatalyst materials with other visible light photocatalyst materials was also essential to provide the potential application for the next-generation solar photocatalyst systems. A demonstrated ability to employ the magnetic-based photocatalysts at a solar pilot scale for water decontamination processes would certainly benefit the commercial sector both in terms of environment and economy. Furthermore, most of the literature studies focused only on the degradation rate and efficiency of target organic dyes disregarding the mineralization as well as identification of the degradation intermediates. This aspect should not be overlooked while reporting any future work. Undeniably, there were still some challenges, and prospects should be focused on the magnetic-based photocatalysts’ photocatalytic system. However, the outlook of progressing intense and diverse research on magneticbased photocatalyst has shown the evidence of its potential as an effective photocatalyst on the water treatment processes. Acknowledgements This work was supported by the Universiti Tunku Abdul Rahman (UTARRF/2018–C2/S02 and UTARRF/2018–C1/L02) and Ministry of Higher Education of Malaysia (FRGS/1/2016/TK02/UTAR/02/1).

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Chapter 9

Antimicrobial Activities of Photocatalysts for Water Disinfection Veronice Slusarski-Santana, Leila Denise Fiorentin-Ferrari, and Mônica Lady Fiorese

Contents 9.1 9.2 9.3 9.4 9.5 9.6 9.7

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Mechanisms of Photocatalytic Disinfection . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Pure and Modified Photocatalysts . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Photocatalytic Films and Biofilms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Photocatalytic Composites and Nanocomposites . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Materials with Antimicrobial Activity in the Absence of Light . . . . . . . . . . . . . . . . . . . . . . . . . . Case Study: Application of Supported Photocatalysts in Disinfection of WheyProcessing Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.8 Final Considerations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract Water is an essential element for the survival of humanity and the continuity of the planet’s development. Impacts of human activity on water resources have been a growing concern because of environmental contamination and public health problems. For that reason, strategies and technologies to promote water sustainability and ensure water availability for current and future generations, with suitable quality for its respective uses, are being researched and developed, with priority being given to modern, efficient, inexpensive, and fast techniques that can achieve adequate water disinfection. Water disinfection is an efficient process for removal, deactivation, or killing of microorganisms that are responsible for waterborne diseases, and it is applied in public water supply and reuse systems. Among disinfection techniques, a promising and efficient method used against a wide range of different species of organisms is photocatalytic water disinfection using a photocatalyst activated by ultraviolet (UV) or visible radiation for generation of reactive oxygen species (ROS) that damage microorganisms. We performed a review of the literature on the mechanisms of photocatalytic water disinfection and photocatalysts with microbial activity. It was found that disinfection may occur due to (1) attack by ROS on bacterial cells, (2) the effects of metal ion release on cellular V. Slusarski-Santana (*) · L. D. Fiorentin-Ferrari · M. L. Fiorese Department of Chemical Engineering, West Paraná State University, Toledo, Paraná, Brazil e-mail: [email protected] © Springer Nature Switzerland AG 2020 Inamuddin et al. (eds.), Nanophotocatalysis and Environmental Applications, Environmental Chemistry for a Sustainable World 30, https://doi.org/10.1007/978-3-030-12619-3_9

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proteins, (3) direct interactions between nanoparticles and bacterial cells, and (4) mechanical damage to cell membranes. Generation of ROS in the presence of light occurs with activation of the photocatalyst with appropriate radiation; however, the mechanism of disinfection when it is performed in the absence of light has not yet been completely defined. In order for industrial application of photocatalytic water disinfection to become feasible, it is necessary to develop photocatalysts that exhibit high antimicrobial activity and can be activated preferentially with solar radiation. Among the different types of photocatalysts are pure photocatalysts and nonmetals, halogens, metals, and rare earth–modified photocatalysts, as well as photocatalytic films, biofilms, and nanocomposites. Modified photocatalysts, especially fluorineand cobalt-modified zinc oxide and silver- and nitrogen-modified titanium dioxide, are more efficient than pure oxides under visible radiation, which makes their use promising in solar-induced photocatalytic water disinfection because modified photocatalysts can be reused and are not harmful to human health and the environment, being in accordance with the principle of green chemistry. Photocatalytic films, biofilms, and nanocomposites include silver- and copper-modified potassium hexaniobate film; silver orthophosphate, titanium dioxide, and magnetite film; and titanium dioxide and tungsten(VI) oxide anchored on a reduced graphene oxide nanocomposite, which demonstrate high antimicrobial activity with a mortality rate higher than 97%. As few studies have been performed with native microorganisms and under real conditions, a case study was performed experimentally by the authors, involving application of a supported photocatalyst for disinfection of wastewater from whey processing during exposure to UV radiation. The photocatalyst, constituted by hematite and titanium dioxide supported on a glass sphere, showed high bacteriostatic activity against mesophilic microorganisms and irreversible damage to psychrophilic microorganisms’ cell walls and their components. Therefore, this chapter contributes to the knowledge about use of photocatalysts as water disinfection agents and for control of waterborne diseases, especially those that are activated by solar radiation, thus making photocatalysis an efficient, viable, and environmentally safe alternative. Keywords Photocatalysis · Nanoparticles · Biofilm · Nanocomposite · Supported photocatalysts · Microorganism · Microbicidal · Biostatic · Cell damage · Reactive oxygen species

9.1

Introduction

Water is a natural resource, is vital for the survival of humanity and all living organisms, and maintains the balance of ecosystems (Chaplin 2001; Popkin et al. 2010). Consequently, it is vitally important that this natural resource is suitable for physical, chemical, and microbiological conditions; in other words, water should be exempt from substances that can cause harmful effects on the environment and living organisms (Clasen et al. 2006; Vyas et al. 2015).

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Most organisms commonly found in water are, in origin, pathogenic or undesirable, and they are capable of causing damage to the health of consumers, as well as losses of production in food systems or alterations in the microbial flora of ecosystems. The most important bacteria transmitted through contaminated water are coliforms (total coliforms), fecal or thermotolerant coliforms, Escherichia coli, enterococci (fecal streptococci or intestinal enterococci), bacteriophages, Salmonella, Shigella, Campylobacter jejuni, Yersinia enterocolitica, Vibrio, Candida albicans, Pseudomonas aeruginosa, and Staphylococcus aureus. They are responsible for diseases such as typhoid fever, paratyphoid, cholera, bacillary dysentery, hepatitis, and gastroenteritis (Cabral 2010; Saxena et al. 2015; An et al. 2017). Treatment of water either for consumption or for industrial reuse is becoming increasingly important, since the impacts of human activity on water resources are causing widespread worry for the future of humanity (Cosgrove and Loucks 2015). Therefore, the search for efficient treatments to ensure good water quality—including modern, efficient, inexpensive, and rapid techniques to disinfect water for the intended purposes—has become a subject of interest for numerous researchers over the years (Gevod et al. 2005; Goncharuk et al. 2018). The process of disinfection means removal, deactivation, or killing of undesirable organisms or pathogenic microorganisms responsible for waterborne diseases. The quality of water-disinfecting agents is described in the literature as being dependent on a variety of factors (National Research Council 1980; Fraise et al. 2004; An et al. 2017): • High capacity to rapidly destroy pathogenic organisms in the conditions and concentrations that are present in the water • Nontoxicity to the environment and to humans • High microbicidal potential • Relatively low cost • Facility and safety of transportation, storage, handling, and application Photocatalytic disinfection of water meets certain exacting criteria and definitions, and is considered a promising and efficient method for use against a wide range of different species, including algae, viruses, fungi, and Gram-positive and Gram-negative bacteria (Verdier et al. 2014; An et al. 2017). Photocatalysis for microbial inactivation in water occurs by the action of a light source (usually ultraviolet (UV) radiation) and a photocatalyst, which together act as agents for destruction or inhibition of pathogens. The antimicrobial agents used in disinfection processes are classified according to the effects they exert on microorganisms; when they act against bacteria and fungi, they are considered antibacterial and antifungal agents, respectively. They act in two forms: as microbicidal/biocidal agents when they attack and damage cells, and as biostatic agents when they inhibit growth (Shateri-Khalilabad and Yazdanshenas 2013; Verdier et al. 2014; Ameta and Ameta 2017).

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Mechanisms of Photocatalytic Disinfection

The photocatalytic disinfection process submits living organisms to physical and chemical events. According to the literature, the most common sequence of events (Foster et al. 2011; Thabet et al. 2014; Gilbertson et al. 2016; Teodoro et al. 2017) is: 1. Damage to the cell wall: After direct contact with the photocatalyst particle (a physical event), disruption and damage to cell wall structures occur. How to the peptidoglycan is a highly cross-linked molecule, damage to the cell wall occurs specific sites and it can not be viewed easily. 2. Cell membrane damage: Intracellular oxidative stress is caused by elevated levels of reactive oxygen species (ROS), which react with bacterial structural components (such as the cell wall and cell membrane), especially unsaturated phosphate lipids in the cell membrane. This first chemical event that occurs during the photocatalytic antibacterial process is initiated by an autocatalytic lipid peroxidation reaction with hydrogen abstraction or addition of an oxygen radical, resulting in oxidative damage to polyunsaturated fatty acids with formation of a lipid peroxyl radical, which leads to the production of toxic organic hydroperoxides. Consequently, changes in the physical and chemical properties of the membrane imply loss of selectivity in ion exchange, changes in membrane permeability, and formation of cytotoxic products. 3. Direct attack on intracellular content after exposure to ROS: This causes degradation of proteins, i.e., expansion of intracellular fluid, implying release of the content of organelles and culminating in cell death. In some cases, depending on the type of photocatalytic treatment applied, the damage caused to the nucleic acids leads to complete disappearance of microbial DNA and RNA, resulting in mineralization of the microorganism. Inactivation of bacterial contaminants by the photocatalytic process is better described in the literature than inactivation of other species of microorganisms, and the effectiveness of this process has already been proved by several techniques such as atomic force microscopy (AFM), transmission electron microscopy (TEM), scanning electron microscopy (SEM), x-ray diffraction (XRD), and Fouriertransform infrared spectroscopy (FTIR). The photocatalytic disinfection process can be divided into four basic mechanisms (Padmavathy and Vijayaraghavan 2008; Zhang et al. 2008; Foster et al. 2011; Espitia et al. 2012; Azizi et al. 2014; Carré et al. 2014; Verdier et al. 2014; Prasanna and Vijayaraghavan 2015; Sirelkhatim et al. 2015; Bui et al. 2017): 1. ROS attack on the bacterial cell: ROS (such as O2•
, HO
, •OH, HO2•, H2O2, and 1O2) are formed mainly by activation of a photocatalyst with UV and/or visible radiation, and they damage bacterial cells from both the inside and the outside. Penetration of the cell membrane by the hydrogen peroxide (H2O2) molecule causes damage and destruction of the cell cytoplasm, accelerating the death process. The outside cell wall is attacked by the O2•
and HO
species, which cannot enter the bacterial membrane, because of their negative charge, but

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they may remain in direct contact with the bacterial outer surface, promoting breakage of the cell wall. The high reactivity and oxidizing property of ROS are responsible for their toxicity to microorganisms. 2. Release of metal ions from a photocatalyst into the solution: As they disperse into the solution, metal ions present in the photocatalyst interact with protein, causing allosteric changes that alter its function and interrupt vital processes for the cell (such as regulation of cell growth, energy production, intercellular signaling, immunity and defense, synthesis of biological substances, and phagocytosis) and interact with nucleic acids by stopping cell replication. This release of metal ions depends on the physicochemical properties of the particles (such as their porosity, surface chemistry, concentration, size, and morphology) and the medium conditions (such as its pH, the type of radiation, the exposure time, and the presence of other ionic species). It is still necessary to consider the intrinsic characteristics of each microorganism, since—depending on the type of microorganism and the released concentration of metal ions during the inactivation—these ions can act as nutrients for microorganisms instead of inactivating them. 3. Direct interaction between nanoparticles of the antimicrobial agent and the bacterial cell: The existence of electrostatic forces originating from the opposite charges of the negative carboxyl groups in the microbial cells and the positive surface of the photocatalyst promotes direct interaction between them. Accumulation of nanoparticles in the cell membrane may cause its disruption and raise its permeability, allowing the nanoparticles to reach the cytoplasm and cause reversible initial damage; may cause phagocytosis of nanoparticles and cell wall disorganization, followed by internalization of the nanoparticles into the cells; and may cause alteration of the cellular morphology and intracellular content leakage. All of these types of damage cause oxidative stress, promoting inhibition of cell growth and microbial death, and thus irreversible damage. The internalization is controlled by the size, surface chemistry, defects, and functionalization of the nanoparticles. 4. Mechanical damage to the cell membrane: The photocatalyst may have an abrasive texture due to surface defects such as rough edges and corners, so the impact of the photocatalyst particles on microorganism cell walls can lead to mechanical injury/deformation of the walls. The four proposed mechanisms all promote destabilization in the microbial cell; however, a greater number of studies have concluded that the efficiency of photocatalytic disinfection is mainly due to oxidative stress caused by ROS, because uncontrolled accumulation of hydroxyl radicals causes breaks in nucleic acids, carbohydrates, and proteins, and peroxidation of lipids, in both the presence and the absence of light. Activation of a photocatalyst with appropriate light, usually UV or light with energy greater than its bandgap energy, causes excitation of electrons from the valence band to the conduction band of the semiconductor, generating electron/ hole (e
/h+) pairs. The photogenerated electrons (e
) reduce oxygen to form a superoxide radical anion (O2•
), which is oxidized by the photogenerated hole

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(h+) to produce singlet oxygen (1O2), while the h+ reacts with adsorbed water (H2O) or hydroxyl ions (HO
) to form ROS. Initially, the hydroxyl radicals (•OH) are formed and, by the union of two of these radicals, formation of hydrogen peroxide (H2O2) occurs. Then, the hydrogen peroxide reacts with •OH and forms HO2•, which acts as O2•
in the antimicrobial inactivation (Etacheri et al. 2013; Verdier et al. 2014; Carré et al. 2014; Yadav et al. 2016). However, when formation of the e
/h+ pair is caused by visible radiation, the photogenerated holes (h+) in these conditions do not have enough reduction potential to generate hydroxyl radicals by oxidation of water molecules. Therefore, less oxidative species such as 1O2 are responsible for the antibacterial effect during visible light–induced photocatalysis (Podporska-Carroll et al. 2017). With regard to production of ROS in the absence of light, Prasanna and Vijayaraghavan (2015) state that ROS (especially •O2
) are produced significantly by a singly ionized oxygen vacancy from the aqueous suspension of the photocatalyst even in the dark, and surface defects play an important role in the production of ROS. According to Gilbertson et al. (2016), generation of ROS in the absence of light can be caused by the photocatalyst itself or by release of ions into the solution, and it can be attributed to (1) redox reactions that occur between photocatalyst particles and microbial cells, (2) reactions between photocatalyst particles and the culture medium, and (3) species released by photocatalyst particles. In the latter two cases, there is no need for cell–particle contact for ROS production to occur. However, it is agreed that nanoparticles demonstrate lower antimicrobial activity in the dark than in the presence of light. Other researchers (Banoee et al. 2010; Khan et al. 2014; Prasanna and Vijayaraghavan 2015; Shao et al. 2015) have shown that the mortality of microorganisms, when they are exposed to antimicrobial agents in the dark, is mainly caused by leaching of ions from the suspension, which are internalized and damage the microbial cell. Pasquet et al. (2014) attributed reduction of the cell viability of E. coli, P. aeruginosa, and S. aureus to release of dissolved zinc ions into the medium by partial dissolution of zinc oxide (ZnO) particles, and stated that this process was dependent on the contact time and the concentration and surface area of the ZnO. Rtimi et al. (2015) demonstrated the mechanism of inactivation of E. coli using a composite formed by polyester (PES) and titanium dioxide (TiO2) as a bactericidal agent under dark conditions and a TEM technique. Initially, there was interaction between the PES–TiO2 composite and the cell wall of the intact bacterium by means of electrostatic forces that overcame the van der Waals forces, forming an agglomeration of TiO2 nanoparticles on the surface of the cell. After 30 min, the cell wall and membrane were attacked by TiO2 nanoparticles, and at 120 min, cell inactivation occurred as a result of the damage caused to the external layers of the bacterium because Ti4+/Ti3+ redox catalysis occurred in this material. Antimicrobial activity depends on the characteristics of the photocatalyst, such as its concentration, morphology (size, surface area, and shape), stability, crystalline structure, surface coverage, state of aggregation, and electronic properties, as well as the medium conditions, such as the time, type, and intensity of the radiation; the pH;

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and the microorganism species and concentration (Espitia et al. 2012; Aich et al. 2014; Verdier et al. 2014). The characteristics that favor antimicrobial activity of the photocatalyst are high dispersion of the particles, which promotes a high probability of contact with microorganisms; small particle size, resulting in a high surface area–to–volume ratio and high surface reactivity; and the concentration of nanoparticles. These characteristics are related to generation of ROS (particularly hydrogen peroxide), progressive release of lipopolysaccharides and membrane proteins, and damage to nucleic acids associated with proteins and intracellular factors. The attack of hydrogen peroxide occurs in the nitrogenous bases and the sugar–phosphate backbone, provoking single- and double-stranded DNA breaks. Therefore, nanoparticles (smaller than 100 nm) of the photocatalyst are more efficient than microparticles (100–1000 nm in size) because nanoparticles have the ability to enter the cell membrane, since the bacterial cell size is in the range of micrometers, while the outer membranes have pores in the nanometer range, allowing only nanoparticles to enter (Kohane 2007; Azam et al. 2012; Espitia et al. 2012; Reddy et al. 2014; Prasanna and Vijayaraghavan 2015; Sirelkhatim et al. 2015; Verdier et al. 2014). In addition, smaller particles exhibit high activity, and the particle size influences the path to be traversed by active centers for oxidation–reduction reactions (i.e., the smaller the particle, the better), thereby reducing the likelihood of charge recombination (Lu et al. 2008; Azizi et al. 2013; Bui et al. 2017). The influence of copper (II) oxide (CuO) morphology and particle size on antimicrobial activity was studied by Gilbertson et al. (2016), who found that CuO nanosheets exhibited higher surface reactivity and higher electrochemical and antimicrobial activity than bulk microsized CuO and CuO nanoparticles (of a sphere-like shape). Azam et al. (2012) observed that antimicrobial activity varied inversely with the size of ZnO, CuO, and Fe2O3 particles. With regard to the number of nanoparticles, according to Cordeiro et al. (2004) and Udayabhanu et al. (2018), an increase in the number of nanoparticles favors inactivation of microorganisms. The effectiveness of the photocatalytic process is influenced by the time and light intensity, being higher when the microorganism is exposed to high radiation for long periods of time (Cordeiro et al. 2004; Verdier et al. 2014). In relation to the type of light, the components of the solar spectrum that may affect the viability of microorganisms are UV-B (280–320 nm), UV-A (320–400 nm), and visible (400–700 nm) light. UV-B light directly affects the genetic material of the cell, while UV-A and visible light promote activation of the photocatalyst (Lanao et al. 2012). Carré et al. (2014) studied the effect of UV-A photocatalysis with TiO2 on degradation of E. coli and found that the longest exposure to UV radiation promoted the lowest viability of E. coli. The presence of the O2•
radical contributes to photocatalytic antimicrobial activity, promoting simultaneous effects of this radical on lipid peroxidation and the proteome during the process. The highest lipid peroxidation was obtained with TiO2 0.4 g L
1 and 60 min of irradiation, evidencing that the contact between the microorganism and TiO2 particles plays an important role in photocatalytic disinfection. In the process of photolysis in the absence of a catalyst, UV radiation also

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causes excitation of photosensitive molecules inside the cell, causing damage to microorganisms (Yadav et al. 2016). The antimicrobial activity of the photocatalyst varies greatly, depending on the microbial species. The difference in the activity is related to the structure and composition of the microorganism cell wall (which may promote greater or lesser resistance to biocide attack), the sensitivity of the microorganism to stress caused by ROS, the affinity between the cell membrane and the bactericide agents, and the cell membrane polarity. Gram-negative bacteria are generally less sensitive to ROS than Gram-positive bacteria because of the polarity of the lipid bilayer of the cell membrane; Gram-positive bacteria have a lower negative charge, which facilitates attack by negatively charged free radicals, causing damage and cell death with a lower concentration of the photocatalyst than that required for Gram-negative bacteria (Espitia et al. 2012). Azizi et al. (2013) demonstrated that ZnO nanoparticles and cellulose nanocrystal (CNC) nanocomposites had higher antibacterial activity against S. aureus (which is Gram positive), because of its simpler structure, than against Salmonella enterica serotype Choleraesuis (which is Gram negative), as the latter has an external lipopolysaccharide layer that protects the peptidoglycan layer and aids survival of bacteria in inhospitable environments. The effect of the medium pH on the antimicrobial activity of photocatalysts has been little studied. Microbial growth is dependent on this parameter, and the medium pH should be favorable for the target microorganism. Sirelkhatim et al. (2015) reported that a neutral pH (7–7.5) did not influence antibacterial activity in the absence of light. Salehi et al. (2014) observed growth of S. aureus in a pH range from 4.5 to 9.3, and they set an optimal pH of 7–7.5, consistent with the physiology of the bacterium. However, in the presence of cadmium oxide (CdO) nanoparticles, the pH range for S. aureus growth was reduced (to a range of 6–8), though the optimum pH was maintained. Antimicrobial agents may be pure, doped, or modified photocatalysts, commonly employed in the form of nanoparticles dispersed in an aqueous solution or in the form of coatings/films applied to various substrates or composites (Azizi et al. 2013; Prasanna and Vijayaraghavan 2015; Verdier et al. 2014; Kim et al. 2018; Doran et al. 2018).

9.3

Pure and Modified Photocatalysts

Currently, the antimicrobial agents most often used in the photocatalytic disinfection process are silver (Ag), zinc oxide (ZnO), and titanium dioxide (TiO2). These materials can either be used in their pure form or modified with cobalt, cerium, fluorine, carbon, iron, copper, strontium, vanadium, nitrogen, sulfur, platinum, silver, gold, or rare earth, which, in some cases, also allows use of visible (solar) radiation in the activation of the photocatalyst. This fact contributes to the green chemistry principle and makes visible light–induced photocatalysis a promising technique for disinfecting water and effluent. The high antimicrobial activity of

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doped photocatalysts in comparison with the pure oxides can be attributed to increases in the generation of ROS (•OH, •O2
, 1O2), the binding forces, and the surface area, and to reductions in the particle size and bandgap energy (Oves et al. 2015; Hui et al. 2016; Yadav et al. 2016). Prasanna and Vijayaraghavan (2015) evaluated the bactericidal activity of ZnO microparticles, ZnO nanoparticles, and oxalic acid–capped ZnO against S. aureus (3  1012 colony-forming units (CFU) mL
1) both in the dark and in the presence of light. In the absence and in the presence of light, ZnO nanoparticles were more efficient (17% and 80%, respectively) than oxalic acid–capped ZnO (9% and 48%, respectively) and ZnO microparticles (0% and 3.8%, respectively). This showed that Zn2+ ion leaching had a minor role in the antimicrobial activity, since the ZnO nanoparticles were negatively charged, as well as the bacteria, which would make it difficult for the ions to be internalized in the bacterial cell. However, the visualization of ZnO inside the bacteria was due to incorporation of the particle after rupture of the cell membrane by ROS. Although the highest concentrations of the •OH radical (8.6 mg L
1) and hydrogen peroxide (14 mg L
1) were found for ZnO nanoparticles in the presence of light, these species were also found in the dark (0.95 and 1.25 mg L
1, respectively), evidencing the effect of ROS on antimicrobial activity even in the absence of light. Pure and cobalt-doped zinc oxide showed antibacterial activity against Shigella dysenteriae, Salmonella typhi, Vibrio cholerae, and E. coli (1  106 CFU mL
1) both in the dark and under solar radiation for 2 h, as reported by Oves et al. (2015). Antibacterial activity increased with cobalt (Co) content, and the largest inhibition zones were obtained with a 5% Co-doped ZnO catalyst against V. cholerae (21 and 23 mm in the absence and in the presence of solar light, respectively) and E. coli (22 and 24 mm, respectively). This catalyst had a smaller particle size (20.5 nm) than ZnO (25.7 nm), which contributed to its greater bactericidal activity. Hui et al. (2016) also found that pure ZnO showed lower antimicrobial activity than ceriumdoped, flower-shaped ZnO against C. albicans and Aspergillus flavus under visible light. An increase in the cerium content favored the antimicrobial activity, a 75% reduction in the population of C. albicans and an 80% reduction in the population of A. flavus with 0.8% and 1.0% Ce-doped ZnO, respectively. The higher efficiency of the photocatalyst against A. flavus could be attributed to differences in the structure and chemical composition of the cell surface between the two microorganisms. The effect of fluorine-doped nanocrystalline ZnO (with molar ratios of ZnO to trifluoroacetic acid (TFA) of 1:1 and 1:2) on the cellular viability of E. coli and S. aureus (1  105 CFU mL
1) was studied by Podporska-Carroll et al. (2017) under visible light for 6 h and in the absence of light. The authors verified that the efficiency depended on the microorganism, light conditions, and type of material. Pure ZnO showed higher antimicrobial activity, achieving log reductions in E. coli and S. aureus of 1.74 and 1.26, respectively, in dark conditions. In the presence of visible light, the efficiency was higher for ZnO:TFA 1:1, with log reductions of 2.88 for E. coli and 4.62 for S. aureus. The stronger bactericidal activity of ZnO:TFA 1:1 could be explained by its chemical composition (specifically, its higher content of oxygen (surface defects), which promoted higher generation of ROS upon

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irradiation), the toxic effect of the released Zn2+ ions, and addition of fluorine within the ZnO framework. The efficiency of titanium dioxide (TiO2) as an antimicrobial agent against Grampositive and Gram-negative bacteria (1  108 CFU mL
1) and fungi (1  107 CFU mL
1) in the absence and in the presence of UV-A (320–400 nm) radiation increased with addition of potassium iodide (KI), according to Huang et al. (2016). The higher the concentrations of TiO2 and KI, the greater the mortality of S. aureus and E. coli bacteria and C. albicans fungus (6 logs of killing with UV-A radiation). With the addition of KI, there was production of reactive short-lived intermediates (iodine radicals (I•) or iodonium cations (I+)) and oxidized products of the iodide anion (molecular iodine, tri-iodide anion, or hypoiodide) in the medium, which favored the process. Reactive short-lived intermediates were mainly responsible for the extra microbial mortality observed during the irradiation of the medium, while the oxidized long-lived products of the anion iodide promoted mortality also after exposure to light (the period in the dark). Yuan et al. (2010) showed that incorporation of nitrogen and silver into TiO2 nanoparticles improved their antimicrobial activity against E. coli and Bacillus subtilis in the presence of fluorescent light (30 W) for 24 h. The co-doped photocatalyst (1% Ag–N–TiO2) was more effective against E. coli and B. subtilis (achieving clear antibacterial circles of 33 and 22.8 mm, respectively) than pure TiO2 (which achieved smaller clear bacteriostatic circles of 16 and 15.2 mm, respectively, for these two organisms). This behavior was attributed to the presence of the Ag+ nanoparticles—which acted as centers of charge separation, inhibiting the process of e
/h+ pair recombination—and an increase in surface hydroxyl groups. In a study by Etacheri et al. (2013), a carbon-doped anatase–brookite heterojunction photocatalyst showed antibacterial activity against S. aureus (1  105 CFU mL
1) twice as high as that of TiO2 P25 with visible radiation over a period of 5 h. As these two materials were unable to inactivate S. aureus in the dark, the bactericidal activity was attributed to the formation of e
/h+ pairs and consequently to ROS. The microbicidal activity of TiO2 nanoparticles dispersed in solution (in a slurry) and deposited on transparent coatings (sterile glass disks) against proliferation of E. coli in conditions of low light intensity (UV radiation 2.5 W m
2) was evaluated by Verdier et al. (2014). The authors observed low antimicrobial activity of the dispersed nanoparticles in the solution after 6 h of irradiation, which was attributed to low dispersion (and consequently less probability of contact between the TiO2 nanoparticles and bacterial cells) and to low light intensity. However, there was an increase in antimicrobial activity when disks with TiO2 nanoparticles were used, because of a reduction in the distance between the TiO2 nanoparticles and the cells, since this contact led to direct oxidation of the cells by the photogenerated holes, also contributing to the reduction of the e
/h+ recombination process in the photocatalyst. Teodoro et al. (2017) used TiO2 supported on microtubes under UV radiation in the absence and in the presence of hydrogen peroxide (H2O2) (150 mg L
1) to disinfect gray water, monitoring the effects on E. coli (1.8 logs), P. aeruginosa (3.4 logs), total coliforms (5.5 logs), and enterococci (2.7 logs). The inactivation of total coliforms with UV/TiO2 was 2 logs within 120 min, whereas with UV/TiO2/

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H2O2 there was total inactivation of total coliforms, enterococci, and P. aeruginosa within 60 min. In addition to zinc oxide and titanium dioxide, other materials have shown photocatalytic antimicrobial activity, such as tin(IV) oxide (SnO2), nickel(II) oxide (NiO) nanoparticles, and silver-loaded and erbium/ytterbium–doped barium molybdate (Ag–Er3+/Yb3+–doped BaMoO4). Amininezhad et al. (2015) confirmed the bactericidal activity of SnO2 nanoparticles in inactivation of E. coli and S. aureus (1.5  106 CFU mL
1) under UV light, with 100% killing of S. aureus and E. coli within 5 and 3 h, respectively. This material also showed activity in the absence of light, but the inactivation was slower: 8 h for S. aureus and 4 h for E. coli. According to Gondal et al. (2011), the decay rate constant for E. coli (4  107 CFU mL
1) was 0.38 min
1 with use of 30 mM of NiO nanoparticles during 30 min of laser-induced photocatalysis. This value was higher than that observed with use of TiO2 (0.24 min
1) under the same conditions. Ray et al. (2017) investigated the antimicrobial activity of Ag-loaded and Er3+/ Yb3+–doped BaMoO4 against E. coli (3.6  106 CFU mL
1), P. aeruginosa, and S. aureus (8  107 CFU mL
1) under visible radiation. Increases in the silver content and the photocatalyst dose favored bacterial killing, and P. aeruginosa showed the largest inhibition zone, due to its finer peptidoglycan wall. Complete inactivation of E. coli, P. aeruginosa, and S. aureus was achieved within 1, 4, and 5 h of irradiation and with 0.2, 2.0, and 2.0 mg mL
1 of 1% Ag–BaMoO4:Er3+/Yb3+, respectively. The authors attributed the bactericidal activity of this material to creation of defects (surface oxygen vacancies) by doping with rare earth and to the reduction of e
/h+ pair recombination by the presence of silver. The majority of studies on antimicrobial activity have made use of techniques in which the antimicrobial agents were added to the culture medium (liquid or solid), remaining incubated in the absence of light. Two studies on materials that showed antimicrobial activity in the dark and photocatalytic activity under radiation (mainly UV) have been reported. The antibacterial and antifungal activities of pure cerium(IV) oxide (CeO2) and zirconium-doped CeO2 nanoparticles (5, 10, and 15 wt%) against S. aureus, E. coli, P. aeruginosa, Streptococcus faecalis, B. subtilis, and Proteus vulgaris bacteria and against C. albicans, Aspergillus terreus, Candida tropicalis, Aspergillus fumigatus, A. flavus, and Aspergillus niger fungi (all at 1  105 CFU mL
1) were evaluated by Bakkiyaraj et al. (2017). All samples showed moderate antimicrobial activity in dark conditions, but for S. aureus and E. coli, 15% Zr-doped CeO2 was shown to be more efficient (with 11-mm and 7-mm inhibition zones, respectively) than pure CeO2 (with 8-mm and 6-mm inhibition zones, respectively). Five percent Zr-doped CeO2 was more efficient than pure CeO2 for A. terreus (7 mm versus 5 mm) and A. flavus (12 mm versus 11 mm). Cell wall synthesis inhibition and cell capsular degradation by nanoparticles may have led to these differences in activity. TiO2 nanoparticles, biosynthesized using an aqueous extract of plant leaves, showed antibacterial activity against S. aureus, Staphylococcus epidermidis, P. aeruginosa, E. coli, P. vulgaris, and Klebsiella pneumoniae, as reported by Udayabhanu et al. (2018). However, inhibition of cell growth was dependent on

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the microorganism: S. epidermidis showed the largest inhibition zone (16.8 mm), followed by P. vulgaris (13.8 mm), whereas for S. aureus the inhibition zone measured only 9.6 mm. On the basis of the aforementioned studies, it has been established that modified photocatalysts are more efficient than pure photocatalysts—especially fluorine- and cobalt-modified zinc oxide, and silver- and nitrogen-modified titanium dioxide— when used together with solar radiation, which makes their use promising for solarinduced photocatalytic water disinfection.

9.4

Photocatalytic Films and Biofilms

Photocatalytic films and biofilms can be prepared from a combination of different materials. In the case of films, they may consist of a single oxide deposited on inert materials, such as glass and silica fibers, or a combination of pure or modified oxides with silver, copper, tin, and nitrogen. Biofilms can be composed of polymers, chitosan or cellulose, and oxides, with incorporation of metal particles and pure or modified metal oxides. Films and biofilms can be used as antimicrobial agents against various microorganisms in the presence of UV or visible radiation. This type of material is generally used in control of bacterial biofilms, which are microchannels that facilitate the circulation of nutrients, promoting multiplication and dispersion of bacteria (del Valle et al. 2016). Glass fiber thin films coated with nitrogen-doped SnO2/TiO2 showed antimicrobial activity against E. coli, S. typhi, and S. aureus (1  103 CFU mL
1) in the presence of UV radiation, as reported by Kongsong et al. (2014). The best result was obtained with a 20N/3SnO2/TiO2 film, which presented rate constants of 0.450 min
1 for S. typhi, 0.128 min
1 for E. coli, and 0.082 min
1 for S. aureus. These values were higher than those observed with TiO2 film (0.24, 0.082, and 0.058 min
1, respectively). E. coli was most rapidly inactivated (within 20 min), followed by S. typhi (within 40 min) and S. aureus (within 60 min). Wong et al. (2015) prepared silver nanoparticle–impregnated and nitrogen-doped TiO2 sandwich films (TiO2(N)/Ag/TiO2(N)), which presented high antimicrobial activity against S. aureus, E. coli, Streptococcus pyogenes, and Acinetobacter baumannii (1  106 CFU mL
1) under conditions of darkness and visible radiation. The synergistic effect of Ag nanoparticles and nitrogen-doped TiO2 was responsible for the antimicrobial activity in both the dark and visible light conditions. The sample with the highest silver content exhibited the greatest activity, and S. pyogenes was efficiently inactivated with visible radiation. However, the antimicrobial activity decreased with reuse of the material over three cycles. The bactericidal activity of potassium hexaniobate (K4Nb6O17) thin films modified with silver (Ag), copper (Cu), and both (Ag–Cu) against E. coli (1–5  107 CFU mL
1) in the presence of visible light (10 W) for 45 min was investigated by Lin and Lin (2012). The authors achieved greater than 99% elimination of the bacteria with Ag–Cu(3)K4Nb6O17 film but only 33% with K4Nb6O17

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film. The efficiency of the Ag–Cu(3)K4Nb6O17 film was attributed to high dispersion of the Ag–Cu nanocomposites (which avoided the e
/h+ recombination process) and synergism between Ag and Cu ions. However, other factors—including chemical states, loading amounts, and particle size—also affect photocatalytic inactivation. A film nanocomposite containing silver orthophosphate, titanium dioxide, and magnetite (Ag3PO4/TiO2/Fe3O4) showed excellent antibacterial activity and recyclability when used against E. coli (1  107 CFU mL
1) under visible light, according to Xu et al. (2014). With use of 0.30 mg of Ag3PO4/TiO2/Fe3O4 and Ag3PO4 film, 99.8% bactericidal efficiency was achieved within 5 and 15 min of visible radiation, respectively. This activity was maintained for two cycles of use, and with five cycles of use, the Ag3PO4/TiO2/Fe3O4 film still had 55% efficiency, whereas the Ag3PO4 film achieved only 10% inactivation. The bactericidal activity was attributed to photogenerated ROS (•OH and O2•
). According to Xing et al. (2012), polyethylene (PE) films incorporating TiO2 demonstrated strong antibacterial activity against E. coli and S. aureus under UV-A (320–400 nm) radiation, with inhibition rates of 89.3% and 95.2%, respectively, after 60 min of irradiation. A similar result was reported by Rtimi et al. (2015). In their study, radiofrequency plasma pretreatment improved the amount and adhesion of TiO2 sputtered on PES and PE biofilms, and consequently increased the antimicrobial activity against E. coli (1  106 CFU mL
1) under dark and visible UV (UV-Vis) light conditions. The best results were obtained with PES–TiO2 biofilm: a sample without pretreatment inactivated E. coli within 5 h of UV-Vis irradiation, whereas with radiofrequency plasma pretreatment of a PES–TiO2 sample for 30 min the inactivation time was reduced to 1.5 h. The increased activity with radiofrequency–plasma pretreatment was due to increases in hydrophilicity, the number of surface active sites, and photogenerated ROS, especially •OH. Azizi et al. (2014) prepared biofilms from zinc oxide–silver (ZnO–Ag) nanoparticles stabilized with CNCs as multifunctional nanosized fillers in poly (vinyl alcohol)/chitosan (PVA/Cs) matrices and evaluated their bactericidal properties against S. choleraesuis and S. aureus (1  108 CFU mL
1) with incubation for 24 h under UV light. The films presented excellent antimicrobial properties, and the higher the coverage content of the ZnO–Ag nanoparticles, the larger the inhibition zones of the two bacteria. With CNC/7% ZnO–Ag biofilm, 8.3- and 6.0-mm inhibition zones were obtained for S. aureus and S. choleraesuis, respectively. The lower resistance of S. aureus, compared with that of S. choleraesuis, is related to the different structures and cell membrane chemical compositions of these microorganisms. Anatase TiO2 thin films prepared on a polymeric substrate (fluorinated ethylene propylene copolymer) showed photocatalytic antimicrobial activity against S. aureus (8.7  105 CFU mL
1) under UV-A radiation for 5 h, according to a study by Doran et al. (2018). The best result was obtained with film annealed at 150  C, which reduced the number of bacterial colonies by 85% (1.3  105 CFU mL
1). Photoactivation of the anatase produced ROS through the photocatalytic process; these ROS have the ability to break down the phospholipid bilayer of bacterial cells,

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which causes exposure of their internal organelles and promotes bacterial inactivation. UV radiation may also be responsible for some cellular damage, although UV-A (320–400 nm) radiation is less effective than UV-B (280–320 nm). In relation to films that have photocatalytic activity under UV-Vis radiation and antimicrobial activity in the absence of light, Wojcieszak et al. (2016) studied TiO2 fine films with anatase or rutile structures deposited on silica (SiO2) substrates. The bactericidal effects of these materials on E. coli and S. aureus and their fungicidal effects on C. albicans were evaluated in the dark. Weak antibacterial activity (with only a 1-log reduction after 8 h of contact), which was considered bacteriostatic, was observed only with the TiO2 rutile film tested against S. aureus. An increase in the surface area and a reduction in the surface hydrophilic properties, which were results of the finer crystalline structure of the TiO2 rutile film, were responsible for the antimicrobial activity of this film. TiO2 anatase film reduced S. aureus and C. albicans within the first 4 h of incubation, but, with extended exposure, it subsequently promoted the growth of these microorganisms. Thus, in relation to photocatalytic films and biofilms with high antimicrobial activity, it is possible to consider silver- and copper-doped potassium hexaniobate film and silver orthophosphate, titanium dioxide, and magnetite film activated with visible radiation, and nitrogen-doped SnO2/TiO2 glass fiber thin film activated with UV radiation.

9.5

Photocatalytic Composites and Nanocomposites

Composites and nanocomposites can be prepared by combination of organic–inorganic and inorganic–inorganic materials with use of different synthesis methodologies. Generally, the organic materials used are cellulose, polyurethane (PU), PE, and polyacrylonitrile (PAN), among others, and the inorganic materials used are oxides, sulfides, and metal particles. The advantage of working with organic–inorganic material, which combines different organic materials and biocidal metal particles, is related to control of the metal particle size by not agglomerating these particles during the synthesis. Many of these metal particles can be adsorbed on the surface of the organic material because of electrostatic interactions between the oxygen atoms of the polar hydroxyl and the metal particles, which favor dispersion (Lu et al. 2008; Azizi et al. 2013; Bui et al. 2017). However, when metallic particles are dispersed on oxides (inorganic–inorganic materials) to avoid the agglomeration of these particles, there is an optimal content for deposition, and higher content does not always promote greater antimicrobial activity. The nanocomposite can undergo the antibacterial process synergistically, caused by strong interaction between the metallic particles and the oxides (Lu et al. 2008). According to Lu et al. (2008), (inorganic–inorganic) Ag/ZnO nanocomposites combine the characteristics of their constituent materials in a synergistic way, involving the efficacy of silver against Gram-negative bacteria and ZnO against Gram-positive bacteria. In addition, there is greater interaction

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between bacteria and the nanocomposite because of electrostatic forces, since silver nanoparticles in the Ag/ZnO nanocomposite are more positive than those in the pure Ag. Hierarchical ZnO nanorod (ZNR)–deposited polydopamine (Pdopa)– functionalized PU nanofibers were evaluated for inactivation of E. coli (3  106 CFU mL
1) by Kim et al. (2018), using low-intensity UV-LED [lightemitting diode] light. A Pdopa–ZNR/PU nanocomposite was most efficient, reducing the cellular concentration to 1  104 CFU mL
1 within 180 min, and the antibacterial activity was maintained even after three cycles of use. Zapata et al. (2012) used PE and TiO2 nanoparticle nanocomposites with and without modification with hexadecyltrimethoxysilane against E. coli (1.5  107 CFU mL
1) under white and UV light for 60 and 120 min. The PE/TiO2 (8 wt%) nanocomposite showed a strong bactericidal effect (a 3-log reduction) only under UV radiation, whereas the PE/Mod–TiO2 (8 wt%) nanocomposite showed activity with both types of radiations and eliminated 99% of the bacteria within 60 min of exposure to UV-A radiation. A Z-scheme heterojunction photocatalyst (TiO2/rGO/WO3) was formed by titanium dioxide (TiO2) and tungsten(VI) oxide (WO3) anchored on reduced graphene oxide (rGO), as described by Zeng et al. (2017). The TiO2/rGO/WO3 nanocomposite was more effective (inactivating 97.3% of E. coli (2  103 CFU mL
1)) than a TiO2/ WO3 nanocomposite under simulated sunlight for 80 min and in the absence of light. This higher efficiency of the TiO2/rGO/WO3 photocatalyst was attributed to reduction of the e
/h+ recombination process and an increase in oxygen reduction reactions due to the presence of rGO. According to Fatimah and Yudha (2016), a combination of silver, ZnO, and smectite clay minerals to form a Ag/ZnO–hectorite composite favored inactivation of E. coli (1801 CFU mL
1) under UV illumination for 10 min. In the illuminated condition, Ag/ZnO–hectorite and ZnO–hectorite caused 99.7% and 99.1% killing of the bacteria, respectively, while the inactivation rate with UV radiation alone (photolysis) was only 18.8%, confirming the antimicrobial characteristics of these composites. A similar result was reported by Wu et al. (2010), who achieved almost 100% killing of E. coli (1–4  108 CFU mL
1) using a montmorillonite (MMT)– supported Ag/TiO2 composite (Ag/TiO2/MMT) under visible light for 3 h. The high antimicrobial activity, stability, and recyclability of this composite were attributed to adsorption of the bacteria on the surface layer of the MMT due to its strong adsorption capacity. Some composites, like nanoparticles, have photocatalytic activity under UV and visible radiation, but their antimicrobial activity has been evaluated in the dark. Rehana et al. (2017) tested ZnO/MS (M ¼ zinc (Zn), cadmium (Cd), or lead (Pb) in different shell concentrations) core/shell nanoparticles as antimicrobial agents against E. coli, K. pneumoniae, Enterococcus faecalis, and S. aureus bacteria, and against A. niger and C. albicans fungi (all at 5  105 CFU mL
1). All samples exhibited moderate to weak activity against bacteria and fungi in the dark, but each microorganism was less resistant to a specific material; e.g., E. coli and K. pneumoniae were inactivated by ZnO/CdS (0.06 mg L
1) and ZnO/PbS

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(0.02 mg L
1), respectively, whereas ZnO/ZnS (0.02 mg L
1) was efficient against E. faecalis. The core/shell morphology favored antimicrobial activity, since it combined the activity of the metals used in the nanocomposite, besides avoiding agglomeration of metal nanoparticles and leaching of metals ions into the medium. Khan et al. (2014) prepared silver-doped cerium(IV) oxide (Ag@CeO2) nanocomposites, using electrochemically active biofilms (EABs) as a reducing agent, and these materials showed antimicrobial activity against E. coli and P. aeruginosa. An increase in the concentration of the nanocomposite (from 1 to 3 mM) increased the inhibition zones of E. coli (from 0.5 to 1 mm) and P. aeruginosa (from 1.5 to 2 mm), whereas pure CeO2 showed no antimicrobial activity. The effectiveness of the Ag@CeO2 nanocomposite against bacteria could be attributed to Ag+ ions that are released and then absorbed by the cell surface, damaging the cell membrane and killing the bacteria. In relation to the survival rate of the microorganisms, there was basically no difference between the 1- and 3-mM concentrations of the Ag@CeO2 nanocomposite, but the survival rate varied as a function of the microorganism: 10% for E. coli and 0.1% for P. aeruginosa. Saud et al. (2016) demonstrated the bactericidal activity of a silver vanadate(V) (Ag3VO4) nanoparticle PAN nanofiber composite against E. coli in the dark, and the antibacterial effect was attributed to the contact between the composite and the bacteria. Lu et al. (2008) synthesized ZnO nanoparticles and Ag/ZnO nanocomposites with different Ag content (0.61–5.15 at%), using tyrosine as a shape-directing agent for ZnO faceted nanorods and as a reducing agent for Ag+ ions. The authors evaluated the antimicrobial activity of these materials against E. coli and S. aureus (1  108 CFU mL
1), and they obtained the best result with 1.2 at% Ag/ZnO, which demonstrated high antibacterial capacity against both species. The minimal inhibitory concentrations (MICs) of 1.2 at% Ag/ZnO against E. coli and S. aureus were 600 and 400 μg mL
1, respectively. These values were much lower than those for ZnO (3500 and 1000 μg mL
1, respectively), indicating the high toxicity of the nanocomposites. Among photocatalytic nanocomposites and composites, titanium dioxide and tungsten(VI) oxide anchored on an rGO nanocomposite presented high antimicrobial activity with mortality higher than 97% under solar radiation, and with an MMT-supported Ag/TiO2 composite under visible radiation, 100% killing was achieved.

9.6

Materials with Antimicrobial Activity in the Absence of Light

In the literature there are several reports on antimicrobial activity in the dark with use of various types of antimicrobial agents. Table 9.1 lists some materials that have antimicrobial activity in the dark and are potential photocatalysts; however, their photocatalytic activity was not evaluated in those reports.

Klebsiella pneumoniae, Staphylococcus aureus Escherichia coli, Pseudomonas aeruginosa, Salmonella enterica, Bacillus cereus Escherichia coli, Listeria monocytogenes, Staphylococcus aureus Staphylococcus aureus, Bacillus cereus, Escherichia coli, Salmonella typhimurium

ZnO nanoparticles

Escherichia coli, Staphylococcus aureus Staphylococcus aureus

Escherichia coli, Staphylococcus aureus Escherichia coli, Staphylococcus aureus, Pseudomonas aeruginosa, Bacillus thuringiensis

Metallic copper nanoparticles (Cu2+ ions reduced with a plant extract) Cadmium oxide (CdO) nanoparticles

Silver and silver tetratungstate (Ag@Ag8W4O16)– nanoroasted rice beads Leaf extract–capped cadmium sulfide (CdS) nanoparticles

Metallic silver nanoparticles (Ag+ ions reduced with a plant extract)

Erbium (Er)–doped ZnO and neodymium (Nd)– doped ZnO nanoparticles

ZnO and ZnO–Al2O3 nanoparticles

Microorganism Klebsiella pneumoniae

Antimicrobial agent Zinc oxide (ZnO) nanoparticles

The inhibition zone was 19 mm with 20 mg mL
1 of CdO, and complete elimination was observed within 30 h of contact with use of 20 μg mL
1 of CdO The inhibition zones were 19 mm for E. coli and 20 mm for S. aureus with use of 100 μg mL
1 CdS had strong bactericidal activity, with inhibition zones of 31 mm for S. aureus and 24 mm for E. coli due to the presence of the bioactive phase in CdS, which generated more free radicals

Nd-doped ZnO demonstrated higher activity, and E. coli was the least resistant microorganism (inhibition zone 28 mm) The amount of extract was inversely related to the particle size and directly related to the antimicrobial activity; higher activity was obtained with Ag nanoparticles reduced with 4 mL of extract (inhibition zone 10.4–10.8 mm) The inhibition zones were between 8.3 and 13.3 mm

Main results ZnO (0.75 mM) inhibited the growth of K. pneumoniae within 4 h of contact, and the minimal inhibitory concentration was 40 μg mL
1 The inhibition zones were 11 mm for K. pneumoniae and 10 mm for S. aureus ZnO–Al2O3 demonstrated higher activity than ZnO, and a concentration of 10 μg mL
1 inhibited E. coli within 6 h of contact

Table 9.1 Materials that exhibit antimicrobial activity in the dark and are potential photocatalysts

(continued)

Selvamani et al. (2016) Ayodhya and Veerabhadram (2017)

Michael et al. (2017) Salehi et al. (2014)

Benakashani et al. (2017)

Raza et al. (2016)

Janaki et al. (2015) Şahin et al. (2017)

References Reddy et al. (2014)

9 Antimicrobial Activities of Photocatalysts for Water Disinfection 233

Copper(II) oxide (CuO), nickel(II) oxide (NiO), zinc oxide (ZnO), and antimony(III) oxide (Sb2O3) nanoparticles Zinc-substituted cobalt ferrite (Co(1
x)ZnxFe2O4) [x ¼ 0, 0.3, 0.5, 0.7, and 1] ZnO nanoparticles/cellulose nanocrystal (CNC) nanocomposites Salmonella enterica serotype Choleraesuis, Staphylococcus aureus (1  108 CFU mL
1)

Staphylococcus aureus, Bacillus subtilis, Pseudomonas aeruginosa, Escherichia coli (1  106 CFU mL
1) Escherichia coli, Bacillus subtilis, Staphylococcus aureus Escherichia coli

Zinc oxide (ZnO), copper(II) oxide (CuO), and hematite (Fe2O3)

Copper (II) oxide (CuO in the nanosheets, bulk microsized and nanopowder forms)

Microorganism Staphylococcus aureus, Pseudomonas aeruginosa, Escherichia coli, Candida albicans (1  106 CFU mL
1) Aspergillus niger, Trichophyton rubrum (2  106 CFU mL
1) Escherichia coli (1  106 CFU mL
1)

Antimicrobial agent ZnO nanoparticles biosynthesized using a plant extract

Table 9.1 (continued)

Gilbertson et al. (2016)

The antimicrobial activity of CuO was related to the mechanism of inactivation: a single active site (CuO nanosheets) and double active sites (CuO nanopowder); nanosheets showed 100% inactivation of E. coli with 0.04 mg mL
1 and a rate constant 3–4 times higher than that of CuO nanopowder For B. subtilis the inhibition zones were 25 mm with ZnO, 21 mm with CuO, and 15 mm with Fe2O3; E. coli was the most resistant microorganism

Baek and An (2011) Sanpo et al. (2012) Azizi et al. (2013)

The order of toxicity was CuO > ZnO (except for S. aureus) > NiO > Sb2O3 The bacterial survival rate was 50% with CoFe2O4 and 30% with ZnFe2O4 The lower the mass ratio of Zn:CNC, the smaller the particle size, the higher the bandgap energy, and the higher the activity (optimum ratio 2:4); the nanocomposite demonstrated its highest activity against S. aureus

Azam et al. (2012)

References Dobrucka et al. (2018)

Main results S. aureus exhibited the highest sensitivity to ZnO nanoparticles, which were more toxic to prokaryotic cells than to more sophisticated eukaryotic cells

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Escherichia coli, Staphylococcus aureus (4  106 CFU mL
1)

Aeromonas, Bacillus subtilis, Klebsiella, Pseudomonas aeruginosa Escherichia coli, Staphylococcus aureus

Graphene oxide coated with silver nanoparticles (GO–Ag)

Graphene oxide and zinc oxide (GO–ZnO)

CFU colony-forming units

ZnO/silica gel

ZnO/bentonite

Escherichia coli (1  1011 CFU mL
1) Escherichia coli (2.2  106 CFU mL
1), Staphylococcus aureus (1.9  106 CFU mL
1), Candida albicans (9.9  105 CFU mL
1) Escherichia coli

Alginate–zirconium(IV) phosphate (AG/ZPNC) ion exchanger Cu(II) and Zn(II)–doped hydroxyapatite (HAP) nanopowders

100% CFU reductions were achieved in S. aureus within 10 min and in E. coli within 40 min, with inhibition zones of 22.87 mm and 18.66 mm, respectively, within 40 min of contact

ZnO/bentonite demonstrated a strong antimicrobial effect with mortality higher than 99.94%, while pure bentonite achieved only 1.33% mortality With 100 μL of GO–Ag the inhibition zones were 18 mm and 17.2 mm for E. coli and S. aureus, respectively; the activity was attributed to oxidative damage caused by reactive oxygen species and Ag+ ions released from Ag nanoparticles The increase in bacterial inactivation was due to a reduction in the concentration of the microorganism and an increase in the concentration of GO–ZnO

There was a 3-log reduction in E. coli after 6 h of contact with 3 mg mL
1 of AG/ZPNC With CuHAP2 the inhibition zones were 2 and 3 mm for E. coli and C. albicans, respectively; all materials promoted 2-log reductions in cell viability for both bacteria, but ZnHAP2 was more efficient

Lotfiman and Ghorbanpour (2017)

Sumalatha et al. (2017)

Shao et al. (2015)

Pouraboulghasem et al. (2016)

Pathania et al. (2017) Stanić et al. (2010)

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Hrenovic et al. (2012) investigated the antimicrobial activity of copper(IV) oxide (Cu2O), zinc oxide (ZnO), and nickel(II) oxide (NiO) nanoparticles supported on natural clinoptilolite (NZ) against E. coli and S. aureus in pure cultures and against native E. coli in real effluent water from the second stage of a biological wastewater treatment plant. The pure cultures presented bacterial concentrations between 1  106 and 1  107 CFU mL
1, while the real effluent presented a more complex composition—4.50  104 CFU mL
1 of heterotrophic bacteria, 4.88  103 CFU mL
1 of total coliform bacteria, 1.35  103 CFU mL
1 of fecal coliform bacteria, and 2.10  102 CFU mL
1 of fecal streptococci—although monitoring was done only in relation to native E. coli (1.55  102 CFU mL
1). NiONZ nanoparticles showed weak antimicrobial activity, whereas Cu2ONZ and ZnONZ nanoparticles reduced the numbers of viable E. coli cells by 87% and 54%, respectively, and viable S. aureus cells by 94% and 68%, respectively, in pure cultures within 24 h of contact, and they showed 100% antibacterial activity against native E. coli after 1 h of contact over 48 exposures. Antimicrobial agents such as nanoparticles can also be used in combination with antibiotics to enhance the bactericidal effect. Banoee et al. (2010) evaluated the antibacterial activity of ten different antibiotics in the absence and presence of ZnO nanoparticles (500 μg/disk) against S. aureus and E. coli. The authors observed that the bactericidal activity of amoxicillin, penicillin G, and nitrofurantoin against S. aureus was reduced in the presence of ZnO, while the nanoparticles favored the bactericidal property of ciprofloxacin against both microorganisms, achieving 27% and 22% reductions in the inhibition zones of S. aureus and E. coli, respectively. With increasing concentrations of ZnO nanoparticles, the inhibition zones increased (63% for S. aureus and 93% for E. coli, using 2000 μg/disk). The positive interaction between ciprofloxacin and ZnO nanoparticles against S. aureus may be related to ionic interactions between the protonated nitrogen atoms of quinolone present in the three amino groups with the electron donor fluoro group and the hydroxylated surface of the ZnO. The increase in the activity of the antibiotic–ZnO against S. aureus could be attributed to three factors: (1) interference of nanoparticles with the pumping action of NorA protein in S. aureus, which controls the active efflux of hydrophilic fluoroquinolones from the cell; (2) ability of nanoparticles to favor electron transfer kinetics at the active sites of the enzymes; and (3) ability of nanoparticles to enhance the absorption of the antibiotic into bacterial cells.

9.7

Case Study: Application of Supported Photocatalysts in Disinfection of Whey-Processing Water

We prepared hematite (Fe2O3) and titanium dioxide (TiO2) supported on a glass sphere for use as a photocatalyst in disinfection of whey-processing water under UV radiation. Initially, the glass sphere was submitted to basic treatment with sodium hydroxide (5 mol L
1) for 24 h, and an Fe2O3 and TiO2 aqueous dispersion was

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added (mass ratio of the sphere: dispersion ¼ 5:1 and Fe2O3: TiO2 ¼ 1:9). Afterward, the solvent was evaporated in a rotary evaporator and the material was calcined at 400  C for 5 h. This process generated a Fe2O3–TiO2/glass photocatalyst with a supported active phase of 6.3%. Photocatalytic disinfection was applied to wastewater samples from a membrane filtration system with reverse osmosis, used in fractionation and concentration of whey components. From the third stage, two permeate samples (denoted as effluent samples 1a and 1b) and two retentate (concentrate) samples (denoted as effluent samples 2a and 2b) were collected, and the collections were carried out on two different days. The Fe2O3–TiO2/glass photocatalyst (4 g) was added to each effluent sample (500 mL), and the suspension was irradiated with UV light (250 W) for 3 h. The antimicrobial activity of the photocatalyst was estimated by the reduction in the total number of microorganisms in Plate Count Agar (PCA) culture medium, and the incubation temperature was 30  2  C. After the photocatalytic treatment, two samples of the effluents were stored for 12 days, one sample at room temperature (25–30 oC) and the other sample was kept at 5 oC, and then reanalyzed. The results are expressed in CFU per plate. We found a significant variation in the effluent microbial load between the two collected samples: on the first day, effluent samples 1a and 2a had counts of 6000 and 1300 CFU mL
1, respectively. On the second day, both effluent samples (1b and 2b) had significantly increased numbers of microorganisms: a microbial load higher than 3  105 CFU mL
1. This variation was attributed to the initial raw material whey. The Fe2O3–TiO2/glass photocatalyst showed high bacteriostatic activity, reducing microbial counts in all samples: effluent samples 1a (182 CFU mL
1), 1b (110 CFU mL
1), 2a (110 CFU mL
1), and 2b (126 CFU mL
1). This reduction was attributed to high UV light intensity, formation of ROS, and the Fe2O3-TiO2/ glass photocatalyst—more specifically, the effect of metal ions released in the solution, once 0.3% Fe2O3–TiO2 was detached from the support and dissolved in the effluents. Upon reanalysis after 12 days of treatment, the effluent showed microbial growth of approximately 2.2  105 CFU mL
1 in the samples stored at ambient temperature, while those maintained at a temperature of 5  C showed the absence of microbial growth. These results showed that the surviving microorganisms were of a mesophilic type. We concluded that the dose of UV radiation (250 W) and the exposure time (3 h) were not sufficient to destroy all of the microbial load existing in the effluent samples. This disinfection process showed irreversible damage to the cell wall and its components only for the group of psychrophilic microorganisms; thus, the Fe2O3–TiO2/glass photocatalyst had a biocidal effect on this group. However, after a period of stress, the mesophilic microorganisms were able to recover and multiply. Lanao et al. (2012) attributed this type of behavior to a latency state; after a period in the dark, mesophilic microorganisms regain their capacity for growth and reproduce again. Thus, we concluded that the treatment applied to the group of mesophilic microorganisms was not efficient in the conditions tested; therefore, the photocatalysis process had only a bacteriostatic effect on this group.

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Final Considerations

On the basis of the considerations described in this chapter, there is strong global interest in development of photocatalysts for use in photocatalytic disinfection of water and wastewater, especially those activated with visible (solar) radiation. In addition to this, the disinfection process does not generate by-products that are harmful to human health and the environment, being in accordance with the principle of green chemistry. In recent decades there has been a significant increase in research on antimicrobial activity of photocatalysts, which has shown that conditions involving ultraviolet and visible radiation increase generation of reactive oxygen species. However, there is still no consensus on the best type of photocatalyst to use and its optimum concentration. This is because the published research studies have used different concentration conditions, species of microorganisms, catalyst concentrations, and synthesis methodologies. Among the different microorganisms that have been studied, Staphylococcus aureus and Escherichia coli have been most frequently used, and in relation to the microbial load, the range of application has varied considerably (from 1  103 to 3  1012 CFU mL
1), which has caused large variations in the efficiency of inactivation and the inhibition zones (5–33 mm). To date, few studies have used native microorganisms and been performed under real conditions, such as disinfection of drinking water for human consumption, disinfection of wastewater and secondary sanitary sewage for reuse, and disinfection of industrial effluent. Although antimicrobial activity differs among photocatalysts, the photocatalytic disinfection mechanism is well defined in the presence of light; however, with regard to this mechanism in the absence of light, deeper investigation into production of reactive oxygen species is still necessary so that the toxicity of the mechanism can be defined in these conditions. Photocatalytic disinfection with solar light is an economically viable and competitive alternative to the conventional methods currently used to treat wastewater and drinking water; however, the application of this technology on a large scale is still lacking.

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Wong M-S, Chen C-W, Hsieh C-C et al (2015) Antibacterial property of Ag nanoparticle– impregnated N-doped titania films under visible light. Sci Rep 5:11978. https://doi.org/10. 1038/srep11978 Wu T-S, Wang K-X, Li G-D et al (2010) Montmorillonite-supported Ag/TiO2 nanoparticles: an efficient visible-light bacteria photodegradation material. ACS Appl Mater Interfaces 2 (2):544–550. https://doi.org/10.1021/am900743d Xing Y, Li X, Li Z et al (2012) Effect of TiO2 nanoparticles on the antibacterial and physical properties of polyethylene-based film. Prog Org Coat 73(2–3):219–224. https://doi.org/10. 1016/j.porgcoat.2011.11.005 Xu J-W, Gao Z-D, Han K et al (2014) Synthesis of magnetically separable Ag3PO4/TiO2/Fe3O4 heterostructure with enhanced photocatalytic performance under visible light for photoinactivation of bacteria. ACS Appl Mater Interfaces 6(17):15122–15131. https://doi.org/ 10.1021/am5032727 Yadav HM, Kim J-S, Pawar SH (2016) Developments in photocatalytic antibacterial activity of nano TiO2: a review. Korean J Chem Eng 33(7):1989–1998. https://doi.org/10.1007/s11814016-0118-2 Yuan Y, Ding J, Xu J et al (2010) TiO2 nanoparticles co-doped with silver and nitrogen for antibacterial application. J Nanosci Nanotechnol 10(8):4868–4874. https://doi.org/10.1166/ jnn.2010.2225 Zapata PA, Palza H, Delgado K et al (2012) Novel antimicrobial polyethylene composites prepared by metallocenic in situ polymerization with TiO2-based nanoparticles. J Polym Sci A Polym Chem 50(19):4055–4062. https://doi.org/10.1002/pola.26207 Zeng X, Wang Z, Wang G et al (2017) Highly dispersed TiO2 nanocrystals and WO3 nanorods on reduced graphene oxide: Z-scheme photocatalysis system for accelerated photocatalytic water disinfection. Appl Catal B Environ 218:163–173. https://doi.org/10.1016/j.apcatb.2017.06.055 Zhang L, Ding Y, Povey M et al (2008) ZnO nanofluids—a potential antibacterial agent. Prog Nat Sci-Mater 18(8)):939–944. https://doi.org/10.1016/j.pnsc.2008.01.026

Chapter 10

Medicinal Applications of Photocatalysts Busra Balli, Aysenur Aygun, and Fatih Sen

Contents 10.1

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.1.1 Background . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.2 Antifungal Activity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.3 Virucidal Activity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.4 Antimicrobial Activity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.5 Anticancer Activity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.6 Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract With the help of nanotechnology and nanoscience, future devices and technologies will be small and possess advanced features; photocatalysis is an important area of application. This chapter discusses the importance of photocatalysts and their medicinal applications for human beings in daily life. The properties of photocatalysts as a result of their nanoscale are discussed. The main medicinal applications of photocatalysts—antifungal, antimicrobial, anticancer, and several another applications—are also focused on in detail. Keywords Antifungal · Antimicrobial · Anticancer · Carbon nanotube · Graphene · Inhibition · Medicine · Nanoscience · Nanomaterial · Nanotechnology · Photocatalyst

B. Balli · A. Aygun · F. Sen (*) Sen Research Group, Biochemistry Department, Faculty of Arts and Science, Dumlupınar University, Kütahya, Turkey e-mail: [email protected] © Springer Nature Switzerland AG 2020 Inamuddin et al. (eds.), Nanophotocatalysis and Environmental Applications, Environmental Chemistry for a Sustainable World 30, https://doi.org/10.1007/978-3-030-12619-3_10

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Antimicrobial agents are usually used in medical treatment of microbial infections. They are derived from various different substances of natural, semisynthetic, or synthetic origin. They can successively kill or inhibit the growth of microorganisms while causing little or no damage to the host cell. Antimicrobial agents can be classified according to the types of microorganisms that they act against, such as bacteria (antibacterials), viruses (antivirals), fungi (antifungals), and protozoa (antiprotozoals). In general, agents that kill microbes are termed microbicidal, while those agents that merely inhibit microbial growth are termed biostatic. Antimicrobial chemotherapy is the use of antimicrobial medicines to treat infection, whereas antimicrobial prophylaxis is the use of antimicrobial medicines to prevent infection. Disinfection plays an important role in the control of pathogens and microbial species in water, and can prevent waterborne epidemics and the spread of infectious disease. Wastewater is severely contaminated and contains high levels of microorganisms such as bacteria, fungi, protozoa, viruses, and algae, which greatly affect the quality of the water. Hence, it is important to clear the water along with the organic compounds it contains; therefore, sterilization of water is carried out in biological and biochemical industries by the use of some important and essential technologies. A traditional and well-known water disinfection method is the addition of chlorine to groundwater. However, the chlorination of groundwater containing high total organic carbon produces high levels of trihalomethanes and also some carcinogenic disinfection by-products. Additionally, highly resistant pathogens such as Cryptosporidium and Giardia cannot be effectively inactivated at normal dosages used for water treatment applications. As a consequence, inactivation of organisms and decomposition of organic compounds need to be performed in such a way that the formation of disinfection by-products is minimized, if not completely stopped. Many approaches such as the use of chlorine dioxide, ozonation, ultraviolet (UV) irradiation, advanced filtration processes, and photocatalytic oxidation have been tried for this purpose. Among these approaches, photocatalytic oxidation is considered to be the most convenient option because of its eco-friendly nature and low cost for inactivation of microorganisms. Among the alternative processes currently in development, photocatalysis has attracted interest for the potential to use sunlight to achieve a variety of chemical reactions for disinfection of drinking water and protection from microorganisms. Photocatalytic techniques also have potential for widespread applications such as in indoor air and environmental health, and in the biological, medical, laboratory, hospital, pharmaceutical, and food industries, as well as in plant protection, etc. (Gamage and Zhang 2010; Ameta and Ameta 2017; McEvoy and Zhang 2014). The general principles of the photocatalysis mechanism are shown in Fig. 10.1. The photocatalysis method was first described by Matsunaga et al. (1985) for effective sterilization of drinking water by inactivation of three microbial cells (Lactobacillus acidophilus, Saccharomyces cerevisiae, and Escherichia coli) by

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Fig. 10.1 General principles of the photocatalytic reaction mechanism

using a photoelectrochemical process in the presence of Pt/TiO2. Many other researchers have since reported different types of water disinfection techniques with the use of photocatalysis methods. By far the most well-studied microorganism in the literature is E. coli, which is largely used as an indicator organism in photocatalytic drinking water disinfection systems (Maness et al. 1999; Marugán et al. 2008; McLoughlin et al. 2004; Rincon and Pulgarin 2004; Rodrigues et al. 2007; Srinivasan and Somasundaram 2003; Sunada et al. 1998; Watts et al. 1995; Wei et al. 1994; Choi and Kim 2000). Many other microorganisms (bacteria and yeasts) have been investigated since then—such as Candida albicans, Enterococcus faecium, Listeria monocytogenes, Staphylococcus aureus, Pseudomonas aeruginosa, Salmonella enterica, Vibrio parahaemolyticus, Enterococcus faecalis (formerly known as Streptococcus faecalis), and Streptococcus mutans—as indicator microorganisms in photocatalytic water disinfection systems. In addition, Cryptosporidium and Giardia (which are known for their resistance to many chemical disinfectants, including chlorine) have also been efficiently inactivated by photocatalysis, as reported by a number of researchers (Cho and Yoon 2008; Ryu et al. 2008; Méndez-Hermida et al. 2007; Navalon et al. 2009; Lonnen et al. 2005). Nanomaterials offer very promising solutions to the ongoing problems of the world. Fuel cells (Eris et al. 2018a, b; Sen et al. 2014), materials for catalysis (Günbatar et al. 2018; Göksu et al. 2018), capacitors (Chen and Dai 2013), solar cells (Demir et al. 2017), thermopower applications (Abrahamson et al. 2013), and sensors (Bozkurt et al. 2017; Koskun et al. 2018) are some examples of these solutions. These materials are also used collectively to combine their superior properties. Metal–metal combinations, polymer–metal combinations, and their hybrids with carbon-based materials are used for a variety of nanomaterial applications (Sen et al. 2018a, b; Celik et al. 2016a, b). Composite materials are being widely studied to develop appropriate materials for asymmetric supercapacitors, such as conductive polymers and other electronically active material composites.

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Graphene and graphene oxide (Yildiz et al. 2017a, b; Akocak et al. 2017), carbon nanotubes (CNTs) (Sen et al. 2018a, b; Celik et al. 2016a, b, Goksu et al. 2016), activated carbon (AC) (Yildiz et al. 2017a, b), vulcanized carbon (VC) (Eris et al. 2018a, b), carbon black (Baskaya et al. 2017), and graphene-derived materials— which have different structures and morphologies, such as reduced graphene oxide (Eris et al. 2018a, b), graphene nanosheets, graphene nanoribbons, and graphene nanoplatelets—are considered active materials for different applications. In particular, nanoparticles are used in a wide variety of applications (Sahin et al. 2018). Nanosized photocatalytic materials are now commonly used, and these can be applied in many fields, one of which is the treatment of microbial infections. Nanoparticles have been employed to control the formation of biofilms in the oral cavity as a function of their biocidal, antiadhesive, and delivery capabilities, and also have been used in prosthetic devices, as topically applied agents, and in dental materials. Numerous nanosized and modified nanoparticles have been extensively investigated as potent antibacterial, antifungal, antitumor, and antiviral agents because of their very small sizes, high surface-to-volume ratios, and broad spectra of biological actions. Some particular nanoparticulate systems—such as metal nanoparticles (Ag, Au, Cu, Pd, etc.) and their oxides (Cu2O, TiO2, Y2O3, and CeO2), and biodegradable hybrid nanoparticles such as dextran—have also been explored for their antibacterial effects. It has been observed that the antimicrobial activity of a photocatalyst is enhanced when it is used in a nanometric size range (1–100 nm). These nanosized antimicrobial agents inactivate microorganisms more successfully than their micro- or macrosized counterparts (Dalrymple et al. 2010). The prepared materials are mostly used in applications such as antimicrobials, anticancer agents, etc., because they have a small ratio/surface ratio and exhibit antifungal, antialgal, antibacterial, and antiviral effects over a wide range. The latest research publications have been focused on metal, metal oxide, and biodegradable hybrid nanoparticles. The metal oxide nanomaterials are divided into two different groups based on the mechanisms involved in their growth inhibition of microorganisms. Recent developments may help to functionalize nanoparticles and eliminate antimicrobial resistance (Halbus et al. 2017). Nanodimensional photoreactive materials are widely employed in many areas, one of which is treatment of bacterial diseases. The antibacterial action of a photoreactive material is improved when its size is within the range of 1–100 nm. Various nanodimensional and altered photoreactive materials are utilized as antibacterial and antitumor agents. These nanodimensional antibacterial agents are able to pacify microorganisms better than micro- or macrodimensional agents. UV purification utilizes solar radiation (