Remediation of Chromium-Contaminated Soil: ​Theory and Practice​ (Environmental Science and Engineering) 9819954622, 9789819954629

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Environmental Science and Engineering

Weichun Yang · Liyuan Chai · Zhihui Yang · Feiping Zhao · Qi Liao · Mengying Si

Remediation of ChromiumContaminated Soil: Theory and Practice

Environmental Science and Engineering Series Editors Ulrich Förstner, Buchholz, Germany Wim H. Rulkens, Department of Environmental Technology, Wageningen The Netherlands

The ultimate goal of this series is to contribute to the protection of our environment, which calls for both profound research and the ongoing development of solutions and measurements by experts in the field. Accordingly, the series promotes not only a deeper understanding of environmental processes and the evaluation of management strategies, but also design and technology aimed at improving environmental quality. Books focusing on the former are published in the subseries Environmental Science, those focusing on the latter in the subseries Environmental Engineering.

Weichun Yang · Liyuan Chai · Zhihui Yang · Feiping Zhao · Qi Liao · Mengying Si

Remediation of Chromium-Contaminated Soil: Theory and Practice

Weichun Yang Central South University Changsha, China

Liyuan Chai Central South University Changsha, China

Zhihui Yang Central South University Changsha, China

Feiping Zhao Central South University Changsha, China

Qi Liao Central South University Changsha, China

Mengying Si Central South University Changsha, China

ISSN 1863-5520 ISSN 1863-5539 (electronic) Environmental Science and Engineering ISBN 978-981-99-5462-9 ISBN 978-981-99-5463-6 (eBook) https://doi.org/10.1007/978-981-99-5463-6 © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature Singapore Pte Ltd. 2023 This work is subject to copyright. All rights are solely and exclusively licensed by the Publisher, whether the whole or part of the material is concerned, specifically the rights of translation, reprinting, reuse of illustrations, recitation, broadcasting, reproduction on microfilms or in any other physical way, and transmission or information storage and retrieval, electronic adaptation, computer software, or by similar or dissimilar methodology now known or hereafter developed. The use of general descriptive names, registered names, trademarks, service marks, etc. in this publication does not imply, even in the absence of a specific statement, that such names are exempt from the relevant protective laws and regulations and therefore free for general use. The publisher, the authors, and the editors are safe to assume that the advice and information in this book are believed to be true and accurate at the date of publication. Neither the publisher nor the authors or the editors give a warranty, expressed or implied, with respect to the material contained herein or for any errors or omissions that may have been made. The publisher remains neutral with regard to jurisdictional claims in published maps and institutional affiliations. This Springer imprint is published by the registered company Springer Nature Singapore Pte Ltd. The registered company address is: 152 Beach Road, #21-01/04 Gateway East, Singapore 189721, Singapore

Preface

Chromium (Cr) is a hard, steel-gray metal. Its free form is extremely rare and chromite ore in earth’s crust is the common form in nature. It is widely utilized in many industries, such as ferrous and non-ferrous alloy metal fabrication, leathertanning, chrome-plating and pigment industries. 90% of the total chrome production is consumed by metallurgical industry. Chromium has also been recognized as a toxic, mutagenic and carcinogenic metal. It is toxic to microorganism, plants, animals and humans. Anthropogenic activities, especially industrial production of chromate, electroplating and tanning leather directly and indirectly, cause Cr-pollution in wastewater, air and soil, and soil and underground water serve as the sink of Cr-pollutants. With the rapid increasing demand of Cr, the chromium pollution in soil is becoming more and more serious, which has become one of the severe factors that restrict the sustainable development of industry and harm the public health. Therefore, it is an urgent and huge challenge to effectively remediate the soils contaminated by chromium and its associated pollutants. Two stable valence states of Cr in the environment usually are Cr(III) and Cr(VI). Cr(III) has lower toxicity and mobility than Cr(VI). One of the potential treatments is to transform Cr(VI) to Cr(III). Reductive materials and microorganisms have been used to remediate Cr(VI) pollution in soil. According to the global needs of the remediation of soil contaminated by chromium, this book systematically introduces the latest theories and technical achievements of microbial and chemical treatments for the chromium pollution in the chromium slag and chromium-contaminated soil, combined with the author’s research achievements over the decades. The book focuses on the biological and chemical behaviors of chromium in soil, microbial and chemical remediation for the chromium-contaminated soil and the cases of chrome-contaminated site remediation project. This book can be used by the scientific researchers and engineering technicians engaged in chromium chemical industry and environmental protection. It can also be used as a teaching and reference book for graduate students in the majors of environmental science and engineering, soil science, chemistry and chemical engineering, et al.

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This work was supported by the Key Project of Chinese National Research Programs (2021FYC1809203 and 2018YFC1802204), Chinese National High Technology Research and Development Program (2006AA06Z374), National Natural Science Foundation of China (20477059 and 51304250), Chinese National Science and Technology Benefit Program (2013GS430203), Hunan Provincial Science and Technology Key Project (2009FJ1009 and 2008SK2007) and Natural Science Foundation of Hunan Province (2021JJ30829). The authors also sincerely thank the graduate students for their contribution to publish this book. They are Dr. Wenjie Zhu, Dr. Zemin Ma, Dr. Shunhong Huang, Dr. Zhenxin Wang, Dr. Yangyang Wang, Dr. Chunlian Ding, Canwen Sheng, Tengfa Long, Xiong Li, Rong, Deng, Kun Zhao, Lijuan Chen, Zeyou Xu, Bing Wnag, Changqing Su, Yingping Liao, Hangbin Li, Xiaoming Zhang, Qi Li, Dongdong Xi, Xiaomin Li and Jiaqi Tang, et al. Changsha, China

Weichun Yang Liyuan Chai Zhihui Yang Feiping Zhao Qi Liao Mengying Si

Contents

1 Contamination Characteristics of Soils and Biological and Chemical Behavior of Chromium in Soil . . . . . . . . . . . . . . . . . . . . . . 1.1 Sources of Soil Chromium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.1.1 Chromium Pollution from the Chromium Salt Industry . . . . 1.1.2 Chromium Pollution in the Electroplating Industry . . . . . . . . 1.1.3 Chromium Pollution in the Tanning Industry . . . . . . . . . . . . . 1.1.4 Chromium Pollution from Other Industries . . . . . . . . . . . . . . 1.2 Characteristics of Chromium Pollution in Soil . . . . . . . . . . . . . . . . . . 1.2.1 Chromium Concentrations in Chromium Pollution in Soil . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2.2 Vertical Distributions of Total Cr and Water-Soluble Cr(VI) in the Soil Profiles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2.3 Chromium Fractions in the Contaminated Soils . . . . . . . . . . . 1.3 Microbial Communities in Chromium-Contaminated Soil . . . . . . . . 1.3.1 Bacterial Structural Diversity in the Contaminated Soils . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.3.2 Isolation and Identification of Hexavalent Chromium-Resistant Bacteria . . . . . . . . . . . . . . . . . . . . . . . . . . 1.3.3 Bacterial Genetic Diversity on the Basis of Rep-Polymerase Chain Reaction Fingerprints . . . . . . . . . . 1.3.4 Hexavalent Chromium Reduction by Cr(VI)-Reducing Bacterial Strains . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.4 Adsorption of Cr(VI) on Soils . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.4.1 Adsorption Kinetics of Cr(VI) . . . . . . . . . . . . . . . . . . . . . . . . . 1.4.2 Adsorption Thermodynamics of Cr(VI) . . . . . . . . . . . . . . . . . 1.5 Cr(VI) Migration in Chromite Ore Processing Residue (COPR)-Soil . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.5.1 Diffusion of Cr(VI) in Chromium-Containing Slag . . . . . . . . 1.5.2 Cr(VI) Migration in Soils . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

1 1 2 4 5 5 6 7 9 13 16 16 19 19 19 21 21 25 28 28 33 38

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2 Mechanisms of Cr(VI) Reduction by Microorganisms . . . . . . . . . . . . . 41 2.1 Cr(VI)-Reducing Bacteria Pannonibacter Phragmitetus BB Isolated from Chromium-Contaminated Soil . . . . . . . . . . . . . . . . . . . . 42 2.1.1 Isolation and Identification of Cr(VI)-Reducing Bacterial Strain . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 42 2.1.2 Cr(VI) Reduction by Pannonibacter Phragmitetus BB . . . . . 43 2.1.3 Morphology of Pannonibacter Phragmitetus BB . . . . . . . . . 44 2.1.4 Elemental Composition of the Cr(VI)-Reduction Product . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 46 2.1.5 Identification of Cr(VI)-Reduction Product . . . . . . . . . . . . . . 46 2.2 The Cr(VI)-Reducing Bacteria Leucobacter sp. CRB1 Isolated from Chromium-Containing Slag . . . . . . . . . . . . . . . . . . . . . . 47 2.2.1 Physiological and Biochemical Characteristics of the Strain . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 47 2.2.2 The Phylogenetic Tree of Strain . . . . . . . . . . . . . . . . . . . . . . . . 47 2.2.3 Evaluation of Chromium Tolerance . . . . . . . . . . . . . . . . . . . . . 48 2.2.4 Identification of Cr(VI)-Reduction Product . . . . . . . . . . . . . . 50 2.3 Behavioral Characteristics of Cr(VI) Reduction by Bacteria . . . . . . 52 2.3.1 Cr(VI) Reduction Characteristics of Pannonibacter Phragmitetus BB . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 52 2.3.2 Cr(VI) Reduction Characteristics of Leucobacter sp. CRB1 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 55 2.4 Electrochemical Characteristics of Cr(VI) Reduction by Leucobacter sp. Ch-1 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 58 2.4.1 Effect of Initial pH Values on the Growth of Leucobacter sp. Ch-1 and Reduction of Cr(VI) . . . . . . . . . 58 2.4.2 Effect of Applied Potentials on the Growth of Leucobacter sp. Ch-1 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 60 2.4.3 Effect of Applied Potentials on the Reduction of Cr(VI) . . . 60 2.4.4 The Range of Initial pH and Eh for Leucobacter sp. Ch-1 Growth and Cr(VI) Reduction . . . . . . . . . . . . . . . . . . . . . 63 2.4.5 Potential-pH Diagram for “Leucobacter sp. Ch-1-Cr-H2 O” System . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 64 2.5 Molecular Mechanisms of Bacterial Reduction of Cr(VI) . . . . . . . . . 65 2.5.1 Multiomics Response of Pannonibacter Phragmitetus BB to Hexavalent Chromium . . . . . . . . . . . . . . . . . . . . . . . . . . 66 2.5.2 Discerning Chromate Reduce and Transport Genes of Highly Efficient Pannonibacter Phragmitetus BB . . . . . . 96 2.5.3 Dynamic Proteome Responses to Sequential Reduction of Cr(VI) and Removal of Coexisting Heavy Metals by PannonibacterPhragmitetus BB . . . . . . . . 104 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 125

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3 Microbial Remediation of Chromium-Polluted Soil . . . . . . . . . . . . . . . . 3.1 The Microbial Remediation Efficacy of Soil Contaminated with Chromium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1.1 Microbial Remediation of Cr(VI) in Soil . . . . . . . . . . . . . . . . 3.1.2 Kinetics of Cr(VI) Microbial Reduction . . . . . . . . . . . . . . . . . 3.1.3 Changes in pH and ϕh During Cr(VI) Microbial Reduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1.4 Changes in Concentrations of Fe2+ , Mn2+ , SO4 2− and NO3 − . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2 Optimization of Cr(VI) Bioremediation in Polluted Soil . . . . . . . . . . 3.2.1 Effect of Particle Size . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2.2 Effect of Spray Intensity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2.3 The Impact of the Initial Cr(VI) Content . . . . . . . . . . . . . . . . . 3.2.4 Effect of Depth . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2.5 Effect of Circulation Mode . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.3 Stability of Chromium in the Microbial Remediation Soil Under Simulated Acid Rain Leaching . . . . . . . . . . . . . . . . . . . . . . . . . 3.3.1 The pH Values of the Leachate in the Soils Subjected to Simulated Acid Rain Leaching Exhibited Variations . . . . 3.3.2 Releases of Chromium in Soils Under Simulated Acid Rain Leaching . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.3.3 Risk to Groundwater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.4 Changes in the Bacterial Community During the Bioremediation of Cr(VI)-Contaminated Soil . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.4.1 Phylogenetic Analysis of 16 rRNA Clone Libraries . . . . . . . 3.4.2 Real-Time Quantitative PCR Analysis . . . . . . . . . . . . . . . . . . 3.5 The Impact of Microbial Remediation on the Physicochemical and Biological Qualities of Cr(VI)-Contaminated Soils . . . . . . . . . . 3.5.1 Chemical Characteristics of Soils After Remediation . . . . . . 3.5.2 Changes in Enzyme Activity in Soils After Remediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.5.3 Quality of Soil Assessment . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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4 Mechanism of Chemical Reduction of Cr(VI) . . . . . . . . . . . . . . . . . . . . . 4.1 Chemical Reductive Materials for Cr(VI) . . . . . . . . . . . . . . . . . . . . . . 4.1.1 Organic Amendments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.1.2 Iron-Bearing Reductants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.1.3 Sulfur-Based Compounds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2 Enhanced Cr(VI) Removal by an “In Situ Synthesized” Iron-Based Bimetal Material . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2.1 Activity Toward Cr(VI) Removal . . . . . . . . . . . . . . . . . . . . . . . 4.2.2 Enhancement of Coexisting Cations on Cr(VI) Removal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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4.2.3 In Situ Synthesized Fe-Cu Bimetal During Cr(VI) Removal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2.4 Electrochemical Behaviors and Contribution of in Situ Synthesized Fe–Cu Bimetal on Cr(VI) Removal . . . . . . . . . . 4.2.5 Enhanced Cr(VI) Removal Mechanism by in Situ Synthesized Fe–Cu Bimetal . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3 Interaction Between Pyrite and Zerovalent Iron that Has a Higher Ability to Reduce Through Fe(II) Regeneration . . . . . . . . . 4.3.1 Incorporation of Sulfur into Fe0 . . . . . . . . . . . . . . . . . . . . . . . . 4.3.2 Removal Activity of FeS2 /ZVI Toward Cr(VI) . . . . . . . . . . . 4.3.3 Mechanism of Cr(VI) Removal by FeS2 /ZVI . . . . . . . . . . . . . 4.3.4 Sulfur Speciation Affects the Electron Transfer of FeS2 /ZVI . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.4 Synergistic Cr(VI) Reduction and Organic Pollutant Oxidative Degradation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.4.1 Simultaneous Cr(VI) Reduction and Phenol Oxidation . . . . 4.4.2 The Evolution of Cr(VI) in the FeS2 /Fe0 + PS System . . . . 4.4.3 The Roles of Surface-Bound Fe2+ . . . . . . . . . . . . . . . . . . . . . . 4.4.4 The Roles of SO4 2− in the FeS2 /Fe0 + PS System . . . . . . . . 4.4.5 Reactive Species for Phenol Degradation . . . . . . . . . . . . . . . . 4.4.6 Synergistic Redox Conversion Mechanism . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5 Chemical Remediation of Chromium-Contaminated Soil . . . . . . . . . . . 5.1 Kinetics, Thermodynamics And Long-Term Effects in Remediation of Cr(VI)-Contaminated Soil . . . . . . . . . . . . . . . . . . . 5.1.1 Factors Affecting Remediation of Cr(VI)-Contaminated Soil . . . . . . . . . . . . . . . . . . . . . . . . . . 5.1.2 Kinetics and Thermodynamics . . . . . . . . . . . . . . . . . . . . . . . . . 5.1.3 Soil pH Variation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.1.4 Soil Cr Speciation Change . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.1.5 Long-Term Stability of Cr in Remediated Soil . . . . . . . . . . . . 5.2 Remediation of Cr(VI) and Organic Pollutant Cocontaminated Soil . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2.1 Factors Affecting the Remediation of Cr(VI) and Organic Pollutants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2.2 Remediation Performance Comparison of Different Reaction Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2.3 Soil Cr(VI) Speciation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.3 Synergistic Remediation of Cr(VI) and Cationic Metal Co-contaminated Soil . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.3.1 Factors Affecting the Remediation of Cr–Cu–Ni–Co–Contaminated Soil . . . . . . . . . . . . . . . . . . . 5.3.2 Sequential Extraction, Followed by an Evaluation of Risk Reduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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5.4 Chemical-Microwave Combined Remediation of Strongly Alkaline High Cr(VI)-Contaminated Soil . . . . . . . . . . . . . . . . . . . . . . 5.4.1 Comparison of Remediation Effect by Various Amendments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.4.2 Factors Affecting Microwave Irradiation-Assisted Reduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.4.3 Mechanism of Microwave Irradiation Accelerating Cr(VI) Reduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.4.4 Cr Speciation and Phase Transformation Under Microwave Irradiation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.4.5 Long-Term Stability of Cr in Remediated Soil . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6 Case of Chrome-Contaminated Site Remediation Project . . . . . . . . . . 6.1 Project Overview . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.2 Remediation Technology and Route . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.3 Project Photos . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.4 Remediation Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Chapter 1

Contamination Characteristics of Soils and Biological and Chemical Behavior of Chromium in Soil

Chromium salt and chromium metal are the important industrial raw materials and chemicals that are widely applied in electroplating, metallurgy, anticorrosion, leather, dye and other industries. 10% of the commodity species in China are related to chromium salt products, which are mainly used in electroplating, tanning, printing and dyeing, pharmaceuticals, pigments, catalysts, organic synthesis oxidants and metal corrosion inhibition. Chromium metal is also used for superalloys, selectroresistance alloys, precision alloys and other nonferrous alloys. Superalloys containing 10–25% Cr are mainly used for manufacturing jet engines, aerospace machinery and materials, rocket engines and heat exchangers. As chromium can improve the toughness, abrasive resistance and corrosion resistance of steel, chromium is also an indispensable element in stainless steel, tool steel, bearing quality steel and other steel types. The manufacture of chromium salt and chromium metal plays a key role in the construction of the national economy but also produces serious environmental problems, such as soil and groundwater contamination [1, 2].

1.1 Sources of Soil Chromium Because of its extensive usage and potential for negative environmental impacts, the heavy metal chromium is contaminating soil ecosystems at an alarming rate. Chromium salt manufacturing, ore refining, steel and alloy production, pigment manufacturing, corrosion inhibition, leather tanning, wood preservation, and coal and oil combustion are mostly responsible for discharging chromium into the environment. The majority of Cr industry activities lack proper disposal facilities, and leaching has frequently occurred as a result of inadvertent discharges of wastes containing Cr. Yet, it has the potential to gravely harm groundwater. As a result, there are many different causes of chromium contamination in soils.

© The Author(s), under exclusive license to Springer Nature Singapore Pte Ltd. 2023 W. Yang et al., Remediation of Chromium-Contaminated Soil: Theory and Practice, Environmental Science and Engineering, https://doi.org/10.1007/978-981-99-5463-6_1

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1 Contamination Characteristics of Soils and Biological and Chemical …

1.1.1 Chromium Pollution from the Chromium Salt Industry The production of chromium salts in China has experienced a history of 40 years, and the production and consumption of chromium compounds has become the first in the world. More than 70 units have produced chromium salts in China, and the total production of chromium slag is approximately 6 million tons. Chromium salt production has grown from small to large, and its processes are constantly being improved. However, the production and use of chromium salt products produce chromium-containing wastewater, exhaust gas and slag emissions, resulting in soil and groundwater contamination of the surrounding sites. Chromium salt production is divided into two major technologies, calcium roasting and calcium-free roasting. The process and the main equipment are generally similar. The biggest difference is that calcium-free roasting does not use calcium filler, with calcium roasting slag discharge, high hexavalent chromium content, and pollution being a concern. In the production process of calcium roasting, finely ground chromite, dolomite and soda ash are mixed in a certain ratio, the mixture is added to the rotary kiln to oxidize and roast the chromium trioxide in chromite into sodium chromate, and the residual solid waste is chrome slag. Every 1 ton of sodium red alum will discharge 1.7–4.2 tons of chrome slag, and every 1 ton of chromium metal will discharge 7 tons of chrome slag. The composition and properties of chromium slag vary greatly due to the difference in raw material composition and production conditions. The main composition of chromium slag with calcium roasting is approximately 30% CaO, 20% MgO, 10% Fe2 O3 , 10% Al2 O3 , 10% SiO2 , 4.5% total chromium, 0.3% water-soluble Cr(VI), and 0.5% acid-soluble Cr(VI) [3]. Therefore, the chromium slag produced under the calcium roasting process has a very high alkalinity and contains a large amount of incompletely leached hexavalent chromium, which is toxic and classified as hazardous waste. The main phase composition of chromium slag is listed in Table 1.1, and the X-ray diffraction analysis pattern is shown in Fig. 1.1. Cr(VI) in chromium slag can be classified into two types: water-soluble chromium and acid-soluble chromium. The process involves finely grinding the chromium slag and subjecting it to extensive washing with hot water to extract the Cr(VI) present, which is then referred to as water-soluble chromium in the resulting washout water. The Cr(VI) obtained after the washing filter cake is dissolved by heating with dilute sulfuric acid which is called acid-soluble chromium. There is no strict boundary between water-soluble chromium and acid-soluble chromium, and the acid-soluble chromium will be slowly changed to water-soluble chromium when the chromium slag is heated with water for a long time or by the long-term effect of outdoor rain and carbon dioxide. Water-soluble Cr mainly refers to the remaining Na2 CrO4 due to incomplete leaching of roasting products during the production process. Acidsoluble Cr(VI) refers to the free CaCrO4 and the Cr(VI) presenting in the lattice of calcium silicate and calcium iron aluminate in the form of calcium chromium aluminate—calcium chromate, calcium silicate—calcium chromate, calcium iron aluminate—calcium chromate solid solution. Upon conducting an investigation into

1.1 Sources of Soil Chromium

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Table 1.1 Phase composition of chromium slag Phase

Molecular formula

Content (%)

Magnesite

MgO

~20

Dicalcium silicate

β-2CaO·SiO2

~25

Calcium ferroaluminate

4CaO·Al2 O3 ·Fe2 O3

~25

Calcium chromite

CaCr2 O4

5–10

Chrome spinel

(Mg,Fe)Cr2 O4

Calcium chromate

CaCrO4

1

Sodium chromate tetrahydrate

Na2 CrO4 ·4H2 O

2–3

Calcium chromium aluminate

4CaO·Al2 O3 ·CrO3 ·12H2 O

1–3

Alkaline ferric chromate

Fe(OH)CrO4

0.5

Calcium carbonate

CaCO3

2–3

Calcium aluminate hydrate

3CaO·Al2 O3 ·6H2 O

~1

Fig. 1.1 X-ray diffraction analysis of chromium slag

the leaching behavior of Cr(VI) from chromium slag, James et al. discovered a significant correlation between the leaching behavior and the dissolution characteristics of acid-soluble CaCrO4 in the surrounding environment. Consequently, the persistent contamination of chromium slag and the challenges associated with achieving complete detoxification can be primarily attributed to the presence of acid-soluble Cr(VI). The wastewater containing chromium in chromium salt production mainly consists of chromic anhydride production plant wastewater, ground and equipment cleaning wastewater, laboratory wastewater and boiler discharge wastewater. The wastewater produced by the chromium salt production system is very small, but the flushing water produced by the washing of equipment and site and evaporation with froth due to material and slag loss, valve leakage, etc., is large and has a high chromium concentration. Chromium-containing waste gas pollutants are mainly generated in rotary kilns, chrome slag drying, grinding, neutralization, acidification, chromic anhydride

4

1 Contamination Characteristics of Soils and Biological and Chemical …

heating furnaces and coal-fired boiler sections. The main pollutants are soot and dust, and Cr(VI) and total chromium contained in the soot (dust), while gaseous pollutants such as SO2 , nitrogen oxides, chromic acid mist, HCl and chlorine gas are also generated due to fuel combustion and evaporation.

1.1.2 Chromium Pollution in the Electroplating Industry With the advancements in mechanical manufacturing, chromium plating has been witnessed widespread adoption owing to its array of exceptional properties. As a result, the demand for chromium coatings has been steadily increasing. A large number of strong acids, strong alkalis, heavy metal solutions, and even toxic and harmful chemicals such as cyanide and chromic anhydride have been applied in the electroplating industry. These toxic and harmful substances are discharged into the environment through wastewater and waste residues, which makes chromium plating a heavily polluting industry. Only approximately 1/3 of chromic anhydride is used for chromium plating, and the other 2/3 of chromic anhydride is consumed in wastewater or waste gas during chromium plating, which has caused serious pollution to the environment. The chromium plating solution widely used in industry consists of chromic anhydride supplemented by a small number of anions. Chromic acid, derived from the aqueous solution of chromic anhydride, serves as the sole source for chromium coating. The concentration of chromic anhydride can span a broad range. When employing a high-concentration chromic acid electrolyte, there is a tendency for plating bath loss upon withdrawal of the workpiece. This leads to unnecessary material consumption and introduces certain environmental pollution. Conversely, a low-concentration plating bath is susceptible to impurity metal ions and exhibits inadequate coverage. The existing form of Cr(VI) in the plating solution varies according to the concentration of chromic anhydride, generally in the form of chromic acid (CrO4 2− ) and dichromatic acid (Cr2 O7 2− ). When the pH < 1, Cr2 O7 2− is the main form; when the pH = 2–6, Cr2 O7 2− and CrO4 2− have the balance as follows: Cr2 O7 2− + H2 O = 2H2 CrO4 − + 2CrO2 − + 2H+ . When the pH > 6, CrO4 2− is the main form in chromic anhydride. Therefore, the chromate ions in the chromium plating electrolyte are Cr2 O7 2− and CrO4 2− . In the chromium plating solution, Cr(III) is generated through the cathodic reduction of Cr(VI), while simultaneously being reoxidized at the anode. The concentration of Cr(III) quickly reaches equilibrium, and the equilibrium concentration is contingent upon the cathode-to-anode area ratio. The colloid film formed by the cathode predominantly comprises Cr(III) ions, and successful chromium deposition occurs only when the plating solution contains a specific quantity of Cr(III). Sulfate, fluoride, fluorosilicate, fluoroborate, and their combinations are frequently employed as catalysts in chromium plating. Insufficient catalyst content results in the formation of minimal or no coating, primarily resulting in a brown oxide

1.1 Sources of Soil Chromium

5

layer. Conversely, excessive catalyst levels lead to inadequate coverage, diminished current efficiency, and the possibility of partial or complete absence of the coating. At present, sulfuric acid is widely used as a catalyst. Therefore, Cr2 O7 2− , CrO4 2− , H+ and SO4 2− exist in the chromium plating electrolyte.

1.1.3 Chromium Pollution in the Tanning Industry Tanning is the most important process in the leather industry. Chrome tanning was invented in the middle of the nineteenth century. After more than 100 years of development, the modern leather industry has formed a complete set of tanning process systems based on chrome tanning. Due to its simple operation, easy control, and high moisture and heat resistance, chrome tanning is suitable for making various light leathers, such as shoe upper leather, garment leather, glove leather, furniture leather, etc. Soon, it was widely applied in the leather industry and occupied a dominant position. The utilization rate of chromium in the conventional chrome tanning method is only 65–75%, and the remaining chromium is left in the waste liquid. The concentration was up to 25 g L−1 (Cr2 O3 ). Therefore, the main cause of chromium pollution in the leather industry is the incomplete utilization of chromium and its elution in the subsequent process, which not only places great pressure on the ecological environment but also causes the waste of chromium resources. In particular, chromium retention is carried out on the basis of leather having almost saturated absorption of chromium. Hence, the chromium absorption rate is lower, and the elution rate is higher. Chrome tanning mainly causes Cr(III) pollution. According to the current treatment method for Cr(III) in the leather industry, except for the separated chrome tanning waste liquor, which is precipitated into chromium-enriched sludge by alkali, other Cr(III) basically enters the integrated sludge. The Cr(III) in tannery integrated sludge mainly exists in the form of Cr(OH)3 , which is easily converted into chromic acid and forms chromic acid mist under high-temperature and aerobic conditions. With the development of pollution control and clean tanning technology, chrome tanning technology will develop toward high absorption, waste liquid sectional throttling and recycling. The research and development of chrome-free tanning technology will receive increasing attention.

1.1.4 Chromium Pollution from Other Industries In addition to chromium salt production, the electroplating industry and industrial chrome tanning, chromium pollution also widely exists in metallurgy, building materials, medicine, anticorrosion, dyes and the use of agrochemicals. In metallurgy, chromium pollution generally originated from chrome ore smelters, whose washing

6

1 Contamination Characteristics of Soils and Biological and Chemical …

wastes contained 136 mg L−1 of total chromium, the majority of which was Cr(VI) at 112 mg L−1 and Cr(III) at 24 mg L−1 on average. During the zinc smelting operation, chromium pollution in the zinc smelting dust reached an average of 42 mg kg−1 , while the content of Cr(III) in the spray tower-emitted water and ferrochromium-silicon alloy smelter slag processing waste effluents reached 4.4 and 1964 mg L−1 , respectively. Chromium pollution in the field of building materials came from the use of high temperature resistant magnesium-chromium bricks and hexavalent chromium, a byproduct of cement production. The dissolution of chromium in cement was obviously influenced by the change in pH. Under alkaline conditions, the amount of Cr(VI) dissolution was small, but when the pH dropped to 6–8, the chromium dissolution increased dramatically, reaching 2.0–2.5 mg L−1 , which greatly exceeded the normal drinking water standard of chromium content below 0.05 mg L−1 . With the advancement of urbanization, paint sludge was also produced in large quantities, bringing chromium pollution even as high as 1500 mg kg−1 . Chromium pollution in the pharmaceutical industry mainly came from the organic synthesis reaction in drug synthesis, and it has been the core concern of heavy metal pollution control in the pharmaceutical industry that chromium pollution from sodium dichromate used in the oxidation reaction of chromic anhydride. Relatively speaking, it was convenient to recycle the chromium in a unified manner because of the relatively small amount of chromium used in anticorrosion. In textile printing and dyeing, chromium was generally applied as a dyeing aid and mordant, and the total amount of chromium in its wastewater reached 600 mg L−1 , among which contained Cr(VI) and Cr(III). The waste gas, slag and wastewater generated by the production process of chromiumrelated industries caused serious pollution threats to the area where the enterprises were located and the surrounding soil.

1.2 Characteristics of Chromium Pollution in Soil The discharge of waste containing chromium is the primary contributor to the accumulation of chromium in soils. China stands out as one of the major producers of chromate. During a period of more than 30 years, the metallurgical and chemical sectors collectively released more than 6 million tons of chromium-containing slag [4]. As most Cr manufacturing activities lack suitable disposal facilities, significant volumes of chromium-containing slag may be deposited there without adequate safeguards against leaching, which may lead to soil contamination with the heavy metal Cr. Slag that contains chromium has a high concentration of soluble Cr(VI), which seeps into groundwater from the soil. Therefore, it is crucial to assess the extent of chromium contamination in soils impacted by chromium-containing slag and to examine the transport of Cr into the subsoil, as well as the levels of total Cr and water-soluble Cr(VI) in both bulk soil and soil profiles.

1.2 Characteristics of Chromium Pollution in Soil

7

1.2.1 Chromium Concentrations in Chromium Pollution in Soil Since soil Cr(VI) contamination in China mainly resulted from chromium salt manufacturing, a site of chromium-containing slag heap was chosen for investigation of chromium pollution in soil. The sites containing chromium-containing slag, along with the surrounding area, are situated within a steel alloy factory in central Hunan Province, located in the central and southern parts of China (latitude 27°75' N, longitude 112°50' E). This region experiences a subtropical warm-moist climate, with an average annual temperature of 17 °C. Monthly mean temperatures range from 29.7 °C in July to 4.1 °C in January. The average annual precipitation in the area is 1300 mm.

1.2.1.1

The Total Cr Concentrations in Topsoils Contaminated by Chromium-Containing Slag

Table 1.2 provides an overview of the total Cr content in the topsoil across different sites. The concentration of total Cr in the soil exhibits significant variations depending on the distance from the location of the chromium slag. The soils beneath the slag heap exhibit total Cr concentrations ranging from 1248.4 to 2130.3 mg kg−1 . In the area directly surrounding the slag heap, the total Cr concentrations range from 656.1 to 3500.1 mg kg−1 , with a mean value of 2239.5 mg kg−1 . The soils adjacent to the sewage channel display a mean total Cr concentration of 995.3 mg kg−1 , with a range of 208.6 to 6207.6 mg kg−1 . In contrast, the total Cr concentrations in the control location, representing uncontaminated soil located 5 km away from the chromium-containing slag heap, exhibit minimal variation, ranging from 90.0 to 117.5 mg kg−1 . The total Cr concentrations in the soils beneath the slag heap, around the slag heap, and near the sewage channel are 15, 21, and 9 times higher, respectively, compared to the control location. These results clearly indicate the severe pollution caused by the chromium-containing slag in the investigated areas. Among the three selected locations, the soil in proximity to the slag heap displays the highest level of contamination, while the contamination level is lowest near the sewage channel. The total Cr contents in the soils beneath the slag heap are slightly lower than those in the vicinity of the slag heap. This discrepancy can be attributed to the natural leaching and mobility of the chromium-containing slag, which contributes to the majority of the total Cr in the surface soils beneath the slag heap. However, the overall enrichment of Cr in the topsoil around the slag heap can be attributed to runoff, leaching, and diffusion of Cr(VI) released by the chromium-containing slag. It is important to note that high Cr concentrations in soils can enhance plant adsorption and have adverse effects on ecosystems, as well as the health of animals and humans if ingested through food sources [5, 6]. Given that the soil samples collected along the sewage channel were paddy soils, it is reasonable to anticipate that the elevated levels of total Cr in this area could potentially contaminate the food chain. These findings underscore the

8

1 Contamination Characteristics of Soils and Biological and Chemical …

Table 1.2 Total Cr and water-soluble Cr in topsoil (mg/kg) Sampling site

N

Mean

Minimum

Maximum

SD

Total Cr Under slag heap

3

1589.3

1248.4

2130.3

473.5

In the vicinity of slag heap

5

2239.5

656.1

3500.1

1413.5

Near the sewer channel

23

995.3

208.6

6207.6

1643.1

Control site

10

104.9

90.0

117.5

13.9

Water soluble Cr(VI) Under slag heap

3

123.8

49.5

252.7

In the vicinity of slag heap

5

36.9

0.5

101.8

112 38.9

Near the sewer channel

23

0.6

0.2

1.7

0.3

Control site

10

0.7

0.1

1.1

0.5

Note N = number of samples; SD = standard deviation Reprinted from Ref. [4], Copyright 2009, with permission from the Nonferrous Metals Society of China and Elsevier

significant detrimental impact of chromium released from the ferroalloying factory on the soil ecosystem.

1.2.1.2

Water-Soluble Cr Concentrations in Topsoils Contaminated by Chromium-Containing Slag

Table 1.2 also presents the concentrations of water-soluble Cr(VI) at various sampling locations. The soils beneath the slag heap exhibited an average water-soluble Cr(VI) concentration of 123.8 mg kg−1 , ranging from 49.5 to 252.7 mg kg−1 . The soils surrounding the slag heap had an average water-soluble Cr(VI) concentration of 36.9 mg kg−1 , with a range of 0.5–101.8 mg kg−1 . The average water-soluble Cr(VI) concentrations beneath the slag heap and in the vicinity of the slag heap were 176.9 times and 52.7 times higher, respectively, compared to the unpolluted soils at the control site. In contrast, the water-soluble Cr(VI) levels in the soils near the sewage channel were similar to those observed at the control site. These results clearly indicate significant chromium contamination at the slag site of the ferroalloy plant. Furthermore, the soil beneath the slag heap exhibited the highest concentration of water-soluble Cr(VI). It was worth noting that although the water-soluble Cr(VI) concentration in the soil along the sewer channel was low, the total Cr concentration exceeded the critical limit set by the Secondary Environmental Quality Standard for Soil in China. In the presence of MnO2 , Cr(III) can be transformed into Cr(VI) [7]. As soil properties, such as pH and organic matter content, undergo changes, other forms

1.2 Characteristics of Chromium Pollution in Soil

9

of Cr, such as exchangeable Cr and carbonate-bonded Cr, can also be converted into water-soluble Cr(VI) [8]. It is important to note that even low levels of water-soluble Cr(VI) in soils near the sewer channel may have potential health implications.

1.2.2 Vertical Distributions of Total Cr and Water-Soluble Cr(VI) in the Soil Profiles 1.2.2.1

The Distributions of Total Cr in the Soil Profiles

Figures 1.2, 1.3, 1.4 and 1.5 show the vertical profile distributions of total Cr in the studied locations. Beneath the slag heap, total Cr concentrations in profiles 1 and 3 steadily rose with soil depth within 60 cm before declining again. The results indicated that total Cr accumulated at a depth of 60 cm. This most likely contributed to Cr immobilization by additional Fe and Mn molecules. Nevertheless, a greater Cr content was discovered in profile 2 at depths of 20–40 cm. Due to the limited number of profiles, there was no consistent distribution pattern of total Cr observed in the soil profiles near the slag heap (Fig. 1.2). However, in the vicinity of the sewer channel, the profiles 6, 7, 8, and 9 in Fig. 1.4 exhibited a decreasing trend of total Cr concentration with decreasing depth. In comparison to the highly contaminated soils beneath and surrounding the slag heap, the chromium pollutant originating from chromium-containing slag was predominantly immobilized in the topsoil layers (Fig. 1.3). With the exception of profile 12, which exhibited considerably higher total Cr accumulations at the surface level, there was no discernible change in total Cr concentrations in the uncontaminated soil profiles (profiles 10 and 11) (Fig. 1.5). Generally, the mean total Cr concentrations at different soil depths beneath the slag heap followed this order: 40– 60 cm (2236.8 mg kg−1 ) > 20–40 cm (2210.9 mg kg−1 ) > 0–20 cm (1589.3 mg kg−1 ) > 60–100 cm (1258.4 mg kg−1 ) > 100–150 cm (623.1 mg kg−1 ). A similar trend was observed in the soil profiles surrounding the slag heap. These findings clearly demonstrate a significant retention of Cr in the intermediate soil depths at the chromiumcontaining slag sites of the iron alloy factory (under the slag heap and in the vicinity of the slag heap). Previous studies have shown that the vertical distribution of total Cr in the soil profiles surrounding the slag heap exhibited an accumulation pattern towards the intermediate depths [9]. At intermediate depths, the observed leaching and modest accumulation of Cr can be primarily attributed to the retention of the element by iron compounds [1]. In geological environments, the solubility of Cr(III) is commonly regulated by the solid solution (Cr, Fe)(OH)3 . This solid solution, which forms in the presence of Fe(III), plays a significant role in controlling the solubility of Cr(III) since Cr exists in two oxidation states, Cr(III) and Cr(VI) [10–12]. The soil profile studied in this research was collected from the Quaternary red earth region in southern China, known for the accumulation of iron-manganese nodules in intermediate-depth soil. This phenomenon could potentially explain the observed

10

1 Contamination Characteristics of Soils and Biological and Chemical …

Fig. 1.2 The distributions of total Cr and water-soluble Cr(VI) in soil profiles under the chromiumcontaining slag heap. Reprinted from Ref. [4], Copyright 2009, with permission from the Nonferrous Metals Society of China and Elsevier

Fig. 1.3 The distributions of total Cr and water-soluble Cr(VI) in soil profiles in the vicinity of chromium-containing slag heaps. Reprinted from Ref. [4], Copyright 2009, with permission from the Nonferrous Metals Society of China and Elsevier

1.2 Characteristics of Chromium Pollution in Soil

11

Fig. 1.4 The distributions of total Cr and water-soluble Cr(VI) in soil profiles near the sewer channel of the manufactory. Reprinted from Ref. [4], Copyright 2009, with permission from the Nonferrous Metals Society of China and Elsevier

accumulation of total Cr in the intermediate-depth soils beneath the slag heap and in its surrounding areas. On the other hand, the vertical distribution profiles near the sewage channel exhibited distinct patterns of total Cr compared to those beneath the slag heap and in its vicinity. In these profiles, the highest concentrations of total Cr were found in the topsoil, followed by a rapid decrease with depth. These variations are likely influenced by changes in land use and human factors. The areas along the sewage channel were primarily used for farming activities, where the application of fertilizers and organic manure was common. This resulted in the accumulation of organic matter in the topsoil. It was possible that this organic matter played a contributing role in the substantial enrichment of total Cr in the topsoil along the sewage channel. The high concentration of organic matter also suggested the formation of metal–organic complexes, which could lead to the accumulation of metals in the soil. Therefore, the distribution of total Cr in the studied soil profile exhibited different patterns, with the intermediate-depth soils showing accumulation related to the presence of iron-manganese nodules, while the topsoil along the sewage channel exhibited higher concentrations potentially influenced by organic matter and metal– organic complex formation [10]. According to several publications, the buildup of total Cr in contaminated soils may vary depending on the amount of organic matter in the soils since chromium may be absorbed by organic matter. The study findings indicated that Cr has the ability to penetrate the soil and reach considerable depths of up to 150 cm. In fact, one sample taken at a depth of 150 cm

12

1 Contamination Characteristics of Soils and Biological and Chemical …

Fig. 1.5 The distributions of total Cr and water-soluble Cr(VI) in soil profiles at uncontaminated sites. Reprinted from Ref. [4], Copyright 2009, with permission from the Nonferrous Metals Society of China and Elsevier

beneath a slag heap exhibited the Cr levels that were three times higher than the critical limit specified by the Secondary Environmental Quality Standard for Soil in China (Fig. 1.2). Moreover, the total Cr concentrations in the profile soil beneath and near the slag heap consistently exceeded the standard limit, ranging from two to six times higher. However, along the sewage channel, the mean Cr contents below a depth of 40 cm were within the standard limit, except for samples obtained from the surface soil (0–20 cm) and subsurface soil (20–40 cm), which exceeded the standard limit. Notably, there were significant variations in mean concentrations across the three sites, with the contamination severity following the sequence of under the slag heap > vicinity of the slag heap > along the sewage channel. These findings highlight the potential for chromium migration into the subsoil from the slag heap, despite the soil’s role in retaining chromium. Consequently, the study suggests that chromium poses a substantial risk to the groundwater system.

1.2.2.2

The Distributions of Water Soluble Cr(VI) Content in Soil Profiles

The presence of water-soluble Cr(VI) in profile 1 and profile 2, located under the slag heap, exhibited an initial increase within the top 40 cm of depth (Fig. 1.2), followed by a subsequent decline. Notably, the concentration of water-soluble Cr(VI) was

1.2 Characteristics of Chromium Pollution in Soil

13

significantly higher at the surface depth (0–20 cm) compared to other levels, while profile 3 demonstrated relatively consistent levels of water-soluble Cr(VI) throughout (Fig. 1.2). Similarly, profiles 4 and 5, situated in the vicinity of the slag heap, displayed a decreasing trend in water-soluble Cr(VI) concentrations with increasing depth (Fig. 1.3). Along the sewage channel, all profiles showed extremely low levels of water-soluble Cr(VI), close to the background level (Figs. 1.4 and 1.5). Across all test sites, the average concentrations of water-soluble Cr(VI) in the soil profiles decreased as the depth increased. The soil beneath the slag heap consistently exhibited significantly higher water-soluble Cr(VI) levels compared to the vicinity of the slag heap, followed by the sewage channel and control sites. These findings indicate extensive contamination by chromium from chromium-containing slag heaps, particularly beneath and near the slag heaps. Furthermore, the persistence of elevated water-soluble Cr(VI) levels at depths of 100–150 cm suggests a substantial release and migration of this form of chromium from the slag heap. This transfer of Cr(VI) raises concerns about potential risks to the groundwater system in the area.

1.2.3 Chromium Fractions in the Contaminated Soils The concentration of various chromium fractions in soils collected from different depths at three different locations, under chromium slag, near the slag heap and unpolluted land, are presented in Table 1.3. The mean Cr values of the water soluble, exchangeable, carbonate-bonded fractions in soils under chromium slag were 97.5, 3.0 and 16.3 mg kg−1 , accounting for 6.26%, 0.19% and 1.05% of the total Cr, respectively. Near the slag heap, the Cr contents in the water-soluble, exchangeable, carbonate-bonded fractions were 16.8, 11.5 and 5.7 mg kg−1 , accounting for 1.93%, 1.32% and 0.65% of the total Cr, respectively. In the unpolluted soil, the Cr concentrations in the water soluble, exchangeable, carbonate-bonded fractions were 1.8, 1.5 and 2.5 mg kg−1 , accounting for 2.1%, 1.7% and 2.9% of the total Cr, respectively. Because the chromium slag in the current study has a high pH value, the high percentage of carbonate-bonded Cr in soils under chromium slag and near chromium slag was higher than that in unpolluted soil profiles. The vertical variation in watersoluble Cr in the soil profiles was not consistent in the different locations. The Cr amount of the water-soluble and exchangeable fractions in the soil profiles. There was no obvious difference in carbonate-bonded Cr observed in the soil profiles. Fe- and Mn-bonded Cr, organic-bonded Cr and residual Cr were generally higher than water-soluble Cr, exchangeable Cr and carbon-bonded Cr (Table 1.3). The mean values of Fe and Mn-bonded Cr in profile soils under chromium slag heap ranged from 269.4 to 1679 mg kg−1 . The equivalent values for soils near slag heaps varied from 59.8 to 439.9 mg kg−1 and from 9.9 to 20.78 mg kg−1 in unpolluted soils. Relatively high amounts of Fe- and Mn-bound Cr were found in the subsoil and 40–60 cm soil depth. The mean values of organic-bounded Cr in soils beneth chromium slag near slag heaps and unpolluted land were 273.2, 201.2 and 16.2 mg kg−1 , respectively.

14

1 Contamination Characteristics of Soils and Biological and Chemical …

Table 1.3 Chromium fractions in soil profiles (mg kg−1 ) Soil depth (cm)

Soil under the chromium-containing slag heap (n = 3)

Soil in the vicinity of the chromium-containing slag heap (n = 3)

Uncontaminated soil (n = 3)

Water soluble Cr (Mean ± SD) 0–20

123.8 ± 112.0

33.3 ± 40.5

2.5 ± 1.0

20–40

132.7 ± 145.8

24.1 ± 38.7

1.8 ± 0.5

40–60

91.2 ± 112.9

18.9 ± 30.7

1.4 ± 0.2

60–100

76.2 ± 90.8

6.8 ± 8.0

1.6 ± 0.3

100–150

63.7 ± 100.2

0.8 ± 0.6

Exchangeable Cr (Mean ± SD) 0–20

3.9 ± 3.1

18.6 ± 22.3

1.6 ± 1.1

20–40

3.0 ± 0.2

16.2 ± 23.9

1.4 ± 0.6

40–60

2.9 ± 2.3

11.9 ± 15.9

1.5 ± 0.3

60–100

3.2 ± 1.9

6.3 ± 5.8

1.4 ± 0.7

100–150

1.9 ± 0.4

4.6 ± 4.2

Carbonate-bonded Cr (Mean ± SD) 0–20

12.1 ± 8.7

7.7 ± 8.4

2.4 ± 0.1

20–40

20.7 ± 16.6

10.4 ± 13.7

2.5 ± 0.1

40–60

21.8 ± 17.7

4.1 ± 3.5

2.5 ± 0.1

60–100

16.0 ± 14.6

3.8 ± 2.9

2.6 ± 0.2

100–150

10.8 ± 7.7

2.3 ± 0.0

Organic-bonded Cr (Mean ± SD) 0–20

291.6 ± 385.8

403.6 ± 618.2

20.3 ± 12.5

20–40

390.6 ± 270.1

374.9 ± 572.7

17.7 ± 6.1

40–60

453.5 ± 215.6

83.7 ± 67.1

15.1 ± 3.6

60–100

172.2 ± 103.2

124.3 ± 125.4

11.7 ± 2.1

100–150

58.2 ± 36.2

28.9 ± 20.5

Fe and Mn-bonded Cr (Mean ± SD) 0–20

471.4 ± 434.1

301.4 ± 343.4

20.8 ± 7.9

20–40

1459.2 ± 582.9

439.9 ± 557.5

17.2 ± 9.1

40–60

1679.4 ± 622.2

137.9 ± 112.6

10.9 ± 1.5

60–100

792.3 ± 406.5

228.0 ± 267.7

9.9 ± 2.5

100–150

269.4 ± 140.2

59.8 ± 70.3

Residue Cr (Mean ± SD) 0–20

326.7 ± 161.7

683.8 ± 742.7

57.4 ± 20.2

20–40

240.7 ± 241.9

677.5 ± 770.6

44.4 ± 21.7

40–60

113.0 ± 19.1

249.9 ± 50.3

57.6 ± 13.1 (continued)

1.2 Characteristics of Chromium Pollution in Soil

15

Table 1.3 (continued) Soil depth (cm)

Soil under the chromium-containing slag heap (n = 3)

Soil in the vicinity of the chromium-containing slag heap (n = 3)

Uncontaminated soil (n = 3) 40.4 ± 16.4

60–100

208.7 ± 195.1

224.2 ± 73.4

100–150

276.6 ± 420.8

169.9 ± 135.8

Reprinted from Ref. [4], Copyright 2009, with permission from the Nonferrous Metals Society of China and Elsevier

Regarding residual Cr, its mean content in soils under chromium slag in the vicinity of slag heap and unpolluted land was 233.1, 389.7 and 49.9 mg kg−1 , respectively. The assessment of potential risks associated with heavy metals in soils requires considering their bioavailability rather than relying solely on total metal concentrations. This is because metals exhibit complex distribution patterns among different chemical fractions or solid phases, making it difficult to accurately gauge their bioavailability. The bioavailability of heavy metals is predominantly influenced by their chemical forms, particularly those related to solubility. Of particular concern are the exchangeable and carbonate forms, which are prone to leaching and can pose a threat to groundwater contamination. These forms play a crucial role in determining heavy metal uptake by plants and are thus of utmost importance in assessing environmental hazards [8]. In comparison to unpolluted land, soils under chromate slag and near slag heaps had greater levels of water-soluble Cr, exchangeable Cr, and carbonate forms. The results suggest that the solubility and mobility of chromium were high in the polluted soils, particularly in the subface soil. The analysis shows that the environmental danger posed by the metal chromium released by the Steel-Alloy factory poses a significant risk to the environment. In conclusion, the mean contents of various Cr fractions were in the order Fe and Mn-bonded Cr > residual Cr > organic-bonded Cr > water-soluble Cr > carbonate-bonded Cr > exchangeable Cr. In soil profiles contaminated by chromate slag, the concentration of water-soluble Cr generally decreased with profile depth. The subsoil and the first 40–60 cm of the soil were found to contain rather significant concentrations of o Fe- and Mn-bonded Cr and organic-bonded Cr. In the profile layers, there was no discernible difference between exchangeable Cr and carbonate-bonded Cr. Topsoil beneth chromate slag and in the vicinity of the slag heap contained relatively high levels of water soluble Cr in comparison to unpolluted soils.

16

1 Contamination Characteristics of Soils and Biological and Chemical …

1.3 Microbial Communities in Chromium-Contaminated Soil The recycling of essential nutrients for plants, the preservation of soil structure, and the detoxification of toxic substances are all significant functions carried out by microbes. Heavy metals are hazardous to soil biota, affecting essential microbial processes and reducing the quantity and activity of soil microorganisms [13]. The microbial community has often been presented as a simple and sensitive indicator of anthropogenic influences on soil ecology. Elevated levels of Cr(VI) have been found to alter the composition of soil microbial communities and adversely affect the metabolic activity of microbial cells [14]. As a result, it is critical to look at the microbial communities in hexavalent chromium-contaminated soil, as well as the genetic diversity of hexavalent chromium-resistant bacteria. Soil samples were obtained from six chromite ore processing residue (COPR) disposal sites in China (D: Hunan; G: Henan; H: Shanxi; J: Shaanxi; N: Hebei; R: Yunnan). Since the 1960s, chromite ore processing residue has been disposed of at these locations, producing serious chromium pollution in the local ecosystem. Microbial communities were discovered in chromium-contaminated soil (Table 1.4).

1.3.1 Bacterial Structural Diversity in the Contaminated Soils Soil samples D, G, H, J, N, and R showed 17, 17, 20, 12, 15, and 21 DNA bands, as shown in Fig. 1.6a. Moreover, distinct discrepancies in the locations of the DNA bands in each lane suggested that the bacterial communities in these soils were distinct. The DICE similarity coefficient was used for cluster analysis (Fig. 1.6b). The bacterial cummunities’ similaritiey coefficient in the six soils varied The from 0.52 to 0.81. Further, soils G and J had the greatest similarity coefficient (0.81). The presence of high hexavalent chromium content did not have a significant impact on the structural diversity of bacteria in the soil samples. Soil R had the highest hexavalent chromium concentration (481.27 mg kg−1 ) among the six soils, but it also exhibited the most diverse bacterial community with 21 DNA bands. Conversely, Soil N had the lowest hexavalent chromium concentration (23.60 mg kg−1 ) and showed only 15 DNA bands. Correlation analysis revealed a strong relationship between the diversity of soil bacterial communities and soil pH (correlation coefficient of − 0.874, p < 0.05), while the correlation coefficient between the diversity of soil bacterial communities and soil hexavalent chromium concentration was not significant. The results indicated that soil pH had a more significant influence on the diversity of soil bacterial communities in polluted areas compared to the concentration of hexavalent chromium. This suggests that the variation observed in the bacterial communities was primarily driven by soil pH rather than the specific levels of hexavalent chromium present.

Hunan

Henan

Shanxi

Shaanxi

Hebei

Yunnan

D

G

H

J

N

R

243.9 ± 4.2 176.5 ± 2.8 135.8 ± 2.0 23.6 ± 2.2 481.3 ± 2.5

8.7 ± 0.1

9.2 ± 0.3

8.9 ± 0.1

a

8.1 ± 0.1

57.6 ± 4.5

88.4 ± 3.4

83.2 ± 2.7

104.7 ± 3.5

99.8 ± 6.2

350.2 ± 5.4

151.8 ± 3.5

8.6 ± 0.2

9.0 ±

Pb

Cr(VI)

0.1a

pH

Mean value ± standard deviation Reprinted from Ref. [15] by permission of Taylor & Francis Ltd

Location

Soil

Table 1.4 Source and properties of different soil samples (mg kg−1 )

2.0 ± 0.4

3.0 ± 0.3

2.7 ± 0.1

5.3 ± 0.4

3.1 ± 0.6

9.5 ± 0.3

Cd

72.1 ± 2.9

44.5 ± 3.6

57.8 ± 4.0

540.0 ± 5.3

64.6 ± 3.0

101.5 ± 2.5

Cu

6.94 ± 0.5

5.19 ± 0.6

9.86 ± 0.4

11.10 ± 0.4

4.38 ± 0.6

10.16 ± 0.7

Hg

27.1 ± 4.0

10.4 ± 2.1

13.2 ± 2.7

6. 3 ± 2.0

14.4 ± 1.7

31.0 ± 2.3

As

1.3 Microbial Communities in Chromium-Contaminated Soil 17

18

1 Contamination Characteristics of Soils and Biological and Chemical …

Fig. 1.6 a DGGE banding patterns of the 16S rRNA gene extracted from soil bacterial populations. b Dendrogram constructed using the coefficient method from DGGE profiles of bacterial communities. Lanes D, G, H, J, N, and R represent soil samples contaminated with hexavalent chromium. Reprinted from Ref. [15] by permission of Taylor & Francis Ltd.

In terms of heavy metal toxicity, several mechanisms have been associated with their adverse effects. These mechanisms include direct interactions with proteins, the generation of reactive oxygen species, and the displacement of essential cations from specific binding sites. These processes contribute to the overall toxicity of heavy metals and their detrimental impact on biological systems [16].

1.3 Microbial Communities in Chromium-Contaminated Soil

19

1.3.2 Isolation and Identification of Hexavalent Chromium-Resistant Bacteria A total of twenty-two bacterial strains resistant to hexavalent chromium were isolated from sample D. Analysis of the 16S rRNA gene sequences of these isolates revealed a diverse range of bacterial strains, including Brevundimonas, Micrococcus, Exiguobacterium, Alcaligenes, and Pannonibacter genera. The nucleotide sequences of these strains were submitted to GenBank, and their corresponding accession codes and genotypic characteristics can be found in Table 1.5. Furthermore, a phylogenetic analysis of the sequences of these strains and their closest matches from GenBank is presented in Fig. 1.7. This analysis provides insights into the genetic relatedness of the isolated strains and their relationships with other known bacterial species.

1.3.3 Bacterial Genetic Diversity on the Basis of Rep-Polymerase Chain Reaction Fingerprints Figure 1.8 displays the banding patterns obtained from the BOX-, ERIC-, and REPpolymerase chain reactions. The BOX-polymerase chain reaction produced complex banding patterns with 4 to 10 DNA bands ranging from 500 to 3000 bp. The ERICpolymerase chain reaction generated two to six polymorphic bands varying from 200 to 2500 bp, while the REP-polymerase chain reaction produced two to ten polymorphic bands ranging from 200 to 3000 bp. Among the three primers, the BOXpolymerase chain reaction yielded the most intricate patterns, making it ideal for assessing the genotypic diversity of hexavalent chromium-resistant bacteria. Each primer set produced nine distinct genotypic patterns (Table 1.5), providing valuable insights into the genetic variations among the bacteria under investigation.

1.3.4 Hexavalent Chromium Reduction by Cr(VI)-Reducing Bacterial Strains In the hexavalent chromium reduction test conducted at an initial dosage of 300 mg L−1 , a total of 22 bacterial strains were evaluated. The results, presented in Table 1.5, revealed a significant reduction in the proportion of hexavalent chromium across these strains, ranging from 0.87 to 99.9%. Particularly, the isolates WY20 and WY21 were exhibited exceptional performance by effectively lowering 300 mg L−1 of hexavalent chromium within a mere 24-h period. These findings highlight the potential of these bacterial strains as highly efficient agents for hexavalent chromium remediation. WY20 and WY21 are Pannonibacter phragmitetus, according to the phylogenetic analysis (Fig. 1.9). Bacillus [17], Pseudomonas aeruginosa [18], Ochrobactrum [19] and Microbacterium [20] are among the bacterial strains with

20

1 Contamination Characteristics of Soils and Biological and Chemical …

Table 1.5 Genotypic characterization and hexavalent chromium reducing capacity of hexavalent chromium resistant isolates Isolates (access number)

Hexavalent chromium reduction (%)

REP type

BOX type

ERIC type

Closest identified strain (similarity and accession number)

WY1 (KC203066)

4.0

R1a

B1a

E1a

Brevundimonas sp. PL46b_S4 (100% and JF274919.1)

WY2 (KC203067)

40.9

R2

B2

E2

Alcaligenes faecalis strain KZJ01 (100% and FJ436432.1)

WY3 (KC203068)

5.5

R3

B3

E3

Micrococcus sp. HEXBA04 (99% and JQ658423.1)

WY4 (KC203069)

7.9

R4

B4

E4

Brevundimonas sp. SGJ (100% and HM998899.1)

WY5 (KC203070)

1.5

R5

B5

E5

Exiguobacterium sp. ERGBD-1 (99% and HM854020.1)

WY6 (KC203071)

0. 9

R5

B5

E5

Exiguobacterium sp. AFB-18 (99% and HQ848274.1)

WY7 (KC203072)

36.4

R2

B2

E2

Alcaligenes faecalis strain KZJ01 (100% and FJ436432.1)

WY8 (KC203073)

37.9

R2

B2

E2

Alcaligenes sp. C8 (100% and EU563336.1)

WY9 (KC203074)

36.6

R2

B2

E2

Alcaligenes faecalis strain KZJ01 (100% and FJ436432.1)

WY10 (KC203075)

9.0

R1

B6

E1

Brevundimonas sp. PL46b_S2 (100% and JF274917.1)

WY11 (KC203076)

39.9

R2

B2

E2

Alcaligenes faecalis strain KZJ01 (100% and FJ436432.1)

WY12 (KC203077)

35.0

R2

B2

E2

Alcaligenes faecalis strain KZJ01 (100% and FJ436432.1)

WY13 (KC203078)

39.0

R2

B2

E2

Alcaligenes faecalis strain KZJ01 (100% and FJ436432.1)

WY14 (KC203079)

11.3

R4

B6

E1

Brevundimonas diminuta strain 3P04AD (100% and EU977701.1)

WY15 (KC203080)

7.5

R1

B1

E6

Brevundimonas sp. 183 (100% and EU593764.1)

WY16 (KC203081)

2.6

R6

B7

E5

Exiguobacterium sp. AFB-18 (99% and HQ848274.1)

WY17 (KC203082)

12.1

R7

B8

E7

Brevundimonas sp. PL46b_S4 (100% and JF274919.1)

WY18 (KC203083)

40.7

R7

B8

E7

Brevundimonas sp. LSH-3 (100% and DQ825665.1)

WY19 (KC203084)

7.5

R7

B8

E7

Brevundimonas diminuta strain 764 (100% and EU430091.1)

WY20 (KC203085)

99.7

R8

B9

E8

Pannonibacter sp. W1 (100% and EU617334.1) (continued)

1.4 Adsorption of Cr(VI) on Soils

21

Table 1.5 (continued) Isolates (access number)

Hexavalent chromium reduction (%)

REP type

BOX type

ERIC type

Closest identified strain (similarity and accession number)

WY21 (KC203086)

99.9

R8

B9

E8

Pannonibacter sp. W1 (100% and EU617334.1)

WY22 (KC203087)

44.3

R9

B1

E9

Micrococcus sp. M10 (100% and JN596114.1)

a

R, B and E refer to the abbreviation of fingerprint patterns of REP-, BOX- and ERIC-PCR, respectively. The initial Cr(VI) concentration and pH were 300 mg. L−1 and 9.0, these bacterial strains were incubated at 30 °C for 24 h Reprinted from Ref. [15], Copyright © 2015 managed by Taylor & Francis

hexavalent chromium-reducing activity that have been described in the previous literatures. At concentrations ranging from 10 to 2000 mg L−1 , these bacterial strains may totally eliminate hexavalent chromium [18–21]. The 22 isolates used in this investigation expanded the number of bacteria know to be capable of reducing hexavalent chromium.

1.4 Adsorption of Cr(VI) on Soils The adsorption behavior of Cr(VI) on soils is an essential property for forecasting its current condition and giving significant information for its migration and management. Traditionally, most chromium-soil adsorption experiments have focused on the behavior of pure Cr(VI) solutions in uncontaminated soils. However, it is crucial to gain insights into the kinetics and thermodynamics of Cr(VI) adsorption in soils. This understanding is essential for comprehensively assessing the adsorption processes and their environmental implications.

1.4.1 Adsorption Kinetics of Cr(VI) Figure 1.10 depicts the quantites of Cr(VI) adsorption as a function of sorption time The Cr(VI) concentration in the soil-free blank solutionsdid not change during10 h. The tendencies of Cr(VI) adsorption over time in soil samples were constant at the three distinct temperatures (15, 25, 35 °C). The adsorption capacity of soil for Cr(VI) rose gradually with increasing soil temperature, as illustrated in Fig. 1.10, and eventually reached the equilibrium state of adsorption. Particularlly, the adsorption amounts of Cr(VI) increase fast during the first 2.5 h. After that, the adsorption quantities gradually rise until they reach equilibrium after 5 h.

22

1 Contamination Characteristics of Soils and Biological and Chemical …

Fig. 1.7 Phylogenetic tree generated by 16S rRNA gene sequences for 22 hexavalent chromiumresistant bacterial strains and similar strains. Accession numbers of the bacterial isolates in this study and others are shown in brackets. Reprinted from Ref. [15] by permission of Taylor & Francis Ltd

1.4 Adsorption of Cr(VI) on Soils

23

Fig. 1.8 Rep-PCR genomic fingerprints of 22 isolates generated with the BOX, ERIC, REP primer set. Lanes 1–22 represent isolates WY1 to WY22, respectively. Lane M is the DNA marker. Lane B is a negative control (PCR without DNA template). Reprinted from Ref. [15], Copyright © 2015 managed by Taylor & Francis

24

1 Contamination Characteristics of Soils and Biological and Chemical …

Fig. 1.9 Phylogenetic tree showing the genetic diversity of 22 hexavalent chromium-resistant bacteria based on the combined fingerprint patterns of BOX-, ERIC- and REP-PCR. Reprinted from Ref. [15] Copyright © 2015 managed by Taylor & Francis

Fig. 1.10 The effects of agitation time and temperature on the adsorption of Cr(VI) by soil. Reprinted from Ref. [22], Copyright 2019, Springer-Verlag GmbH Germany, part of Springer Nature

1.4 Adsorption of Cr(VI) on Soils

25

Fig. 1.11 Pseudo-first-order kinetics (a) and pseudo-second-order kinetics (b) for the adsorption of Cr(VI) onto soil at different temperatures. Reprinted from Ref. [22], Copyright 2019, SpringerVerlag GmbH Germany, part of Springer Nature

The kinetic data collected were analyzed by the pseudo-first-order kinetics model and pseudo-second-order kinetics model to research the mechanism of Cr(VI) adsorption on soil and assess the rate-controlling stage of the adsorption process [23]. Figure 1.11 shows the pseudo-first-order and pseudo-second-order kinetics of Cr(VI) adsorption onto soil at varies temperatures are shown in. In comparision to pseudofirst-order kinetics (Fig. 1.11a), the pseudo-second-order kinetics (Fig. 1.11b) model gives a good match to the adsorption kinetic data of Cr(VI) at different temperatures (15, 25, and 35 °C),with correlation coefficients R2 of 0.9998, 0.9999, and 0.9997, respectively. The equilibrium adsorption capacities (qe ) are 232.56, 286.53, and 312.20 mg kg−1 , respectively, andare extremely similar to the actual amount of equilibrium adsorption (229.62 mg kg−1 , 281.67 mg kg−1 , and 309.21 mg kg−1 ), respectively. The pseudo-first-order mode’s R2 values for Cr(VI) adsorption are 0.8593, 0.9548, and 0.9043, respectively, and the equilibrium adsorption capacities (qe ) are 39.06, 66.98, and 78.20 mg kg−1 (Table 1.6). The estimated adsorption capacity of Cr(VI) is significantly lower than the actual equilibrium adsorption quantity. These findings indicate that the pseudo-second-order kinetic model provides a more suitable description of the Cr(VI) adsorption kinetics. Moreover, the pseudo-secondorder model assumes that the rate-limiting step involves chemical sorption, which entails valence force interactions through electron sharing or exchange between the adsorbent and the adsorbate [24]. Hence, the good fitting of the pseudo-second-order model implies that the Cr(VI) sorption onto soil may undergo a chemical adsorption process.

1.4.2 Adsorption Thermodynamics of Cr(VI) The calculation of thermodynamic parameters, including the change in standard free energy (ΔG°), enthalpy (ΔH°), and entropy (ΔS°), were performed using the

26

1 Contamination Characteristics of Soils and Biological and Chemical …

Table 1.6 Predicted kinetic parameters for Cr(VI) adsorption on soil at different temperatures T °C

Pseudo-first-order kinetic

Pseudo-second-order kinetic

qe (mg/kg)

k 1 (min−1 )

R2

qe (mg/kg)

k 2 (kg mg−1 min−1 )

R2

15

39.06

0.0090

0.8593

232.56

0.0006

0.9998

25

66.98

0.0095

0.9548

286.53

0.0004

0.9999

35

78.20

0.0074

0.9043

312.20

0.0003

0.9997

Reprinted from Ref. [22], Copyright 2019, Springer-Verlag GmbH Germany, part of Springer Nature

following equations [25]: ln K c =

ΔH ◦ ΔS ◦ − R RT

(1.1)

ΔG ads = ΔHads − T ΔSads

(1.2)

The thermodynamic parameters (ΔG°) for the adsorption process of Cr(VI) onto the soil were determined to be −2.78, −3.36, and −3.94 kJ mol−1 at the temperatures of 15 °C, 25 °C, and 35 °C, respectively, as presented in Table 1.7. The negative values of ΔG° indicate that the adsorption of Cr(VI) onto the soil occurs spontaneously, suggesting a favorable adsorption process. Additionally, the positive value of ΔH° confirms that the adsorption of Cr(VI) is an endothermic process, indicating that heat is absorbed during the adsorption reaction [26]. The positive value of ΔS° indicates that adsorption is feasible, and enhanced unpredictability at the sorbent/solution interface occurs in the internal structure of Cr(VI) adsorption onto soil. A positive result for value of ΔH° (14.01 kJ mol−1 at 15 °C) implies an endothermic process of adsorption [27]. Table 1.8 shows the values of the Langmuir and Freundlich parameters, as well as and R2 . Both Langmuir and Freundlich isotherms had a strong correlation with the observed data (Fig. 1.12). Yet, the regression coefficient R2 revelaed that the Langmuir isotherm model was much superior to the Freundlich isotherm model in characterizing adsorption processes (R2 of Langmuir isotherm 0.975, 0.990 and 0.977 vs Freundlich isotherm 0.845, 0.882 and 0.980). The Langmuir isotherm is a typical adsorption isotherm for determining an adsorbent’s potential application. The adsorption of Cr(VI) onto soil generally follows Langmuir isotherms with the Table 1.7 Thermodynamic parameters for the adsorption of Cr(VI) by soil Temperature (K)

ΔG° (kJ mol−1 )

ΔH° (kJ mol−1 )

ΔS° (kJ mol−1 K−1 )

288

−2.78

14.01

58.33

298

−3.36

308

−3.94

Reprinted from Ref. [22], Copyright 2019, Springer-Verlag GmbH Germany, part of Springer Nature

1.4 Adsorption of Cr(VI) on Soils

27

maximum adsorption capacity (qm ) of the soil being 175, 200, and 310 mg kg−1 at temperatures of 15 °C, 25 °C, and 35 °C, respectively. The values of b, a measure of heat of adsorption, are 0.043, 0.087, and 0.038. The dimensionless separation factor (RL ) was developed baseeedon additional investigation of the Langmuir equation to predict the breakthrough behavior of Cr(VI) adsorbed on the soil bed [28]. RL =

1 (1 + bC0 )

(1.3)

where C 0 represents the initial Cr(VI) concentration (mg/L) and b represents the Langmuir isotherm constant. Four idealized types of equilibrium behavior may be identified using the value of RL . Unfavorable adsorption equilibrium is represented by RL > 1; linear adsorption equilibrium is represented by RL = 1; favorable adsorption equilibrium is represented by RL = 1; and irreversible adsorption equilibrium is represented by RL = 0. The value of RL in the current investigation varied between 0.054–0.535 (RL 1), confirming soil’s preferential adsorption of Cr (VI). Table 1.8 Parameters of Langmuir and Freundlich isotherms at different temperatures T

Langmuir parameters

Freundlich parameters

b

qmax

R2

Kd

1/n

R2

15°C

0.043

175.44

0.975

14.60

0.513

0.845

25°C

0.087

200.40

0.990

33.13

0.384

0.882

35°C

0.038

310.56

0.977

39.39

0.391

0.980

Reprinted from Ref. [22], Copyright 2019, Springer-Verlag GmbH Germany, part of Springer Nature

Fig. 1.12 Langmuir isotherm (a) and Freundlich isotherm (b) for Cr(VI) adsorption in soil at different temperatures. Reprinted from Ref. [22], Copyright 2019, Springer-Verlag GmbH Germany, part of Springer Nature

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1 Contamination Characteristics of Soils and Biological and Chemical …

1.5 Cr(VI) Migration in Chromite Ore Processing Residue (COPR)-Soil The hexavalent chromium in the chromium-containing slag may be separated into water-soluble and acid-soluble hexavalent chromium based on its solubility. The water-soluble hexavalent chromium is made up of Na2 CrO4 from the leaching and baking processes. The acid-soluble hexavalent chromium mostly consistes of the dissociated CaCrO4 , Ca2 SiO4 and 4CaO·Al2 O3 ·Fe2 O3 . While CaCrO4 is dissociated from chromium-containing slag, the majority of chromium components are dispersed within the particles rather than only as surface contaminants [29]. Ca2 SiO4 and 4CaO·Al2 O3 ·Fe2 O3 occur in the crystal lattice, for example. Because of the aforesaid features of Cr(VI) components, Cr(VI) is persistently leached from chromium-containing slag. The solubility leaching behavior of acid-soluble CaCrO4 in the outside environment was greatly reliant on Cr(VI) leaching behaviour [30]. In aqueous or acidic media, Na2 CrO4 ·4H2 O, CaCrO4 , 4CaO·Al2 O3 ·CrO3 ·12H2 O, Fe(OH)CrO4 , and adsorbed Cr(VI) have rapid diffusion rates, while CaCrO4 ·Ca2 SiO4 ·CaCrO4 and 4CaO·Al2 O3 ·Fe2 O3 ·CaCrO4 have slower diffusion rates [3]. The rate of Cr(VI) diffusion from slag direct impact on the soil contamination process.

1.5.1 Diffusion of Cr(VI) in Chromium-Containing Slag 1.5.1.1

The Deposition Potential of Cr(VI) Reduction

Samples of chromium-containing slag were obtained from the former Changsha Chromate Factory for the purpose of estimating the apparent diffusion coefficient of Cr(VI) in the slag. Two distinct methods were employed for this estimation. In the first procedure, the deposition potential of Cr(VI) reduction was assessed through linear scan voltammetry. Subsequently, the changes in current resulting from Cr(VI) reduction were monitored using single potential step chronoamperometry, which allowed the generation of a current-time course curve. By applying Fick’s laws to this curve, the apparent diffusion coefficients of Cr(VI) originating from the initial chromium-containing slag were calculated. These experimental approaches provide valuable insights into the diffusion behavior of Cr(VI) within the slag samples. The linear scan voltammetry curve in 0.1 mol L−1 Na2 SO4 baseline solution and 0.1 mol L−1 Na2 SO4 + 0.02 mol L−1 K2 Cr2 O7 solution at 0.05 V s−1 across the potential range of 0 to −1.0 V is shown in Fig. 1.13. At −0.75 V, there was a major shift in the electric current in the basal solution(Fig. 1.13a). Since protons are engaged in redox reactions in the basal solution, this inflexion potential might be termed the hydrogen deposition potential. Nevertheless, in 0.1 mol L−1 Na2 SO4 and 0.02 mol L−1 K2 Cr2 O7 solution, a quickly increasingcurrent was recorded at around −0.5 V, suggesting the startof Cr(VI) reduction on the lead electrode surface (Fig. 1.13b).

1.5 Cr(VI) Migration in Chromite Ore Processing Residue (COPR)-Soil

29

Fig. 1.13 Curve of linear scan voltammetry. Reprinted from Ref. [31], Copyright 2009, with permission from Elsevier

Following the initial stage, the current exhibited a gradual increase, accompanied by multiple waves. These findings indicated that the deposition process occurred in distinct steps. The calculated potential for Cr(VI) reduction was determined to be −0.5 V (vs. SCE). In comparison, the calculated potential for Cr(VI) reduction on the standard hydrogen electrode (SHE) was −0.26 V, which closely approximated its standard potential of −0.13 V. The slight disparity between the standard and calculated potentials resulted in an overpotential. This outcome confirmed that the first peak observed in the linear scan voltammetry curve corresponded to the deposition potential of Cr(VI) reduction.

1.5.1.2

Apparent Diffusion Coefficient of Cr(VI)

To determine the diffusion coefficient of metal ions, electrochemical methods have been frequently employed [32, 33]. It was supposed that the diffusion of Cr(VI) is an immediate and unstable diffusion process. As a result, Fick’s second law may be used to charaterize Cr(VI) diffusion. Cr(VI) diffusion in slag occurs inside the fluid-filled pores and empty spaces. The presence of other ions in the pore fluid influences Cr(VI) diffusion. Meanwhile, the chemical and physical binding force must be dispelled by Cr(VI) diffusion in slag. Slag is the most essential factor in Cr(VI) binding. As a result, the diffusion coefficient of Cr(VI) became known as the apparent diffusion coefficient. The apparent diffusion coefficient of Cr(VI) from the original and detoxified chromium-containing slag to the boundary surface was investigated using single

30

1 Contamination Characteristics of Soils and Biological and Chemical …

potential step chronoamperometry in this study (Fig. 1.14). According to Fig. 1.15, The deposition potential of Cr(VI) reduction was −0.5 V. As a result, the step potential was kept under control at −0.5 in the zone of the limit current density. For the original chromium-containing slag and detoxified, the variations in current were measured whthin 5000 and 2500 s, respectively, (Fig. 1.16). A linear line was drawn between ln (it 1/2 ) and t −1 (data not shown). In this equation, i defines the current density, and t defines time. The apparent diffusion coefficient may be stated using the slope of the above straight line and the following equation derived from Fick’s first law and Fick’s second law [34]: D

L2 α 4tg

(1.4)

where D is the apparent diffusion coefficient of Cr(VI), L is the thickness of the slag plate (5 mm) and tgα is the slope of the straight line between ln (it1/2 ) and t−1 . The apparent diffusion coefficient of Cr(VI) distributed from the original chromium-containing slag was 4.40 × 10–9 m2 s−1 , as shown in Table 1.9, whereas the comparable value for the detoxified slag treated with Achromobacter sp. CH-1 for 16 h was 2.62 × 10–8 m2 s−1 . According to the results, there was a six-times increase in Cr(VI) apparent diffusion conffcient in the detoxified slag compared the original slag. The increase in porosity and hydraulic permeability may be to blame for this Cr(VI) increase. The original slag and the detoxified slag had porosities that were 63.4% and 64.1%, respectively. The original slag’s hydraulic permeability coefficient was 8.5 × 10–5 cm s−1 ; the detoxified slag’s was 9.8 × 10–4 cm s−1 . Due to the greater apparent diffusion coefficients of Cr(VI) in the detoxified slag, bacterial were able to queckly submerge, which caused chromium-containing materials to dissolve and the slag to be destroyed. As a consequence of the recurrent destruction, additional pores formed. Cr(VI) is generally related with chromium-containing compounds in the crystal lattices, such as CaO·Al2 O3 ·CrO3 ·12H2 O, CaCrO4 ·Ca2 SiO4 ·CaCrO4 and CaO·Al2 O3 ·Fe2 O3 ·CaCrO4 . Cr(VI) was converted to Cr(III) as detoxification progressed, and the disruption of the crystal lattices might also contribute to the fast diffusion of Cr(VI) in slag.

1.5.1.3

Cr(VI) Diffusion in Chromium-Containing Slag as Affected by Cr(VI) Reduction Strain

A mixture was prepared by combining 2 g of the aforementioned original chromiumcontaining slag with a total volume of 20 mL of bacterial cultures that had been grown overnight. The mixture was then horizontally shaken at a speed of 150 rpm and a temperature of 30 °C until no traces of Cr(VI) were detected. Microbial detoxification increased Cr(VI) diffusion in chromium-containing slag in the ongoing study. While the result would be negative in temrs of environmental protection, it may be used

1.5 Cr(VI) Migration in Chromite Ore Processing Residue (COPR)-Soil CE WE RE

7

31

Electrochemical workstation

8 3

9

5

2

1 Computer

1 6

4

Chromium-containing slag plate

lead sheet electrode

polymethyl methacrylate clapboard with a square hole of 20×20×4 mm (1. Water bath; 2. electrobath; 3. counter electrode (CE); 4. slag plate; 5. clapboard; 6. lead flake (working electrode (WE)); 7. salt bridge; 8. reference electrode (RE))

Fig. 1.14 Set-up of potential step chronoamperometry. Reprinted from Ref. [31], Copyright 2009, with permission from Elsevier

to effectively and quickly detoxify chromium-containing slag. The decreased watersoluble Cr(VI) content in the leachate after the detoxification operation of chromiumcontaining slag by Achromobacter sp. CH-1 revealed that Cr(VI) was efficiently reduced rather than persisting in the ambient environment. As a result, it is possible to infer that the detoxification of chromium-containing slag increased rather than decreased Cr(VI) diffusion. Meanwhile, Achromobacter sp. CH-1 has the potential for use in Cr(VI) biodetoxification in chromium-containing slag.

32

1 Contamination Characteristics of Soils and Biological and Chemical …

Fig. 1.15 The pattern of single potential step chronoamperometry of the original chromiumcontaining slag. Reprinted from Ref. [31], Copyright 2009, with permission from Elsevier

Fig. 1.16 The pattern of single potential step chronoamperometry of detoxified chromiumcontaining slag. Reprinted from Ref. [31], Copyright 2009, with permission from Elsevier

1.5 Cr(VI) Migration in Chromite Ore Processing Residue (COPR)-Soil

33

Table 1.9 Diffusion coefficients of Cr(VI) released from chromium-containing slag Slope (tgαa )

Treatment

−1419.5

Raw slag Detoxified slag a

−238.1

Diffusion coefficient correlation coefficient

p value

(m2 s−1 )

(rb )

4.40 ×

10–9

0.9953

D4), which cannot be merely explained by the intuitive phenomena.

2.5.1.2

Comparative Genomic Analyses Revealing the Unique Genotype of BB

We examined and contrasted the shared and distinctive genes between BB, D1, and D4. 4183 common genes in total were found, the majority of which are essential for bacterial development (Fig. 2.26). A total of 454 dispensable genes were found in BB, 452 in D1, and 469 in D4, correspondingly (Fig. 2.26). BB and D1 also contain two genes. The clustering association between the three Pannonibacter phragmitetus strains is shown by the heatmap of dispensable genes (Fig. 2.27), demonstrating the close affinity of BB and D4 at higher homology. As a result, there are negligible differences between these three bacterial types’ genome sizes and gene species. Additionally, the comparison of the genes implicated in Cr(VI) resistance and reduction shows that the majority of the genes for Cr(VI) uptake, intracellular/extracellular reduction of Cr(VI), ROS detoxification/scavenging, DNA repair, and efflux of Cr are present in all three strains (VI) (Fig. 2.28) [28]. However, there are substantial variations in the distribution of the functional genes among the three strains, which may be a result of evolutionary changes brought on by various initial growth environments (Fig. 2.28a and Table 2.3). There are several Cr(VI) resistance genes in the BB genome (Fig. 2.28). There are more chrA alleles in the BB. Due to horizontal gene transfer, four of those genes are dispersed on plasmid 2, which is crucial for chromate efflux (Fig. 2.28b). The particular ROS-detoxifying genes sodM-like and ggt, which efficiently reduce oxidative stress, are also present in plasmids 1 and 2 (Fig. 2.28b). Plasmid 2 also contained a DNA or RNA helicase belonging to the Superfamily II that is involved in DNA repair (Fig. 2.28b). The BB genome also contained a number of novel genes, including ccoO and ctaG, which encode electron transport proteins that support the extracellular reduction of Cr(VI). The fact that BB can live in an environment with high Cr(VI) is explained by the discovery of particular genes and the abundance of specific genes. New mechanisms of Cr(VI) resistance and reduction by BB must be clarified in order to obtain a fuller understanding. Therefore, in addition to other techniques, proteomic and metabolomic studies were carried out.

2.5.1.3

Multienzyme System Involved in Cr(VI) Resistance and Reduction

It is impossible to separate BB’s extremely high Cr(VI) resistance and reduction from the different proteins involved in the corresponding metabolic processes (Fig. 2.29). The tetrahedral chromate ion (CrO4 2− ) has a shape resembling that of the sulfate ion (SO4 2− ). Its passing through cell membranes is facilitated by nonspecific anionic

Water treatment agents

Activated sludge 500

Pannonibacter phragmitetus D1

Pannonibacter phragmitetus D4

25

50

2500

Cr(VI) reduction (mg L−1 )

Reprinted from Ref. [31] Copyright 2019, with permission from Elsevier

500

4000

Chromate slag

Pannonibacter phragmitetus BB

Cr(VI) resistance (mg L−1 )

Source of isolated

Strains

Draft genome

Draft genome

Complete genome

Sequencing type

5,586,929

17,217

Plasmid 2 Chromosome

253,070

Plasmid 1

116,285

Plasmid 2

5,175,596

349,969

Plasmid 1 Chromosome

5,064,402

Size (bp)

Chromosome

Genome

Table 2.3 The maximum Cr(VI) resistance/reduction and genomic characteristics of three Pannonibacter phragmitetus strains

64.15

63.55

64.23

GC (%)

5251

5114

5166

Predicted genes

68 2 Mechanisms of Cr(VI) Reduction by Microorganisms

2.5 Molecular Mechanisms of Bacterial Reduction of Cr(VI)

69

Fig. 2.26 Venn maps of genomic Pangene sets between D4, D1, and BB strains. Each ellipse represents a strain; the data on each region represents the number of clusters that only appear in that one; a cluster represents a group of genes with greater than 50% similarity and sequence length differences less than 0.3. Reprinted from Ref. [31] Copyright 2019, with permission from Elsevier

(SO4 2− and PO4 3− ) carriers [32, 33]. Sulfate transporter genes were found in BB’s genome at eight different locations. As a result, BB controls this system negatively to protect it from Cr(VI)’s fatal consequences. It’s possible that sulfate transporters are downregulated to lessen the uptake of Cr(VI) by BB because the enzymatic activity of sulfate transporters in BB dropped slightly from 16.48 to 14.46 U/L (Fig. 2.30). Unexpectedly, the proteomic analysis revealed that the expression of five sulfate transporters was substantially increased (Table 2.4). The five corresponding genes were substantially upregulated according to the quantitative real-time PCR analysis, supporting this phenomenon (data not shown). The kit we used in this research might not be able to identify the enzymatic activity of various sulfate transporters in BB. Therefore, in contrast to other bacteria, BB adopts a novel, diametrically opposed, and superior approach to react to Cr(VI) stress [34]. TEM analysis demonstrated that sulfate transporters allow extracellular Cr(VI) to penetrate cells directly (Fig. 2.31). It may be possible to activate several enzymes needed for intracellular Cr(VI) reduction to convert Cr(VI) to Cr (III). The enhanced uptake of Cr(VI) for additional reduction is facilitated by the induced enzymes (Fig. 2.29). Thus, some of the Cr(VI) tension that BB was experiencing is reduced.

70

2 Mechanisms of Cr(VI) Reduction by Microorganisms

Fig. 2.27 Heatmap of dispensable genes to show the clustering relationship of three Pannonibacter phragmitetus strains. The bottom corresponds to different strains. The left is the dispensable gene cluster tree, and the above is a cluster of bacteria. The middle is the gene similarity heatmap. Different coverages are in different colors. The correspondence between color and coverage is shown in the legend above left. Reprinted from Ref. [31] Copyright 2019, with permission from Elsevier

The intracellular reductase activity of BB was examined to support our theory. Compared to the control sample, the enzyme activity of 194.65 U L−1 in 500 mg L−1 Cr(VI) rose by 9.41% (Fig. 2.30). According to proteomic analysis, Cr(VI) enhances the reducing power of BB cells and causes the overexpression of numerous reductases. Significantly overexpressed enzymes included ferredoxin (Fdx), NADPHdependent FMN reductase (SsuE), flavin reductase (RutF), NADH quinone oxidoreductase (NuoA), and nitrate reductase (NasA) (Table 2.5). They reduce Cr(VI) in bacteria similarly to other traditional chromate reductases, such as YieF, NemA, and NfsA in Escherichia coli and ferric reductase B in Paracoccus denitrificans, but they are not obligatory chromate reductases [28, 35]. The amplification of these five reductases—Fdx, SsuE, RutF, NuoA, and NasA— indicates that they are essential for intracellular Cr(VI) reduction. Other oxidoreductases with Cr(VI)-reducing activity, such as azoreductase AzoR, NemA, and coppercontaining nitrite reductase NitR, had stable or modestly elevated expression levels (Table 2.5). It’s interesting to note that BB has stable expression of the chromate reductases ChrR and nitroreductase (Table 2.5) [30, 36, 37]. These two enzymes are

2.5 Molecular Mechanisms of Bacterial Reduction of Cr(VI)

71

Fig. 2.28 Comparative genomic analysis of three Pannonibacter phragmitetus strains BB, D1, and D4 based on known mechanisms of bacterial Cr(VI) resistance and reduction. a Genes involved in Cr(VI) resistance and reduction distributed on the chromosomes of three Pannonibacter phragmitetus strains, b unique genes of Cr(VI) resistance distributed on the plasmids of BB. Reprinted from Ref. [31] Copyright 2019, with permission from Elsevier

inherent rather than Cr(VI)-inducible, but they may still be able to reduce intracellular Cr (VI). The nitR gene, which codes for nitrite reductase, was also noticeably overexpressed (Table 2.5 and Fig. 2.28). During intracellular Cr(VI) reduction, the initial redox equilibrium in BB is disrupted. ROS are the cause of oxidative stress. The primary sources of ROS in cells are O2 and OH radicals, which are created by a Fenton-like reaction and a short-lived but highly reactive intermediate called Cr(V) [28]. The well-known oxidative stressreduction enzymes glutathione S-transferase (GST), superoxide dismutase (SOD), and catalase (CAT) are regarded in this research as an additional mechanism of chromate resistance (Fig. 2.29). At least 17.39% more of these three enzymes were active, and CAT activity rose by up to 76.58% (Fig. 2.30).

72

2 Mechanisms of Cr(VI) Reduction by Microorganisms

Fig. 2.29 Molecular response mechanism of Pannonibacter phragmitetus BB to Cr(VI). (1) Multienzyme system involved in Cr(VI) resistance/reduction including Cr(VI) uptake, intracellular reduction, ROS detoxification, DNA repair, and Cr(VI) efflux, (2) LuxI/LuxR and TraI/TraR quorum sensing systems, (3) double systems for electronic transfer including membrane bound reductases and extracellular electron shuttles, (4) flagella-based negative chemotaxis to Cr(VI), and (5) reduction of Cr(VI) by metabolites such as L-cysteine and glutathione. Reprinted from Ref. [31] Copyright 2019, with permission from Elsevier

Fig. 2.30 Enzymatic activity of typical enzymes involved in Cr(VI) resistance and reduction from BB cultured with (treat) and without (control) 500 mg L−1 of Cr(VI). Reprinted from Ref. [31] Copyright 2019, with permission from Elsevier

BBGL003497 BBGL002032

Iron(III) transport system permease protein

ABC transporter substrate-binding protein, iron complex transport system substrate-binding protein

ABC-type Fe3+ /spermidine/putrescine transport systems, ATPase components

ABC transporter substrate-binding protein, ABC-type Fe3+ -hydroxamate transport system

afuB/ fbpB

–

–

–

BBGL000627

BBGL004768

ABC-type

BBGL000973

BBGL005002

BBGL004224

BBGL000192

BBGL002899 BBGL000451

BBGL000449

BBGL002898

BBGL002900 BBGL000450

BBGL002901

D1GL000635

D1GL001967

D1GL003583

D1GL004964

D1GL000888

–

D1GL004212

D1GL000236

D1GL000506 D1GL002892

–

D1GL002891

D1GL002893 D1GL000505

D1GL002894 D1GL000504

D4GL000455

D4GL001890

D4GL003371

D4GL004991

D4GL000821

D4GL004901

D4GL004066

D4GL000076

D4GL000275 D4GL002741

D4GL000273

D4GL002740

D4GL002742 D4GL000274

D4GL002743

1.25

1.30

1.31

1.38

1.39

1.45

1.96

2.30

1.32

1.36

1.39

1.87

1.96

(continued)

Gene_ID in BB Gene_ID in D1 Gene_ID in D4 Upregulation of proteins in BB

afuA/ fbpA

transport system

Iron complex transport system permease protein

Fe3+

Iron ABC transporter substrate-binding protein

–

Sulfate ABC transporter permease subunit

cysW

–

Sulfate transport system substrate-binding protein

cysP

ABC transporter substrate-binding protein, ABC-type Fe3+ -hydroxamate transport system

Sulfate transport system ATP-binding protein

cysA

–

Sulfate ABC transporter permease subunit

cysT

Other transporters

Sulfate transport system substrate

cysP

Sulfate transporters

Predicted function

Gene name

Categories

Table 2.4 Predicted genes that encode proteins related to Cr(VI) uptake in Pannonibacter phragmitetus BB, D1, and D4 and the upregulation of proteins in BB

2.5 Molecular Mechanisms of Bacterial Reduction of Cr(VI) 73

Predicted function

Iron transporter, putative hemin transport protein

Iron ABC transporter substrate-binding protein

Gene name

–

afuA/ fbpA

Reprinted from Ref. [31] Copyright 2019, with permission from Elsevier

Categories

Table 2.4 (continued)

BBGL003485

BBGL002455 D1GL003543

D1GL002391 D4GL003332

D4GL002303 1.20

1.20

Gene_ID in BB Gene_ID in D1 Gene_ID in D4 Upregulation of proteins in BB

74 2 Mechanisms of Cr(VI) Reduction by Microorganisms

2.5 Molecular Mechanisms of Bacterial Reduction of Cr(VI)

75

Fig. 2.31 TEM images of BB grown (a) without and (b) with 500 mg/L Cr(VI). Reprinted from Ref. [31] Copyright 2019, with permission from Elsevier

SOD’s presence causes several O2 ’ radicals to dismutate into H2 O2 , partly scavenging ROS brought on by O2 ’ radicals. On the basis of the proteomics analysis, the expression of all SODs was raised (Table 2.6). O2 radical levels are constrained in vivo by superoxide reductase (SOR) and SOD [37]. SOR was not found in BB, though. As a result, SOD is primarily in charge of eliminating O2 radicals. CuZn-SOD (SOD1), found in the cytosol, nucleus, and lysosomes; Mn-SOD (SOD2), found in the mitochondrial matrix; and EC-SOD (SOD3), attached to the matrix and proteoglycans in the extracellular space, are all members of the SOD family [38]. Only SOD1 and SOD2, which neutralize intracellular ROS produced by the intracellular reduction of a significant quantity of Cr, were found in BB (VI). In the meantime, H2 O2 interacts with Cr(V) to produce hydroxide (OH− ) and OH radicals through a Fenton-like reaction, which results in the production of additional ROS. To safeguard the cell from oxidative damage by ROS, CAT stops the Fenton-like reaction from occurring through the breakdown of H2 O2 to H2 O and O2 [39]. Catalase KatE and catalaseperoxidase KatG are just two examples of the CATs that are consistently expressed in BB. They successfully prevent a Fenton-like reaction from occurring (Table 2.6). GST is a potent detoxifying protein as well. The conjugation of reduced glutathione (GSH) with xenobiotic compounds like Cr(VI), Cr(III), O2 , and OH radicals in BB is known to be catalyzed by it. GSH blocks the interaction of xenobiotic compounds with essential cellular proteins and nucleic acids [40]. The marked rise in GSH suggests that GSH is important in scavenging ROS. In the BB genome, glutathione metabolism is controlled by a total of 30 genes; nine coded proteins, particularly sulfurtransferase, are markedly upregulated. Glutathione transferases prevent ROS harm by acting as Cr(VI)-inducing enzymes (Table 2.6). Additionally, metabolomic research revealed that GSTs and GSH are essential for BB’s ability to detoxify ROS. L-glutamic acid and norvaline, two intermediates in the production of glutathione, are increased in the middle and late phases of Cr(VI) reduction (Fig. 2.32). DNA changes were brought about by the intracellular reaction to entering Cr(VI)produced ROS, active intermediate Cr(V), Cr(IV), and end product Cr(III) [28, 41,

76

2 Mechanisms of Cr(VI) Reduction by Microorganisms

Table 2.5 Predicted genes that encode intracellular reductases in Pannonibacter phragmitetus BB, D1, and D4 and the upregulation of those reductases in BB Gene name

Predicted function

Gene_ID in BB

Gene_ID in D1

Gene_ID in D4

Upregulation of proteins in BB

fdx

Reductase, 2Fe-2S ferredoxin

BBGL005040 –

ssuE

NADPH-dependent FMN reductase

BBGL002909 D1GL002901 D4GL002750 1.60

norB

Nitric oxide reductase

BBGL003737 D1GL003808 D4GL003569 1.42

rutF

Flavin reductase

BBGL002597 D1GL002543 D4GL002459 1.37

–

Oxidoreductase

BBGL003509 D1GL003591 D4GL003382 1.30

nuoA

NADH-quinone BBGL002670 D1GL002617 D4GL002533 1.30 oxidoreductase subunit A

nasA

Nitrate reductase catalytic subunit

BBGL003642 D1GL003716 D4GL003513 1.22

qor

Quinone oxidoreductase

BBGL002475 D1GL002421 D4GL002333 1.18

–

2,4-dienoyl-CoA BBGL001710 D1GL001634 D4GL001564 1.18 reductase or related NADH-dependent reductase, Old Yellow Enzyme (OYE) family

–

Ferredoxin subunit of nitrite reductase or a ring-hydroxylating dioxygenase

BBGL002082 D1GL002016 D4GL001941 1.12

–

NADH:ubiquinone oxidoreductas (formate dehydrogenase)

BBGL002728 D1GL002673 D4GL002588 1.1

ccr

crotonyl-CoA carboxylase/reductase

BBGL004420 D1GL004410 D4GL004656 1.07

–

nitrite reductase (cytochrome c552, ammonia-forming)

BBGL000431 –

ndhF

oxidoreductase

BBGL004259 D1GL004249 D4GL004102 1.06

qor

NADPH2:quinone reductase

BBGL003838 D1GL003907 D4GL003668 1.06

nitR/ nirK

nitrite reductase, copper-containing

BBGL003744 D1GL003815 D4GL003577 1.06

nirD

assimilatory nitrite reductase (NAD(P)H) small subunit

BBGL003643 D1GL003717 D4GL003514 1.06

D4GL004937 1.84

–

1.07

(continued)

2.5 Molecular Mechanisms of Bacterial Reduction of Cr(VI)

77

Table 2.5 (continued) Gene name

Predicted function

Gene_ID in BB

Gene_ID in D1

Gene_ID in D4

Upregulation of proteins in BB

nuoH

NADH:ubiquinone BBGL002662 D1GL002609 D4GL002524 1.05 oxidoreductase subunit H

rutF

flavin reductase

iucC

Ferric iron reductase, BBGL000089 D1GL000086 D4GL004169 1.05 siderophore synthetase component

azoR

FMN-dependent NADH-azoreductase

BBGL003225 D1GL003304 D4GL003073 1.02

–

NAD(P)H-dependent oxidoreductase

BBGL002574 D1GL002517 D4GL002432 1.02

nqrA

NADH:ubiquinone reductase (Na(+)-transporting) subunit A

BBGL002764 D1GL002611 D4GL002527 1.00

nqrB

NADH:ubiquinone reductase (Na(+)-transporting) subunit B

BBGL002763 D1GL002610 D4GL002525 0.99

–

Putative NADPH-quinone reductase

BBGL004178 D1GL004166 D4GL004020 0.98

qor/ CRYZ

NADPH:quinone oxidoreductase

BBGL000158 D1GL000154 D4GL000039 0.98

qor/ CRYZ

NAD(P)H-quinone oxidoreductase

BBGL004083 D1GL004073 D4GL003927 0.97

–

Nitroreductase

BBGL001796 D1GL001724 D4GL001651 0.97

chrR

NADPH-dependent FMN reductase

BBGL000223 D1GL000269 D4GL000108 0.97

qor/ CRYZ

NADPH2:quinone reductase

BBGL000149 D1GL000145 D4GL000030 0.96

nirB

Nitrite reductase large subunit

BBGL003644 D1GL003718 D4GL003515 0.95

nqrF

NADH:ubiquinone reductase (Na(+)-transporting) subunit F

BBGL002759 D1GL002771 D4GL002619 0.95

nqrC

NADH:ubiquinone reductase (Na(+)-transporting) subunit C

BBGL002762 D1GL002774 D4GL002622 0.93

BBGL001797 D1GL001725 D4GL001652 1.05

(continued)

78

2 Mechanisms of Cr(VI) Reduction by Microorganisms

Table 2.5 (continued) Gene name

Predicted function

Gene_ID in BB

Gene_ID in D1

Gene_ID in D4

Upregulation of proteins in BB

nemA

Alkene reductase

BBGL001696 D1GL001619 D4GL001550 0.92

nqrD

NADH:ubiquinone reductase (Na(+)-transporting) subunit D

BBGL002761 D1GL002773 D4GL002621 0.89

Reprinted from Ref. [31] Copyright 2019, with permission from Elsevier

42]. The BB genome contains genes that are involved in the SOS repair mechanism. The medium containing 500 mg L−1 Cr(VI) greatly increased (by 30%) the activity of the DNA repair enzymes (Fig. 2.30). The DNA ligase-associated DEXH box helicase, the Holliday junction DNA helicase RuvA, and the ATP-dependent DNA helicase RecG are all overexpressed, according to a proteomics analysis. These enzymes also include the DNA mismatch repair protein MutT, DNA-3-methyladenine glycosylase I, and the DNA recombination/repair protein RecA (Fig. 2.29). Other DNA repairrelated enzymes were produced in a stable manner, and no downregulation was seen (Table 2.7). Additionally, comparative genomic research reveals numerous DNA repair enzymes that are specific to the BB genome, including the superfamily II DNA or RNA helicase, UvrD-like helicase, and archaeal DNA helicase HerA (Table 2.7 and Fig. 2.28). The high variety and specificity of gene-encoded DNA repair enzymes, which also contribute to improving the Cr(VI) resistance performance of BB, are probably a result of the significant DNA damage caused by the high levels of Cr(VI) absorption and reduction in BB. The ability of bacterial BB to absorb Cr(VI) is very strong. However, when BB is firstly inoculated into the medium with Cr(VI), there are very few cells present, and very little of the Cr(VI)-related protein expression is expressed. In order to promptly efflux the absorbed Cr(VI), the transport mechanism should be housed in BB. Additionally, bacteria have a defense mechanism known as chromate ion efflux from the cell cytoplasm, which is mediated by transporters and inhibits the buildup of harmful chromate ions inside the cells [28]. With the induction of Cr(VI), the chromate transporters of BB greatly increased their enzymatic activity (Fig. 2.30). The chromate transporter ChrA, the chromate resistance protein ChrB, and the heavy metal transport/detoxification protein CopZ are three of the nine genes connected to chromate efflux in BB. Based on the proteomic studies, their expression also markedly increased with the induction of Cr(VI) (Table 2.8). According to the reports, chrA genes only defend against Cr(VI) in the submillimolar range, in contrast to other metal resistance systems. On the other hand, significant activation of the chromate efflux pump may result in the coextrusion of sulfate; the development of sulfur-starved conditions is not conducive to bacterial growth [43]. However, even at extremely high Cr(VI) concentrations, BB continued to develop (Table 2.3). This suggests that other proteins, in addition to ChrA, are

Sulfurtransferase

Lactoylglutathione lyase

Glutathione synthase/RimK-type ligase, ATP-grasp superfamily

Class II glutamine amidotransferase BBGL000458

Glutathione transport system permease protein GsiD

Glutathione S-transferase

Glutathione S-transferase

S-(hydroxymethyl)glutathione dehydrogenase

Glutathione import ATP-binding protein GsiA

Glutathione S-transferase

Glutathione S-transferase

Glutathione S-transferase

Glutathione S-transferase

Glutathione transport system permease protein GsiC

Glutathione synthase/RimK-type ligase, ATP-grasp superfamily

–

gloA

–

–

gsiD

gst

gst

frmA/adhC

gsiA

yghU/yfcG

gst

gst

–

gsiC

–

BBGL004597

BBGL002446

BBGL002199

BBGL001255

BBGL002947

BBGL001027

BBGL001597

BBGL002821

BBGL003837

BBGL001965

BBGL001573

BBGL002180

BBGL004650

BBGL000792

BBGL003539

Sulfurtransferase

–

Gene_ID in BB

Glutathione transferases

Predicted function

Gene name

Categories

D1GL004725

D1GL002392

D1GL002153

D1GL001172

D1GL003017

D1GL000941

D1GL001522

D1GL002829

D1GL003906

D1GL001896

D1GL001499

D1GL000513

D1GL002133

D1GL004780

D1GL003621

Gene_ID in D1

D4GL004408

D4GL002304

D4GL002056

D4GL001104

D4GL002790

D4GL000874

D4GL001449

D4GL002680

D4GL003667

D4GL001819

D4GL001425

D4GL000282

D4GL002036

D4GL004355

D4GL003411 D4GL005043

Gene_ID in D4

1.07

1.07

1.07

1.07

1.08

1.08

1.09

1.11

1.12

1.17

1.24

1.32

1.33

1.50

1.80

2.01

(continued)

Upregulation of proteins in BB

Table 2.6 Predicted genes that encode ROS detoxification enzymes in Pannonibacter phragmitetus BB, D1, and D4 and the upregulation of those enzymes in BB

2.5 Molecular Mechanisms of Bacterial Reduction of Cr(VI) 79

Categories

Predicted function

Gamma-glutamyltransferase

Glutathione import ATP-binding protein GsiA

Glutathione import ATP-binding protein GsiA

Glutathione S-transferase

Glutathione metabolism protei

Gamma-glutamyltranspeptidase/ glutathione hydrolase

Glutathione-disulfide reductase

Glutathione ABC transporter ATP-binding protein

Gamma-glutamyltranspeptidase/ glutathione hydrolase

Glutathione synthase

Glutathione S-transferase

Glutathione S-transferase

Glutamate–cysteine ligase

Glutathione S-transferase

Glutathione import ATP-binding protein GsiA

Gene name

ggt

gsiA

gsiA

gst

yecN

ggt

GSR/gor

gsiA

ggt

gshB

gst

gst

gshA

–

gsiA

Table 2.6 (continued)

BBGL001572

BBGL003326

BBGL000991

BBGL001912

BBGL003818

BBGL004710

BBGL005032

BBGL003467

BBGL002804

BBGL001639

BBGL000969

BBGL000875

BBGL000359

BBGL002073 BBGL002448

BBGL004998

BBGL004168

Gene_ID in BB

D1GL001498

D1GL003406

D1GL000906

D1GL001844

D1GL003886

D1GL004840

–

D1GL002811 D1GL004532 D1GL003553

D1GL001563

D1GL000885

D1GL000792

D1GL000405

D1GL002007 D1GL002394

D1GL005001

D1GL004156

Gene_ID in D1

D4GL001424

D4GL003191

D4GL000841

D4GL001770

D4GL003647

D4GL004295

D4GL004929

D4GL004475 D4GL002666

D4GL001490

D4GL000818

D4GL000705

D4GL000240

D4GL002306 D4GL001932

D4GL004897

D4GL004010

Gene_ID in D4

0.93

0.94

0.94

0.96

0.97

0.98

0.99

0.93

0.99

0.99

0.99

1.01

1.01

1.03

1.05

1.06

(continued)

Upregulation of proteins in BB

80 2 Mechanisms of Cr(VI) Reduction by Microorganisms

Catalase-peroxidase

katG

Reprinted from Ref. [31] Copyright 2019, with permission from Elsevier

Catalase

katE/CAT/ catB/srpA

Superoxide dismutase

SOD1

Catalase HPII

Superoxide dismutase SodM-like protein

–

katE/CAT/ catB/srpA

Superoxide dismutase SodM-like protein

–

Catalases

Superoxide dismutase

SOD2

Superoxide dismutases

Predicted function

Gene name

Categories

Table 2.6 (continued)

BBGL001447

BBGL000134

BBGL003808

BBGL003794

BBGL005072 BBGL005157

BBGL005079 BBGL005166

BBGL001494

Gene_ID in BB

D1GL001371

D1GL000131

D1GL003877

D1GL003863

–

–

D1GL001419

Gene_ID in D1

D4GL001296

D4GL000017

D4GL003637

D4GL003621

–

–

D4GL001344

Gene_ID in D4

0.91

1.06

1.10

1.04

1.19

1.20

1.45

Upregulation of proteins in BB

2.5 Molecular Mechanisms of Bacterial Reduction of Cr(VI) 81

82

2 Mechanisms of Cr(VI) Reduction by Microorganisms

Fig. 2.32 Metabolism pathway of glutathione in BB. Red indicates upregulation, green indicates downregulation, and black indicates byproducts by metabolomics analysis. Reprinted from Ref. [31] Copyright 2019, with permission from Elsevier

involved in the release of chromate in BB (Fig. 2.29). Ochrobactrum tritici 5bvl1, a strain with a high tolerance to Cr(VI), has a subfamily of genes for chrB, chrA, chrC, and chrF that allows it to live at chromate concentrations greater than 50 mM [43]. O. tritici’s strong Cr(VI) resistance depends on ChrB and ChrA. The BB genome contained chrB and chrA homologs. A chromium-sensitive activator of the chr operon is called chrB. Chromate or dichromate powerfully induces the chr promoter, but Cr(III), sulfate, oxidants, or other oxyanions have no effect [43]. In BB, copZ is also involved in the transfer and detoxification of heavy metals, and there are six chrA genes and two chrB genes. The comparative genomic study supported the findings stated above. The lack of chrB and the considerably lower number of chrA genes in the genomes of D1 and D4 could be the result of adaptation to the environment. D1 and D4 are significantly less resistant to Cr(VI) because the chr locus lacks a chromatesensitive regulator. As a result, the interaction of these genes promotes chromate efflux more quickly and is a key factor in the high resilience of BB.

2.5.1.4

Double Systems for Extracellular Electron Transfer

Extracellular electron transfer (EET) is the process by which microbes transmit internal electrons from oxidative metabolism to extracellular electron acceptors (such as variable valence metal ions) via predominantly c-type cytochrome-based and soluble electron shuttle-mediated mechanisms [44, 45].

2.5 Molecular Mechanisms of Bacterial Reduction of Cr(VI)

83

Table 2.7 Predicted genes that encode DNA repair enzymes in Pannonibacter phragmitetus BB, D1, and D4 and the upregulation of those enzymes in BB Gene Predicted function name

Gene_ID in BB

Gene_ID in D1

Gene_ID in D4

Upregulation of proteins in BB

mutT

DNA mismatch repair protein BBGL000006 D1GL000006 D4GL004259 1.40 MutT

tag

DNA-3-methyladenine glycosylase I

BBGL001011 D1GL000925 D4GL000860 1.32

recA

DNA recombination/repair protein RecA

BBGL003028 D1GL003108 D4GL002874 1.29

–

DNA ligase-associated DEXH BBGL001263 D1GL001181 D4GL001113 1.18 box helicase

ruvA

Holliday junction DNA helicase RuvA

recG

ATP-dependent DNA helicase BBGL002302 D1GL002260 D4GL002160 1.13 RecG

–

Superfamily II DNA or RNA helicase

BBGL005135 –

dinB

Nucleotidyltransferase/DNA polymerase involved in DNA repair

BBGL001971 D1GL001902 D4GL001825 1.12

recR

Recombinational DNA repair protein RecR

BBGL004719 D1GL004850 D4GL004286 1.11

recQ

ATP-dependent DNA helicase BBGL004490 D1GL004480 D4GL004527 1.11

recR

MULTISPECIES: recombination protein RecR

BBGL004719 D1GL004850 D4GL004286 1.11

sbcD

DNA repair exonuclease SbcCD ATPase subunit

BBGL003813 D1GL003881 D4GL003642 1.08

ruvB

holliday junction DNA helicase RuvB

BBGL004060 D1GL004051 D4GL003905 1.07

oraA/ SOS response regulatory recX protein, interacts with RecA

BBGL004061 D1GL004052 D4GL003906 1.15

–

1.12

BBGL002362 D1GL002318 D4GL002218 1.06

uvrD/ DNA helicase II/ BBGL003180 D1GL003260 D4GL003029 1.06 pcrA ATP-dependent DNA helicase PcrA recN

DNA repair protein RecN

BBGL003185 D1GL003265 D4GL003034 1.06

sbcD

DNA repair exonuclease SbcCD nuclease subunit

BBGL003812 D1GL003880 D4GL003641 1.06

–

TPase involved in cell partioning and DNA repair

BBGL000371 D1GL000416 D4GL000252 1.05

herA

Archaeal DNA helicase HerA BBGL001433 D1GL001356 D4GL001284 1.04

uvrB

Excinuclease UvrABC helicase subunit UvrB

BBGL003311 D1GL003392 D4GL003177 1.03

radA

DNA repair protein RadA

BBGL002565 D1GL002508 D4GL002423 1.02

addB Double-strand break repair protein addb

BBGL000326 D1GL000374 D4GL000210 1.02 (continued)

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2 Mechanisms of Cr(VI) Reduction by Microorganisms

Table 2.7 (continued) Gene Predicted function name

Gene_ID in BB

Gene_ID in D1

Gene_ID in D4

Upregulation of proteins in BB

priA

Primosomal protein N' (replication factor Y) (superfamily II helicase)

BBGL000349 D1GL000396 D4GL000232 1.02

recO

DNA repair protein RecO

BBGL001738 D1GL001664 D4GL001593 1.02

recJ

Single-stranded-DNA-specific BBGL002831 D1GL002838 D4GL002689 1.02 exonuclease

addA Double-strand break repair helicase AddA

BBGL000325 D1GL000373 D4GL000209 1.01

lhr

DNA ligase-associated DEXH BBGL001262 D1GL001180 D4GL001112 1.01 box helicase

recF

DNA replication/repair protein RecF

BBGL000003 D1GL000003 D4GL004263 1.01

UvrD-like helicase C-terminal BBGL004013 D1GL004002 D4GL003857 1.00 domain ruvC

ruvC crossover junction BBGL004062 D1GL004053 D4GL003907 1.00 endodeoxyribonuclease RuvC ATP-dependent exoDNAse BBGL004013 D1GL004002 D4GL003857 0.99 (exonuclease V), alpha subunit, helicase superfamily I UvrD-like Helicase, ATP-binding domain

BBGL000379 –

–

0.97

radD

DNA repair protein RadD

BBGL000393 D1GL002704 –

0.97

mfd

Transcription-repair coupling factor

BBGL002304 D1GL002262 D4GL002162 0.97

herA

Archaeal DNA helicase HerA BBGL003938 – or a related bacterial ATPase, contains HAS-barrel and ATPase domains

mfd

Transcription-repair coupling factor (superfamily II helicase)

BBGL002304 D1GL002262 D4GL002162 0.97

hsdR

Type I restriction enzyme, R subunit

BBGL002096 –

–

0.97

D4GL001955 0.96

dnaB Replicative DNA helicase

BBGL002562 D1GL002506 D4GL002420 0.96

rarA

Recombination factor protein rara

BBGL001854 D1GL001784 D4GL001710 0.95

ligA

DNA ligase (NAD(+)) LigA

BBGL003183 D1GL003263 D4GL003032 0.95

radC

DNA repair protein RadC

BBGL002922 D1GL002918 D4GL002763 0.93

lexA

SOS-response transcriptional repressor (RecA-mediated autopeptidase)

BBGL002280 D1GL002234 D4GL002137 0.91

Reprinted from Ref. [31] Copyright 2019, with permission from Elsevier

2.5 Molecular Mechanisms of Bacterial Reduction of Cr(VI)

85

Table 2.8 Predicted genes that encode chromate transporters in Pannonibacter phragmitetus BB, D1, and D4 and the upregulation of those enzymes in BB Gene name

Predicted function

Gene_ID in BB

Gene_ID in D1

Gene_ID in D4

Upregulation of proteins in BB

chrB

ChrB protein, Chromate resistance protein

BBGL005072

–

–

1.55

copZ

Heavy metal transport/ detoxification protein

BBGL003867

–

–

1.52

chrA

Chromate transporter, chromate ion transporter family

BBGL005156 BBGL005148

–

–

1.44 1.43

chrA

Chromate transporter

BBGL000793 BBGL003540

D1GL003622

D4GL005042 D4GL003412

1.28 1.18

chrB

ChrB protein, Chromate resistance protein

BBGL005157

–

–

1.19

chrA

Chromate transporter

BBGL004927

D1GL005082

D4GL004813 D4GL004579

0.90

chrA

Chromate transporter

BBGL005159

–

–

nda

a nd

not detected by proteomic analysis Reprinted from Ref. [31] Copyright 2019, with permission from Elsevier

Since BB is a facultative anaerobe, it is possible for Cr(VI) to be reduced anaerobically there when oxygen levels are low. Membrane-bound reductases, such as cytochromes, flavin reductases, and hydrogenases, are typically involved in anaerobic reduction. The electron transfer system that employs chromate as the terminal electron acceptor includes these enzymes as part of its components [28]. To quickly react to Cr(VI) pressure, three cytochromes, including two cytochrome c oxidases and a bacterioferritin, were substantially upregulated in BB (Table 2.9). Cytochrome c oxidase activity rises from 44.21 to 51.14 U L−1 (Fig. 2.30). The crucial component connecting electron transfer across various length scales is cytochrome c, a redox protein found in dissimilatory metal-reducing bacteria’s extracellular electron transfer system [46]. As a result, when BB grows in a medium containing Cr(VI), the electrons generated by ubiquinone move through cytochrome b (found in BB’s genome), cytochrome c, and then cytochrome c oxidase, which converts extracellular Cr(VI) to Cr (III). Then, on bacterial cell surfaces like extracellular polymers, Cr(III) makes bonds with functional groups like carboxyl,

Shikimate dehydrogenase

Dihydroorotate dehydrogenase

Saccharopine dehydrogenase

Pyruvate dehydrogenase

aroE

–

–

pdhA

Cytochrome c553

–

Aspartate-semialdehyde dehydrogenase

Cytochrome c oxidase, subunit I domain

–

–

Bacterioferritin, cytochrome b1

–

Glutaryl-CoA dehydrogenase

Ubiquinol-cytochrome c reductase cytochrome b subunit

petB

gcdH

Bacterioferritin, cytochrome b1

bfr

Dehydrogenases

Cytochrome-c oxidase

ccoO

BBGL002619

BBGL003014

BBGL001239

BBGL004677

BBGL004411

BBGL004484

BBGL000761

BBGL004367

BBGL000645

BBGL005071 BBGL003372

BBGL001036

BBGL000762

Cytochrome c oxidase caa3 BBGL000770 assembly factor family protein

ctaG

Gene_ID in BB

Cytochromes

Predicted function

Gene name

Categories

D1GL002566

D1GL003092

D1GL001264

D1GL004807

D1GL004401

D1GL004474

–

D1GL004359

D1GL000652

D1GL003452

D1GL000949

–

–

Gene_ID in D1

D4GL002481

D4GL002861

D4GL001191

D4GL004328

D4GL004664

D4GL004533

–

D4GL004712

D4GL000882

D4GL003237

D4GL000472

–

–

Gene_ID in D4

1.31

1.35

1.45

1.47

1.62

1.74

1.29

1.35

1.36

1.44

1.61

1.81

2.66

(continued)

Upregulation of proteins in BB

Table 2.9 Predicted genes that encode membrane-bound reductases involved in electron transfer on the cell membrane in Pannonibacter phragmitetus BB, D1, and D4 and the upregulation of those reductases in BB

86 2 Mechanisms of Cr(VI) Reduction by Microorganisms

Gluconate 2-dehydrogenase BBGL003791

Lactate dehydrogenase

Short-chain dehydrogenase

Uptake hydrogenase small subunit

3-oxoacyl-ACP reductase, NAD(P)H-dependent dehydrogenase

Short-chain dehydrogenase

2-hydroxy-acid oxidase, FAD/FMN-containing dehydrogenase

–

–

fabG

hyaA/hybO

–

–

–

BBGL001005

BBGL002745

BBGL001979

BBGL003430

BBGL002819

BBGL003322

BBGL000842

Shikimate dehydrogenase

aroE

Gene_ID in BB

Predicted function

Gene name

Reprinted from Ref. [31] Copyright 2019, with permission from Elsevier

Categories

Table 2.9 (continued)

D1GL000919

–

D1GL001912

D1GL003507

D1GL002826

–

D1GL003860

D1GL000759

Gene_ID in D1

D4GL000854

–

D4GL001835

D4GL003296

D4GL002678

–

D4GL003618

D4GL000670

Gene_ID in D4

1.20

1.20

1.22

1.23

1.23

1.27

1.28

1.28

Upregulation of proteins in BB

2.5 Molecular Mechanisms of Bacterial Reduction of Cr(VI) 87

88

2 Mechanisms of Cr(VI) Reduction by Microorganisms

amine, hydroxyl, phosphate, and sulfhydryl groups (Fig. 2.29). The CtaG and CcoO cytochrome c oxidases, which are terminal enzymes of electron transport, are upregulated in Pannonibacter phragmitetus BB but are absent from D1 and D4. Their lack makes it more difficult for electrons to reach extracellular through the cell membrane Cr(VI) [28]. In order to encourage Cr(VI) reduction, BB can also use a variety of other carbon sources as electron donors, such as pyruvate, citrate, formate, lactate, lactose, fructose, glucose, NADPH, and NADH, with lactate being the preferred one (data not shown). These compounds facilitate the dehydrogenation process, which enhances Cr(VI) reduction. Pyruvate dehydrogenase and lactate dehydrogenase are substantially upregulated in the presence of Cr(VI), according to the proteomic analysis. Bacterial extracellular electron shuttles (EESs) are tiny molecules [47] that transport metabolic electrons to extracellular electron acceptors such as Cr(VI). They are frequently regarded as redox antibiotics and may be important in bioremediation [48, 49]. EESs frequently take the shape of heterocyclic aromatic rings and differ from their intracellular cousins [such as NAD(P)H, FAD, and quinones] by having conjugated bonds [47, 50]. The chromium biosynthesis of BB did not contain these conventional EESs. Although pyrrole-2-carboxylic acid (C5 H5 NO2 ) is a shared upregulated element in T1, T2, and T3 compared to C1, C2, and C3, respectively, the metabolomic analyses of BB cultured with and without Cr(VI) demonstrate this molecule is a common upregulated element in T1, T2, and T3 (Table 2.5). In addition, Cr(VI) is reduced 2 h sooner when C5 H5 NO2 is present (3 mM; Fig. 2.33a). Anaerobic ammonium oxidation and iron(III) reduction were boosted as a result of the exogenous inclusion of ESSs in a dissimilatory iron(III)-reducing bacterium [51]. The hue of C5 H5 NO2 , a typical heterocyclic compound with a conjugated bond, changes with pH from white to off-white to pale pink [52]. Generally, the redox activity of the ESS originates from the double bonds, which can be reduced and rearranged at biologically accessible reduction potentials [47]. As an EES, C5 H5 NO2

Fig. 2.33 Reduction of Cr(VI) by Pannonibacter phragmitetus BB with and without 3 mM extracellular electron shuttle pyrrole-2-carboxylic acid (C5 H5 NO2 ) (a); the growth and Cr(VI)-reducing profile of BB with and without 500 mg L−1 Cr(VI) (b). Reprinted from Ref. [31] Copyright 2019, with permission from Elsevier

2.5 Molecular Mechanisms of Bacterial Reduction of Cr(VI)

89

Table 2.10 Predicted genes that encode SoxR transcription factors and Fe-S clusters involved in EES recognition in Pannonibacter phragmitetus BB and their upregulation by proteomic analysis Categories

Gene name

Predicted function

Gene_ID in BB

Upregulation of proteins in BB

Transcriptional regulators

soxR

LysR family transcriptional regulator

BBGL003355

1.07

soxR

MerR family transcriptional regulator

BBGL003275

0.91

soxR

LysR family transcriptional regulator

BBGL004467

0.89

iscA

Fe-S cluster assembly iron-binding protein IscA

BBGL002527

1.11

–

2-hydroxy-acid oxidase, Fe-S oxidoreductase, cysteine-rich domain

BBGL001006

1.09

sufC

Fe-S cluster assembly ATP-binding protein

BBGL002531

1.05

ygfZ

Folate-binding Fe-S cluster repair protein YgfZ

BBGL001174

1.04

sufE

Cysteine desufuration protein SufE, Fe-S center assembly

BBGL001508

1.01

sufB

Fe-S cluster assembly protein

BBGL002532

1.01

Fe-S clusters

Reprinted from Ref. [31] Copyright 2019, with permission from Elsevier

strongly encourages BB reduction of Cr(VI). In addition, the BB genome contained three soxR genes. The most well-known transcription factors for EES detection are the SoxR genes. Two of the genes are consistently expressed, while the third is increased in response to Cr(VI) stimulation (Table 2.10). Through a FeeS cluster connected by a special cysteine motif, SoxR detects redox-active molecules, and its stimulation encourages sequence-specific binding and transcriptional activation [53]. Based on proteomic analysis, the FeeS cluster in BB was also substantially upregulated (Table 2.10). Therefore, on the basis of a FeeS cluster connected to a specific cysteine pattern in BB, overexpressed SoxR detects C5 H5 NO2 (Fig. 2.29).

2.5.1.5

Cr(VI)-Induced Quorum Sensing Facilitates the Interactions Between the Cells

The decrease of Cr(VI) by BB is slower than bacterial growth. The bacteria expanded quickly, but 4 h later, Cr(VI) was begun to decrease (Fig. 2.33b). This suggests that BB must first achieve a certain bacterial concentration in order to begin reducing Cr(VI), as Cr(VI) reduction cannot occur during the initial phase. Bacterial quorum sensing, which is the capacity to recognize and react to the cell population density

90

2 Mechanisms of Cr(VI) Reduction by Microorganisms

by gene regulation and a means of communication to enable interactions between bacterial cells in a population, can be used to explain this phenomenon [54]. It is mediated by small diffusible signaling molecules, so-called autoinducers. Table 2.11 lists several proteins involved in quorum sensing in BB. These proteins were considerably elevated, according to the proteomic analyses. Many proteins involved in amino acid transport and secretion are overexpressed, primarily to supply energy for autoinducer production and efflux. Furthermore, the genome of BB contains homologs of Vibrio fischeri’s LuxI/LuxR bioluminescence system for quorum sensing [55]. Acyl homoserine lactone (AHL), an autoinducer, is produced by BB when it comes into contact with Cr(VI). S-adenosylmethionine is used as an amino donor by autoinducer synthases (LuxI-type) to create the homoserine lactone ring moiety, and a charged (acylated) carrier protein serves as a precursor to create the acyl side chain (Fig. 2.34 and Table 2.11). AHL is present in extremely low amounts. As a result, metabolomic analysis is unable to identify AHL because the majority of it is released from the cell. Yet when N-3-oxohexanoylhomoserine lactone (OHHL), an AHL produced by LuxI synthase, is added, bacterial Cr(VI) reduction starts 2 h earlier. As the extracellular concentration of OHHL reaches a specific threshold, a signal transduction cascade causes population-wide modifications in gene expression and the start of “cooperative” behaviors [56, 57]. Furthermore, the homologs of the TraI/TraR system of activators of conjugal transfers of Ti plasmids from Agrobacterium tumefaciens were found in the genome of BB. One of the LuxI/LuxR homologs, the TraI/TraR system contributes to the production of OHHL [58, 59]. Cr(VI) stimulation in BB activates the LuxI/LuxR and TraI/TraR quorum sensing systems, enhancing bacterial communication and promoting Cr(VI) reduction (Fig. 2.29).

2.5.1.6

Negative Chemotaxis to Cr(VI) Guarantees the Survival of Pannonibacter Phragmitetus BB

With the help of flagella-based movement, bacteria follow gradients of environmental stimuli (such as pH, redox potential, and osmolality) [60]. According to Fig. 2.35, which shows that BB exhibits negative chemotaxis to Cr(VI), a noticeable flagellum at the terminals of BB helps bacteria leave Cr(VI) in a semisolid media and leaves a clear, transparent circle around Cr(VI). Signal transduction and adaptation make up the chemotaxis pathway [61]. The comparable gene cluster was located in the BB genome between GL004026 and GL004030 (Fig. 2.36a and Table 2.12). Moreover, fourteen overexpressed methyl-accepting chemotaxis proteins (MCPs) are among the 24 chemotaxis-associated proteins that facilitate signal transmission (Table 2.12). The transmembrane chemoreceptors known as MCPs help BB detect chemical gradients in the surrounding environment. The MCPs of E. Coli contained five receptors (Tar, Tsr, Trg, Tap, and Aer) with various ligands [62]. Whereas Trg, Tap, and Aer are receptors that are more typical of sugars, dipeptides, and the redox potential, respectively, Tar and Tsr are considered amino-acid receptors and perceive a variety of stimuli [62]. Trg favours ribose and D-galactose while Tar has the highest

2.5 Molecular Mechanisms of Bacterial Reduction of Cr(VI)

91

Table 2.11 Predicted genes that encode proteins related to quorum sensing, transcriptional regulators, and enzymes involved in AHL synthesis in Pannonibacter phragmitetus BB and their upregulation by proteomic analysis Categories

Gene name Predicted function

Quorum ABC.SP.P1 Putative spermidine/putrescine sensing-related transport system permease protein proteins ABC.PE.P Peptide/nickel transport system permease protein

Quorum sensing transcriptional regulators

ACPs/ACP synthases

Gene_ID in BB

Upregulation of proteins in BB

BBGL001191 1.92 BBGL001595 1.50

trbD

Type IV secretion system protein TrbD

BBGL005102 1.38

trbE

Type IV secretion system protein TrbE

BBGL000782 1.31

xagB

Glycosyltransferase XagB

BBGL001244 1.27

ABC.SP.P1 Putative spermidine/putrescine transport system permease protein

BBGL001349 1.27

livF

BBGL004795 1.26

Branched-chain amino acid transport system ATP-binding protein

ABC.PE.P1 Peptide/nickel transport system permease protein

BBGL001573 1.24

livH

Branched-chain amino acid transport system permease protein

BBGL000634 1.23

livG

Branched-chain amino acid transport system ATP-binding protein

BBGL001725 1.22

ABC.PE.P1 Peptide/nickel transport system permease protein

BBGL001570 1.20

trbL

Type IV secretion system protein TrbL

BBGL000779 1.20

ABC.PE.S

Peptide/nickel transport system substrate-binding protein

BBGL002445 1.19

sinR

LuxR family transcriptional regulator

BBGL005114 1.15

cciR

LuxR family transcriptional regulator

BBGL000618 1.13

traI

Conjugal transfer protein TraI

BBGL003889 1.10

traR

LuxR family transcriptional regulator, activator of conjugal transfer of Ti plasmids

BBGL005111 0.88

fabF

Beta-ketoacyl-[acyl-carrier-protein] BBGL000116 1.06 synthase II

–

Holo-[acyl carrier protein] synthase BBGL001742 1.06

–

Acyl carrier protein-like

BBGL002795 1.05 (continued)

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2 Mechanisms of Cr(VI) Reduction by Microorganisms

Table 2.11 (continued) Categories

SAM uptake transporters/ SAM synthases

Gene name Predicted function

Gene_ID in BB

Upregulation of proteins in BB

fabB

Beta-ketoacyl-[acyl-carrier-protein] BBGL004683 1.05 synthase I

fabF

Beta-ketoacyl-[acyl-carrier-protein] BBGL002555 1.00 synthase II

acpP

MULTISPECIES: acyl carrier protein

BBGL002556 0.83

sam

S-adenosylmethionine uptake transporter

BBGL000933 1.19

sam

S-adenosylmethionine uptake transporter

BBGL002047 1.18

sam

S-adenosylmethionine uptake transporter

BBGL003102 1.14

metK

S-adenosylmethionine synthase

BBGL000205 1.03

sam

S-adenosylmethionine uptake transporter

BBGL001065 1.02

Reprinted from Ref. [31] Copyright 2019, with permission from Elsevier

Fig. 2.34 Schematic diagram illustrating the features of the AHL biosynthetic pathway in BB. ACP, acyl carrier protein. Reprinted from Ref. [31] Copyright 2019, with permission from Elsevier

affinity for aspartate. According to the metabolomics investigation, aspartic acid is increased in T2, T2, and T1 compared to T3, C2, and T3, respectively. Additionally, 2-Deoxy-D-galactose is upregulated in T2 compared to T1, suggesting that the absence or reduction of Cr(VI) causes the reduced aspartate and increased 2-DeoxyD-galactose to induce Tar and Trg, respectively, leading to the negative chemotaxis of BB to Cr(VI). However, compared to C3 and C2, D-ribose substantially dropped in T3 and T2, respectively, leading to bacterial genetic changes. Typically, MCPs make homodimers that self-assemble into trimers of dimers by forming bonds with one ligand molecule per dimer. Additionally, they associate with two cytoplasmic

2.5 Molecular Mechanisms of Bacterial Reduction of Cr(VI)

93

Fig. 2.35 Chemotactic responses of BB to 500 mg L−1 Cr(VI) in drop assays and 0.9% NaCl as a control. a Original image, b black and white image for clear observation of the local translucent circle. Reprinted from Ref. [31] Copyright 2019, with permission from Elsevier

Fig. 2.36 The gene cluster of chemotaxis (a) and chemotaxis pathways (b) in Pannonibacter phragmitetus BB. Green represents the pathway induced by Cr(VI), and red represents the default pathway. Reprinted from Ref. [31] Copyright 2019, with permission from Elsevier

94

2 Mechanisms of Cr(VI) Reduction by Microorganisms

Table 2.12 Predicted genes that encode proteins involved in negative chemotaxis of Pannonibacter phragmitetus BB to Cr(VI) and their upregulation by proteomic analysis Modules

Gene name

Predicted function

Gene_ID in BB

Upregulation of proteins in BB

Signal transduction

cheA

Chemotaxis histidine kinase CheA

BBGL004030

1.145

cheW

Purine-binding chemotaxis protein CheW

BBGL004029

1.23

cheY

Chemotaxis protein CheY

BBGL001226

1.14

cheY

Signal transduction response regulator

BBGL000100

1.14

cheY

Chemotaxis protein CheY

BBGL004028

1.145

cheZ

Chemotaxis protein CheZ

BBGL001010

1.02

cheZ

Chemotaxis protein CheZ

BBGL001069

0.97

cheZ

Chemotaxis protein CheZ

BBGL003093

0.94

mcp

Methyl-accepting chemotaxis protein

BBGL001633

1.29

mcp

Methyl-accepting chemotaxis protein

BBGL002672

1.23

mcp

Methyl-accepting chemotaxis protein

BBGL001525

1.22

mcp

Methyl-accepting chemotaxis protein

BBGL004902

1.18

mcp

Methyl-accepting chemotaxis protein

BBGL001401

1.15

mcp

Methyl-accepting chemotaxis protein

BBGL004606

1.15

mcp

Methyl-accepting chemotaxis protein

BBGL001485

1.145

mcp

Methyl-accepting chemotaxis protein

BBGL000648

1.14

mcp

Methyl-accepting chemotaxis protein

BBGL000254

1.13

mcp

Methyl-accepting chemotaxis protein

BBGL000623

1.13

mcp

Methyl-accepting chemotaxis protein

BBGL003361

1.13 (continued)

2.5 Molecular Mechanisms of Bacterial Reduction of Cr(VI)

95

Table 2.12 (continued) Modules

Adaptation

Motor

Gene name

Predicted function

Gene_ID in BB

Upregulation of proteins in BB

mcp

Methyl-accepting chemotaxis protein

BBGL004416

1.10

mcp

Methyl-accepting chemotaxis protein

BBGL004324

1.10

mcp

Methyl-accepting chemotaxis protein

BBGL001852

1.09

cheR

Chemotaxis protein methyltransferase CheR

BBGL004026

1.14

cheB

Glutamate methylesterase

BBGL004027

1.07

fliG

Flagellar motor switch protein FliG

BBGL000562

1.03

fliM

Flagellar motor switch protein FliM

BBGL000565

1.03

motA

Flagellar motor protein MotA

BBGL002458

1.19

motB

Flagellar motor protein MotB

BBGL004143

1.15

pomA

Flagellar motor protein PomA

BBGL004142

1.11

motA

Flagellar motor protein MotA

BBGL004080

1.07

Reprinted from Ref. [31] Copyright 2019, with permission from Elsevier

proteins, CheA, a histidine protein kinase, and CheW, a scaffolding protein, to create stable core signaling complexes [63]. To detect complex environmental gradients, these signaling complexes are arranged in large-scale clusters in the membrane [64]. Regarding the growth of BB, Cr(VI) is chemorepellent. The cells move and swim randomly in a counterclockwise direction in the absence of Cr(VI). The interaction of Cr(VI) with Tar and Trg of the MCPs during swimming causes a change in the balance between the various cytoplasmic signaling domain conformations, which encourages (ON) the autophosphorylation activity of CheA. If not, the chemoattractants would stop (OFF) CheA’s action. As a result, the crucial phosphoryl groups for active CheAs are swiftly transferred to the tiny diffusible protein CheY to create CheYP. CheY-P causes clockwise rotation and tossing by attaching to flagellar motors (FliM). When the environment shifts, the phosphatase CheZ helps CheY-P dephosphorylate quickly (Fig. 2.36b). Signal transduction allows for the detection and fast transmission of signals that are amplified in response to environmental changes and used to control bacterial swimming behavior. The methyltransferase (CheR) and methylesterase (CheB) functions of the adaptation progressively reset the average

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kinase (CheA) activity of the receptor arrays toward an intermediate operation point [61]. In BB, CheR and CheB expression levels increased to 1.14 and 1.07, respectively (Table 2.12). CheR and CheB catalyze the process, which changes the output toward the opposing state by interacting with respective ON-state and OFF-state receptors [65]. CheB’s enzymatic activity is increased by active CheA’s phosphorylation, which speeds up the BB’s exit from Cr(VI) (Fig. 2.5b) [66]. The brief docking of CheR and CheB at the C-terminus of high-abundance receptors results in signal adaptation, which is facilitated by receptor clusters, but CheB phosphorylation is not involved in this process and instead affects a number of nearby receptors [67, 68]. The movement of an organism in reaction to a chemical stimulus is known as chemotaxis. When BB is inoculated in the medium with Cr(VI), the first bacterial reaction to Cr(VI) is the negative chemotaxis of BB in response to Cr(VI). In the early growth period with low biomass, this mechanism aids BB in surviving in an environment with high concentrations of Cr(VI) and strengthens BB’s resistance to Cr(VI).

2.5.2 Discerning Chromate Reduce and Transport Genes of Highly Efficient Pannonibacter Phragmitetus BB 2.5.2.1

General Genomic Features of Pannonibacter Phragmitetus BB

With a G + C content of 63.5% and a length of 5.07 Mb, the Pannonibacter phragmitetus BB entire genome sequences were large. The high-quality reads were assembled into 366 contigs and 25 scaffolds with a 33.20-fold sequencing coverage. The contigs’ N50 and N90 sizes, respectively, were 682,555 and 150,268 bp. The genome’s repeat content was 0.33% and had a length of 5,074,267 bp overall. 4686 genes were predicted in the draft genome sequences, and 4503 CDSs were connected to COG and KEGG. In addition, 1 rRNA, 39 tRNA, and 1 other non-coding RNA (ncRNA) operon were found in the genome. In addition, three CRISPRs (clustered regularly interspaced short palindromic repeats) were discovered in the genome (Table 2.13). Specifically when compared to the genomes of other Pannonibacter phragmitetus strains.

2.5.2.2

Genes Related to Chromate Reduction and Transport in the Genome

The NCBI Prokaryotic Genome Annotation Pipeline was used to annotate all CDSs in the Pannonibacter phragmitetus BB genome using the KEGG and COG databases. The genes involved in chromium biotransformation are listed in Table 2.14 and fall into the two functional categories of chromate transport and reduction. Despite the fact that Pannonibacter phragmitetus BB has strong Cr(VI) reduction abilities, genomic analysis did not reveal any genes that are related to the previously reported classic chromate reductases. However, some bacterial enzymes, such as a

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97

Table 2.13 General genomic properties of the selected Pannonibacter strains Strains

P. phragmitetus BB

P. indicus 31801

CGMCC9175

DSM 14782 2340 [70]

Accession NZ_ CP013068 LGSQ01000001.1 NZ_ NZ_ LDOP00000000.1 KB908215.1 LIPT01000001.1 Size (Mb) 5.07

5.32

5.58

4.78

4.17

GC%

63.5

63.3

63.6

63.1

63.5

Protein

4411

4997

4936

4122

3703

rRNA

1

9

3

4

7

tRNA

39

54

50

49

50

Gene

4686

5150

5060

4232

3845

Pseudo gene

142

81

70

65

84

CRISPR

3

4

6

3

3

Plasmid

0

1

0

0

0

Reprinted from Ref. [69], Copyright 2018, with permission from Elsevier

nitroreductase from Vibrio harveyi KCTC 2720 and NfsA from E. coli, have been demonstrated to possess a bacterium capabilities to reduce Cr(VI) [30, 71]. Scaffold5 gene 106, called nitR, shares a high nucleotide or protein identity with nitrite reductases (Table 2.14) [69]. This implies that nitR may be a brand-new chromate reductase. The two Scaffold2 gene 831 and Scaffold7 gene 96 genes, known as chrA1 and chrA2, sequence chromate transport proteins that are vital to carrying Cr(VI) during the chromium biotransformation process [69]. To confirm their roles in the chromium biotransformation of Pannonibacter phragmitetus BB, further research should concentrate on nitR, chrA1 and chrA2.

2.5.2.3

Phylogenetic Relationship of Bacterial Proteins Related to Chromate Reduction and Transport

By employing the equivalent proteins of nitR (NitR) versus classic chromate reductases and chrA1 and chrA2 (ChrA1 and ChrA2) versus classic chromate transporters, respectively, two phylogenetic trees were generated using MEGA 6.0 to demonstrate the dependability of the chosen genes. Phylogenetically connected to the nitrite reductase of Niveispirillum irakense is the putative chromate reductase NitR of Pannonibacter phragmitetus BB (Fig. 2.37b). This interaction is comparable to that between NfsA, a semitight chromate reductase, and the nitroreductase of Bacillus subtilis QB928. They are similar to a few other chosen chromate reductases. Nitrofurazone, trinitrotoluene, or chromate serve as the substrate for the action of the nitroreductase NfsB of Vibrio harveyi or E. coli [30]. The chromate reductase (ChrR) of Pseudomonas putida KT2440 is a conventional two electron enzyme. Moreover, the periplasmic c type cytochromes MtrC and OmcA of S. oneidensis MR-1 have the ability to decrease chromate [71]. Thus, it is hypothesized based on phylogenetic

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Table 2.14 Putative chromate reduction and transport genes in the genome of Pannonibacter phragmitetus BB Gene location

Putative name

Nucleotide

Protein

Similarity sequence Identities

Similarity sequence Identities

Scaffold5_ gene_106

nitR

Nitrite reductase, copper-containing protein [Starkeya novella DSM 506]

832/1007 (83%)

Nitrite reductase, copper-containing [Niveispirillum irakense]

297/352 (84%)

Scaffold2_ gene_831

chrA1

Chromate transport protein ChrA [Pelagibacterium halotolerans B2]

910/1202 (76%)

Chromate transporter [Devosia riboflavina]

414/454 (91%)

Scaffold7_ gene_96

chrA2

Chromate transporter [Rhizobium sp. S41]

1152/1347 (86%)

Chromate transporter [Paracoccus hallophius]

411/455 (90%)

Reprinted from Ref. [68], Copyright 2018, with permission from Elsevier

Fig. 2.37 Circular representation of the Pannonibacter phragmitetus BB genome against the completed chromosome of Pannonibacter phragmitetus 31801 (a). The inner to outer circles correspond to (1) GC content (percentage) as a peak to valley profile in black; (2) GC skew (−) graph in purple and GC skew (+) in green; and (3) identity to complete sequences of the Pannonibacter phragmitetus 31801 genome. Phylogenetic trees of NitR versus classic chromate reductases (b) and ChrA1 and ChrA2 against verified chromate transporters (c) (neighbor-joining method, 1000 bootstraps). The arsenic transporter or arsenic reductase was the reference protein in the corresponding phylogenetic tree. Reprinted from Ref. [69], Copyright 2018, with permission from Elsevier

analysis that NitR is a new chromate reductase of Pannonibacter phragmitetus BB with a high capacity to reduce Cr(VI). Comparing the chromate transporters of Paracoccus hallophius and Devosia riboflavin to the reference protein, arsenic transporter, ChrA1 and ChrA2 are also

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phylogenetically associated with these proteins (Fig. 2.37c). Moreover, proteins share tight genetic affinities with three homologs of chromate efflux transporters as well as ChrA from Escherichia coli. ChrA1 and ChrA2 may therefore perform similar roles to these chromate transporters. To reduce the toxicity of Cr(VI) to the cells, for instance, ChrA, a well-researched and effective chromate transporter, can supply anions efflux from the cytoplasm. ChrA1 and ChrA2, which are putative chromate transporters, are essential for Pannonibacter phragmitetus BB’s ability to withstand Cr(VI).

2.5.2.4

Transcription Profiles of Putative Chromate Reduction and Transport Genes

The relative cDNA copies of the nitR, chrA1, and chrA2 genes were found by RTqPCR as maximum rates of Cr(VI) reduction and cell growth with various initial Cr(VI) concentrations. This allowed us to decipher the transcription profiles of the putative chromate reduction and transport genes of Pannonibacter phragmitetus BB. According to Fig. 2.38a, Pannonibacter phragmitetus BB removed up to 1000 mg L−1 Cr(VI) entirely and quickly in less than 16 h. The Pannonibacter phragmitetus BB cells were collected after the Cr(VI) reduction process had reached the midpoint of the logarithmic phase (the point of maximum Cr(VI) reduction rate). The use of RT-qPCR is a reliable technique for identifying gene transcription, which reveals the level of the associated enzymes’ expression contents and their ability to withstand pollutants [72]. The 2−ΔΔCt method was used to examine the RT-qPCR findings. With rising Cr(VI) concentrations, the relative quantitative transcription of the three genes in the greatest reduction rate of Cr(VI) phases was considerably increased (p < 0.05). (Fig. 2.38b). At an initial Cr(VI) concentration of 1000 mg L−1 , the relative quantitative expression of nitR, chrA1, and chrA2 was 162.7, 448.2, and 258 times of that in Cr(VI)-free culture media, respectively. Moreover, nitR’s early expression lagged behind that of chrA1 and chrA2, respectively. This suggests that Pannonibacter phragmitetus BB’s chromate transporters may be crucial for the organism’s ability to tolerate the low concentrations of Cr(VI), whereas development of nitR is a key tactic for dealing with the problem of higher concentrations of Cr(VI). Figure 2.38c depicts the growth dynamics of Pannonibacter phragmitetus BB under various Cr(VI) pressures, and the cells were taken out when the bacteria were growing at their fastest rates. Pannonibacter phragmitetus BB, as depicted in Fig. 2.38c, was able to grow from 0 to 1000 mg L−1 Cr(VI), and the entire growth curves were not substantially different from one another. This finding showed that this strain had strong tolerance to toxic Cr(VI). The results of RT-qPCR showed that nitR had low transcription levels, whereas chrA1 and chrA2 transcription levels were maintained at a relatively high level and reduced with rising the initial Cr(VI) concentrations (Fig. 2.38d). This was related to the fact that, as shown in Fig. 2.38a, the samples with the fastest growth rates of bacterial growth were taken at 12 h for initial Cr(VI) concentrations of 0 to 500 mg/L and at 16 h for 1000 mg/L Cr(VI), when Cr(VI) had entirely declined. Although there was no discernible suppression of

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Fig. 2.38 Dynamics of Cr(VI)-reducing (a) and bacterial cell growth (c), transcription profiles of nitR, chrA1 and chrA2 as Pannonibacter phragmitetus BB at the points of maximum speeds of Cr(VI) reduction (b) and cell growth (d) under pressure of initial Cr(VI) concentrations from 0 to 1000 mg/L. The values are presented as the mean ± S.D. (standard deviation) (n = 3), the error bar is the S.D. Different lowercase letters above bar charts indicate significant differences (p < 0.05) between members in the group. Reprinted from Ref. [69], Copyright 2018, with permission from Elsevier

Pannonibacter phragmitetus BB’s cell development under pressure of up to 1000 mg/ L Cr(VI), the findings suggested that Cr(VI) was slightly toxic to the strain BB. It was better than S. oneidensis MR-1, a traditional model for researching Cr(VI) reduction which was markedly suppressed under either anaerobic or aerobic circumstances and back to normal when Cr(VI) was entirely reduced [73–75]. As a result, it’s essential that nitR transcription levels were low until Cr(VI) had been entirely decreased, but large transcription levels of chrA1 and chrA2 were necessary for detoxification through chromium influx/eflux from the bacterial cells. It is analogous to the fact that many nitroreductases and nitrite reductases have Cr(VI) reduction functions [29, 76] that nitR encodes a nitrite reductase in addition to being able to reduce Cr(VI). It was clear that NitR was a Cr(VI)-inducible reducing enzyme because the relative transcription of the enzyme rose greatly with rising Cr(VI) concentrations and stayed at a low level as Cr(VI) was completely reduced. The reduction of Cr(VI) may not be the main biological function of some chromate reductases, which are inducible by pollutants [29, 77].

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ChrA1 and ChrA2 have been identified as chromate transport proteins and are tightly associated with Cr(VI) resistance in Pannonibacter phragmitetus BB. It’s possible that the chromosomal operon or a collection contains the chrA1 and chrA2 genes. The Ochrobactrum tritici 5bv11 chr genes, which are arranged as chrB, A, C, and F in a transposable operon, are similar [43]. The chromium locus is controlled by ChrB, which is chromium-sensitive. Chromate transporters are proteins that are expressed by chrC and F [43]. This suggests that ChrA1 and ChrA2 might have distinct functions in chromate transport and that chromium might regulate the two genes in different ways, which might account for the two genes’ different levels of transcription. The chromate transport operons of Bacillus cereus SJ1 and Arthrobacter sp. FB24 are parallel to the Cr(VI)-inducible chrA1 and chrA2 genes [78, 79].

2.5.2.5

Ex Situ Chromate Reducing Activity of NitR

In E. coli BL21, the suspected chromate reductase gene, nitR, was genetically cloned and expressed heterogeneously. It was discovered that NitR, a purified encoded enzyme, had chromate reduction action. The potential influences on NitR’s ability to reduce Cr(VI) were taken into account, and it was discovered that the enzyme’s activity was best at pH 7.0 and 35 °C (Fig. 2.39a). The maximum enzyme activity of NitR was found to be pH-sensitive, losing more than 50% of that activity at pH values of 6 or 8, but temperate-tolerant, keeping 82% of that activity at 50 °C for 30 min. (Fig. 2.39b). However, the pH range between 9 and 11 is ideal for Pannonibacter phragmitetus BB’s capacity to reduce Cr(VI). The bacterial outer membrane may have proton pumps that can move hydrogen ions in order to keep the intracellular pH in a comparatively stable neutral condition [80]. Under a range of initial Cr(VI) concentrations with enough NADPH, the NitR’s Cr(VI)-reducing kinetics were studied. Before the equilibrium period, the Cr(VI) concentration fluctuations in the solutions were measured (Fig. 2.39c). The Michaelis–Menten equation was used to fit the Cr(VI)-reducing kinetics of NitR as follows: υ0 = Vmax C[Cr(VI)]/(km + C[Cr(VI)]), where V max and K m are constant, the K m value represents the concentration of the substrate at the maximum reduction rate of NitR, V0 represents the reduction velocity, and C[Cr(VI)] represents the substrate concentration.V max is 34.46 μmol/min/mg according to the data, and K m is 14.55 μmol/L (Fig. 2.39d). The fitting curve is quite reasonable because the fitting coefficient (R2 ) is 0.9845. According to the graph, the Cr(VI)-reducing kinetics reaction of NitR is first-order at low Cr(VI) concentrations, mixed-order at intermediate Cr(VI) concentrations, and zero-order at constant reaction rates. This shows that the Michaelis–Menten equation (R2 = 0.9845) is perfectly satisfied by Cr(VI) reduction by NitR (Fig. 2.39d). With NADPH as electron donors, ChrR, YieF, and NfsA had Cr(VI)-reducing Vmaxes of 8.8, 5.0, and 0.25 μmol/min/mg, respectively, and K m of 260, 200, and 36 μM [81]. However, Pseudomonas putida and E. coli, the hosts of the enzymes, have Cr(VI)-reducing capacities that are lesser than 2.1% of that of Pannonibacter phragmitetus BB [81]. The chromate reduction ability of Bacillus

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2 Mechanisms of Cr(VI) Reduction by Microorganisms

Fig. 2.39 Factors affecting the Cr(VI)-reducing activity of NitR. Temperature and pH (a) and tolerance of NitR to temperature (b). Reduction rate of Cr(VI) by NitR with different initial Cr(VI) concentrations and sufficient NADPH (c) and kinetics of Cr(VI) reduction by NitR (d). The values are presented as the mean ± S.D. (standard deviation) (n = 3), the error bar is the S.D. Different lowercase letters above bar charts indicate significant differences (p < 0.05) between members in the group. Reprinted from Ref. [69], Copyright 2018, with permission from Elsevier

sp. ES 29 is comparable to that of Pannonibacter phragmitetus BB, and the crude enzymatic extract’s V max and K m are only 0.171 μmol/min/mg and 7.09 μM for the cell-free extract [82]. This finding implied that NitR’s V max was superior to those of the recognized classic chromate reductases and crude enzymatic extracts, despite the fact that its K m was lower. As a result, NitR is more responsive to Cr(VI) and reduces Cr(VI) more quickly. This explains the effective reduction of Cr(VI) by Pannonibacter phragmitetus BB. In contrast to ChrR and YieF, two well-known class I chromate reductases, NitR has a higher proportion of identity and conserved and semiconserved alterations in the amino acid sequences, as seen in Fig. 2.40 [71]. The homology and specificity of the chromate reductases for reducing Cr(VI) also serve as a basis for classification. Class I chromate reductases either transfer two electrons simultaneously from themselves (known as “tight” enzymes like YieF) or nonsimultaneously (known as “semitight” enzymes like ChrR) [28, 29, 81]. NfsA and other class II chromate reductases have no similarity to class I enzymes. Poor homology also exists between NitR and NfsA.

2.5 Molecular Mechanisms of Bacterial Reduction of Cr(VI)

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The Class I chromate reductases are thought to be more efficient than other chromate reductases and the alignment demonstrates that NitR is one of them. The FMN and NADPH binding sites are essential for chromate reductases to work. FMN delivers the electron from NADH to the substrate scilicet Cr(VI) while being firmly fixed to the reductases via a number of hydrogen bonds. As a result, the enzyme’s FMN-binding region serves as the active site, and hydrogen bonding between FMN molecules and other monomers may be crucial to the activity of chromate reductase. The NADPH-binding site is crucial for enzyme function because NAD(P)H supplies electrons for the reduction of Cr(VI). Compared to ChrR and YieF, which have identical or conserved substitutions, NitR’s FMN- and NADPH-binding sites are more conserved (Fig. 2.40). NitR may therefore be a NADPH-dependent FMN-reductase that reduces Cr(VI) using the same electron transport mechanism as ChrR and YieF.

Fig. 2.40 Multiple alignment of NitR, nitrite reductase (NitK) and the classic chromate reductases (ChrR and YieF). Symbols indicate identical residues (*), conserved substitutions (:), and semiconserved substitutions (.). The FMN-binding and NADPH-binding sites are labeled with yellow and cyan rectangles, respectively, and the sites marked with red rectangles are both FMN and NADPH binding sites. The black dash indicates one gap/deletion in the protein sequences by alignment. Reprinted from Ref. [69], Copyright 2018, with permission from Elsevier

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2.5.3 Dynamic Proteome Responses to Sequential Reduction of Cr(VI) and Removal of Coexisting Heavy Metals by Pannonibacter Phragmitetus BB Heavy metal contamination at chromite ore processing residue disposal locations, including levels of Hg, Cd, Zn, and Pb that were higher than the background levels of soil in China, has been severe [83]. It is crucial to look into how different heavy metal cations affect Pannonibacter phragmitetus BB’s ability to reduce Cr(VI). Pb(II) is bioprecipitated by phosphates [84], adsorbed by extracellular polymeric substances (EPS) (1587.3 mg Pb(II)/g dry weight), and intracellularly bioaccumulated by metallothioneins generated by microorganisms (26.5 mg Pb(II)/g dry weight of cells) [85]. The efficient reduction of Cr(VI) to Cr(III) by Pannonibacter phragmitetus BB has been observed. Although many research have concentrated on eliminating a specific metal, it is extremely difficult to simultaneously treat coexisting metals in the environment using a single living microbial strain. Pannonibacter phragmitetus BB achieved to firstly decrease Cr(VI) and then adsorb Pb(II). Also, the biotransformation of the functional groups during the process and the microbial dynamic proteome responses were examined.

2.5.3.1

Bacterial Growth and Bioremoval of Pb(II) and Cr(VI)

The growth of Pannonibacter phragmitetus BB was not significantly affected by 50 mg/L Pb(II) (Fig. 2.41a), but was delayed in the presence of 100 mg/L Cr(VI). However, the final OD600 values of cells in all treated groups and the control reached 2.2, indicating that the strain exhibited relatively high resistance to the pressure of both Cr(VI) and Pb(II). Cr(VI) was completely reduced within 24 h, with rapid reduction occurring between the 12th and 24th hours. The additional pressure of Pb(II) did not affect the reduction of Cr(VI) (Fig. 2.41b). This was followed by rapid removal stages of Pb(II), with removal efficiencies of 50.89 and 52.15% in the single and binary systems containing Pb(II) during the 24–48 h period, respectively (Fig. 2.41c). The pH values of the cultured medium solutions rapidly increased to 8.4 after 12 h, which may have contributed to the logarithmic reduction of Cr(VI) and removal of Pb(II) (Fig. 2.41d). Previous studies have suggested that Pannonibacter phragmitetus BB, isolated from a strongly alkaline chromite ore processing residue (COPR) disposal site, may be able to produce alkali to adjust the pH values of the environment, which benefits its physiological responses to metal pressure [86]. The functional groups involved in the bioremoval of Pb(II) and Cr(VI) were not explicitly mentioned in the text.

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Fig. 2.41 Growth curves (a), Cr(VI) concentration (b) and Pb(II) removal rates (c) and pH variation of the culture medium (d) in the single system of Cr(VI) or Pb(II), binary system of Cr(VI) and Pb(II), and controls (Pannonibacter phragmitetus BB or sterile media with Pb(II) and Cr(VI). Reprinted from Ref. [87] Copyright 2019, with permission from Elsevier

2.5.3.2

Functional Groups Involved in the Bioremoval of Pb(II) and Cr(VI)

Under the pressure of Cr(VI) and Pb(II), the morphology of Pannonibacter phragmitetus BB became stouter compared to the control cells, due to the toxicity of the two metals (Fig. 2.42a, b). According to the results, Cr(VI) was entirely reduced to Cr(III), with binding energies of 576.91 and 586.04 eV for the single Cr(VI) system and 576.92 as well as 586.08 eV for the binary Cr(VI) and Pb(II) system, respectively (Fig. 2.43a) [88]. Nevertheless, the FT-IR spectra revealed that important functional groups of the bacterial cell surface, such as amide (C=O (1654 cm−1 ), C = N (1401 cm−1 ), hydroxyl (3292 cm−1 ), and amine (1540 cm−1 ), were not significantly altered by Cr(VI) (Fig. 2.42e, f) [89]. Similar findings were made by the EDS spectra, which showed that the cell surface only contained a modest quantity of chromium. More research on intracellular protein reactions to Cr(VI) and Cr(III) is needed and Cr(VI) reduction did not depend on the functional groups of the cell surface. Lead was still present on the cell surface in greater amounts than chromium (Fig. 2.42c). The XPS spectra also showed the peaks of Pb-OH (137.12 and 137.25 eV; Pb 4f7/2 ) and Pb–O (142.10 and 142.12 eV; Pb 4f5/2 ). The outcomes showed that Pb(II) ions were successfully adsorbed on the cell surfaces (Fig. 2.43b) [90]. The O 1s deconvolution spectra revealed that the binding energy of O–H was

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Fig. 2.42 SEM photos of the pristine Pannonibacter phragmitetus BB (a) and after sequential reduction of Cr(VI) and adsorption of Pb(II) (b); EDS spectrum of the strain BB with Cr(VI) and Pb(II) (c); FT-IR spectra of the strain in the single system of Cr(VI) (d) or Pb(II) (e) and binary system of Cr(VI) and Pb(II) (f) in different stages. OP: origin point; RG: rapid growth period; PF: platform period; RR: rapid Cr(VI) reduction period; RA: rapid Pb(II) adsorption period. Reprinted from Ref. [87] Copyright 2019, with permission from Elsevier

Fig. 2.43 XPS spectra of Cr 2p (a), Pb 4f (b), C (c), O (d) and P (e) in the single system of Cr(VI) or Pb(II), binary system of Cr(VI) and Pb(II), and control (Pannonibacter phragmitetus BB) after the sequential reduction of Cr(VI) and adsorption of Pb(II). Reprinted from Ref. [87] Copyright 2019, with permission from Elsevier

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shifted to 532.60 eV following Pb(II) adsorption and its proportion reduced from 50.72% to 34.55% (for single systems of Pb(II)) or to 41.52% (for binary systems of Cr(VI) and Pb(II)) [91]. It was thought that the presence of Pb(II) and alkaline ions most likely caused a decrease in O–H concentrations [91, 92]. The binding energy of O=P(OR)3 changed from 133.58 to 133.22 eV, and its percentage increased from 36.48%-73.17% (single system of Pb(II)) to 70.24% (binary system of Cr(VI) and Pb(II)), according to the results of the P 2p deconvolution spectrum (Fig. 2.43e) [91]. The results were accompanied by the development of the C–O stretching vibration of phosphodiester (980 cm−1 ) and the disappearance of free phosphate (1236 cm−1 ) in the FT-IR spectra during the quick Pb(II) adsorption period (Fig. 2.42e, f) [91, 93]. This result suggested that the formation of Pbx (PO4 )y by the phosphate group may be involved in the precipitation of Pb(II) [94]. After Pb(II) adsorption, a new peak of O–C=O (288.53 eV) was seen in the C 1s deconvolution spectra (Fig. 2.43c) [95, 96]. Similarly, during the quick Pb(II) adsorption period, the vibration of the C=O ester (1726 cm−1 ) emerged, and dramatic variations due to C–O–C stretching vibrations from the ester (1100–1280 cm−1 ) were seen (Fig. 2.42e, f) [93]. Therefore, the functional groups including carboxyl, hydroxyl, phosphonate, ester, and phosphodiester may be crucial for Pannonibacter phragmitetus BB’s ability to remove Pb(II) from the environment [97, 98].

2.5.3.3

Bacterial Comparative Proteomics in the Single and Binary Metal Systems

In single and binary metal systems, bacterial comparative proteomics was used to investigate the molecular processes of Pannonibacter phragmitetus BB’s sequential reduction of Cr(VI) and adsorption of Pb(II). An iTRAQ-based quantitative analysis was used to study the overview of comparative proteomics, and a total of 3822 linked proteins were found. According to the COG functional categorization, the differentially expressed proteins (DEPs) were divided into 22 categories by their activities to extract information regarding the interaction between the BB strain and Pb(II) or Cr(VI) (Fig. 2.44a–d). For the most part, DEPs in the transcription (K), translation, ribosomal structure, and biogenesis (J), and inorganic ion transport and metabolism (P) groups were responsible for the fast reduction of Cr(VI). Nevertheless, amino acid transport and metabolism (E) and carbohydrate transport and metabolism (G) groups were implicated in the primary DEPs associated to the adsorption of Pb(II) (Fig. 2.44c, d). According to the findings, Pannonibacter phragmitetus BB’s unique proteins and regular networks were linked to the ability to tolerate stress by Cr(VI) or Pb(II), respectively [99]. (1) Proteins involved in Cr(VI) reduction In the single and binary Cr(VI) systems, Pannonibacter phragmitetus BB had high Cr(VI)-reducing abilities (Fig. 2.41b). In general, upregulated proteins are induced to account for bacterial resistance to contaminants, whereas downregulated proteins are weakened by contaminant toxicity. In the comparison of the Cr-RR and RG groups

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Fig. 2.44 COG distribution of differentially expressed proteins (DEPs) in the comparison groups Cr-RR vs RG (a), Pb + Cr-RR vs RG (b), Pb-RA versus PF (c) and Pb + Cr-RA versus PF (d); Venn diagrams of the up- (e, g) and downregulated (f, h) expressed proteins in the comparison groups. Reprinted from Ref. [87] Copyright 2019, with permission from Elsevier

2.5 Molecular Mechanisms of Bacterial Reduction of Cr(VI)

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(Cr-RR VS RG), 369 upregulated and 410 downregulated DEPs were found in total, while 409 upregulated and 394 downregulated DEPs were found in the comparison of the Pb+Cr-RR and RG groups. In all of the comparison groups, 289 downregulated DEPs (Fig. 2.44f) and 293 shared upregulated DEPs (Fig. 2.44e) were found to be differentially expressed. Table 2.15 and Fig. 2.45 present the DEPs that are involved in the reduction of Cr(VI). The first category comprises intracellular Cr(VI) reductases, including oxidoreductases (WP_123194584.1, KND18397.1, WP_094461821.1, and KND17003.1), nitrite reductases (ALV30032.1), and azoreductase (WP_ 094463232.1) [28, 100]. These enzymes have been found to play a key role in the reduction of intracellular Cr(VI) using NADH and NADPH as electron donors, which are produced through bacterial energy metabolism (as shown in Fig. 2.45). Their expression was significantly upregulated (1.249–2.907 times) in the presence of Cr(VI) in both single and binary systems (as listed in Table 2.15). According to Table 2.15 and Fig. 2.45, cytochromes and hydrogenases, which function in the electron transport system with Cr(VI) as a terminal electron acceptor, fall under the DEP group that corresponds to the strain BB’s extracellular reduction of Cr(VI) [7, 28, 101]. The rapid Cr(VI)-reduction phase saw an upregulation of several cytochromes and dehydrogenases. For instance, in the single system with Cr(VI) and the binary system with Cr(VI) and Pb(II), respectively, the expression levels of cytochrome c and aldehyde dehydrogenase (WP 094462901.1) increased by 1.472 and 1.397 and 1.467 and 1.478 times, respectively. Furthermore, transporters are the third category of pathways that respond to both intracellularly and extracellularly transported Cr(VI)/ Cr(III) (as shown in Fig. 2.45). During rapid Cr(VI)-reduction periods, the average expression levels of three sulfate transporters (ALV28416.1, WP_094461934.1, and WP_094461933.1) were increased 1-fold. This suggested that Cr(VI) may enter bacterial cells through the sulfate transport system, as chromate (CrO4 2− ) has a similar structure to sulfate (SO4 2− ) (as illustrated in Fig. 2.45) [102]. Increased Cr(VI) inside the cell led to more Cr(VI) being released, as the chromate efflux transporter (ChrA) was stimulated to express 3.116- and 2.885-fold in the single and binary systems with Cr(VI), respectively. Intracellular accumulation of Cr(III) induced overexpression of the iron ABC transporters (WP_026355339.1, WP_094463595.1, WP_094462483.1 and WP_094462377.1) and they may be involved in efflux of Cr(III) due to its oxidative stress. (Table 2.15 and Fig. 2.45) [103]. Furthermore, the other categories are linked to antioxidant functions. Intracellular reduction of Cr(VI), Cr(VI), Cr(III), and reactive oxygen species (ROS) can trigger oxidative stress, resulting in harm to the DNA and proteins of bacterial cells. In order to reduce oxidative stress, the expression of proteins with antioxidant properties—such as glutathione S-transferase (GST), superoxide dismutase (SOD), and sulfurtransferases—was markedly elevated (Table 2.15). DNA recombination protein RmuC (WP_123193516.1), one of the proteins in the bacterial DNA repair system, protected cells from the damage by Cr(VI) and Cr(III) [102]. Thus, numerous Cr(VI) reduction pathways, effective influx and efflux channels of Cr(VI) and Cr(III), oxidation resistance, and DNA repair mechanisms all contribute to the strain BB’s powerful capacities to reduce and resist Cr(VI) (Fig. 2.45).

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Table 2.15 The proteins of strain BB involved in the tolerance and reduction of Cr(VI) Categories

Protein_ID

Function-description

Code Fold change (COG) Cr-RR Pb_ VS Cr-RR RG VS RG

Intracellular reductases

WP_ 123194584.1

FAD-dependent oxidoreductase

R

2.907

2.55

KND18397.1

NADPH-dependent oxidoreductase C

1.425

1.401

WP_ 094461821.1

NAD(P)H-dependent oxidoreductase

1.249

1.253

ALV30032.1

Nitrite reductase, copper-containing DPM

1.643

1.913

WP_ 122520798.1

Nitrate reductase subunit beta

CP

–

1.69

WP_ 094463054.1

Nitrate reductase molybdenum cofactor assembly chaperone

CPO

1.251

3.844

WP_ 094463055.1

Nitrate reductase subunit alpha

CP

–

1.719

WP_ 094463232.1

FMN-dependent NADH-azoreductase

C

1.963

1.561

KND17003.1

SDR family NAD(P)-dependent oxidoreductase

M

1.267

1.352

ALV30022.1

SDR family oxidoreductase

IQR

2.853

2.334

ALV29613.1

SDR family oxidoreductase

R

1.206

1.237

SUB03166.1

Predicted unusual protein kinase HT regulating ubiquinone biosynthesis, AarF/ABC1/UbiB family

1.232

1.231

Cytochrome C

C

1.472

1.467

Cytochrome c oxidase subunit II

E

1.367

1.253

SUB00345.1

Cytochrome d terminal oxidase subunit I

C

–

1.449

WP_ 094462901.1

Aldehyde dehydrogenase

C

1.397

1.478

ALV28703.1

Aldehyde dehydrogenase

C

1.254

1.309

WP_ 094462030.1

PQQ-dependent sugar dehydrogenase

G

1.428

1.411

KND16436.1

Predicted oxidoreductase (related to aryl-alcohol dehydrogenase)

R

1.425

1.409

WP_ 094462520.1

NAD-glutamate dehydrogenase

E

1.359

1.341

WP_ 094461834.1

Homoserine dehydrogenase

E

1.289

1.387

Cytochromes WP_ and 094462148.1 dehydrogenases KND20305.1

C

(continued)

2.5 Molecular Mechanisms of Bacterial Reduction of Cr(VI)

111

Table 2.15 (continued) Categories

Protein_ID

Function-description

Code Fold change (COG) Cr-RR Pb_ VS Cr-RR RG VS RG

WP_ 123193734.1

Acyl-CoA dehydrogenase

I

1.288

1.443

WP_ 094462237.1

3-hydroxyacyl-CoA dehydrogenase I

1.254

1.340

WP_ 094461787.1

Xanthine dehydrogenase molybdopterin binding subunit

F

1.330

1.267

WP_ 094461788.1

Xanthine dehydrogenase small subunit

F

1.244

1.304

SUB02547.1

Type I glyceraldehyde-3-phosphate G dehydrogenase

1.222

1.201

WP_ 123194688.1

Glucose-6-phosphate dehydrogenase

G

1.217

1.295

WP_ 123195680.1

Saccharopine dehydrogenase family protein

E

1.218

1.214

KND21113.1

UDP-glucose/GDP-mannose dehydrogenase family protein

M

1.317

1.316

WP_ 094461795.1

4-hydroxythreonine-4-phosphate dehydrogenase PdxA

H

1.454

1.445

Sulfate ABC transporter permease subunit CysT

p

2.687

3.116

WP_ 094461934.1

Sulfate ABC transporter substrate-binding protein

P

1.779

1.935

WP_ 094461933.1

Sulfate/molybdate ABC transporter P ATP-binding protein

1.382

1.723

WP_ 119396589.1

Chromate efflux transporter

P

3.116

2.885

KND18 421.1 ABC transporter substrate-binding protein

P

2.380

2.192

WP_ 122519206.1

ABC transporter substrate-binding protein

P

2.276

2.361

WP_ 026355339.1

Iron ABC transporter substrate-binding protein

P

2.844

4.74

WP_ 094463595.1

Iron ABC transporter substrate-binding protein

P

1.968

1.733

WP_ 094462483.1

Iron ABC transporter substrate-binding protein

P

1.623

2.098

WP_ 094462377.1

Fe(3+) ABC transporter substrate-binding protein

P

1.426

1.470

Ion transporters ALV28416.1

(continued)

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2 Mechanisms of Cr(VI) Reduction by Microorganisms

Table 2.15 (continued) Categories

Protein_ID

Function-description

Code Fold change (COG) Cr-RR Pb_ VS Cr-RR RG VS RG

Superoxide dismutase

SUB00261.1

MULTISPECIES: superoxide dismutase

P

1.256

–

Glutathione transferases

WP_ 094463327.1

Sulfurtransferase

S

2.72

2.706

WP_ 094462468.1

Sulfurtransferase

S

2.831

3.192

ALV25717.1

Glutathione S-transferase

O

1.723

1.674

KND20786.1

Glutathione S-transferase family protein

O

1.392

1.405

ALV29495.1

Glutathione-disulfide reductase

C

1.543

1.503

WP_ 094463372.1

Glutathione import ATP-binding protein

Q

1.248

1.32

WP_ 122519374.1

Glutamate–cysteine ligase

H

1.369

1.302

KND17998.1

Catechol 2,3-dioxygenase or other lactoylglutathione lyase family enzyme

Q

1.327

1.377

WP_ 122519650.1

Catechol 2,3-dioxygenase or other lactoylglutathione lyase family enzyme

Q

1.524

1.499

SUA99239.1

Class 1b ribonucleoside-diphosphate reductase subunit beta

F

2.137

1.888

WP_ 123193516.1

DNA recombination protein RmuC L

1.745

1.81

WP_ 094461991.1

Recombinase RecA

L

1.692

1.669

KND19751.1

Holliday junction branch migration L DNA helicase RuvB

1.273

1.236

KND16770.1

Excinuclease ABC subunit UvrB

L

1.372

1.418

ALV27274.1

Crossover junction endodeoxyribonuclease RuvC

L

1.501

1.413

WP_ 094462046.1

Na+ /H+ antiporter

O

1.238

1.289

DNA repair proteins

Antiporters

(continued)

2.5 Molecular Mechanisms of Bacterial Reduction of Cr(VI)

113

Table 2.15 (continued) Categories

Protein_ID

Function-description

Code Fold change (COG) Cr-RR Pb_ VS Cr-RR RG VS RG

ALV29663.1

Sodium: proton antiporter

T

1.218

1.213

COG categories: Energy production and conversion (C); Cell cycle control, cell division, chromosome partitioning (D); Amino acid transport and metabolism (E); Nucleotide transport and metabolism (F); Carbohydrate transport and metabolism (G); Coenzyme transport and metabolism (H); Lipid transport and metabolism (I); Transcription (K); Replication, recombination and repair (L); Cell wall/membrane/envelope biogenesis (M); Posttranslational modification, protein turnover, chaperones (O); Inorganic ion transport and metabolism (P); Secondary metabolite biosynthesis, transport and catabolism (Q); General function prediction only (R); Function unknown (S); Signal transduction mechanisms (T); Defense mechanisms (V). Coenzyme transport and metabolism; Signal transduction mechanisms (HT); Energy production and conversion; Inorganic ion transport and metabolism (CP); Cell cycle control, cell division, chromosome partitioning; Inorganic ion transport and metabolism; Cell wall/membrane/envelope biogenesis (DPM); Energy production and conversion; Inorganic ion transport and metabolism; Posttranslational modification, protein turnover, chaperones (CPO); Lipid transport and metabolism; Secondary metabolite biosynthesis, transport and catabolism; General function prediction only (IQR); Cell wall/membrane/envelope biogenesis; Defense mechanisms (MV) Reprinted from Ref. [87] Copyright 2019, with permission from Elsevier

(2) Proteins related to Pb(II) adsorption and detoxification Moreover, it has been demonstrated that Pannonibacter phragmitetus BB removes Pb(II) efficiently (Fig. 2.41c). By contrasting the groups of Pb-RA or Pb+Cr-RA with PF, respectively, the proteome responses for adsorbing and detoxifying Pb(II) were further uncovered. The comparison group of Pb-RA VS PF indicated 510 upregulated and 360 downregulated DEPs, while Pb+Cr-RA VS PF revealed 425 upregulated and 375 downregulated DEPs. Furthermore, these groups shared 173 upregulated (Fig. 2.44g) and 127 downregulated (Fig. 2.44h) DEPs. The DEPs (differentially expressed proteins) that participate in the adsorption of Pb(II) are presented in Table 2.16 and Fig. 2.45. Pb(II) adsorption has been linked to metal-induced polyphosphate degradation and the consequent efflux of metalphosphate [104]. A 1.51-fold increase in the expression level of the transcriptional regulator of phosphonate metabolism, PhnF (KND17366.1) was observed, which in turn stimulated different cellular responses of phosphate ABC transporters. The phosphate ABC transporters, SUB00018.1 and KND19286.1 are responsible for facilitating the efflux of phosphate to induce the precipitation of Pb(II) on the cell surface and the formation of Pb–O (142.10 and 142.12 eV), as demonstrated by the XPS spectra (Fig. 2.43b). Moreover, proteins related to lipids and exopolysaccharides (EPS) have been found to aid in bacterial adsorption of Pb(II) [105]. The

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Fig. 2.45 Mechanisms of Cr(VI) reduction and Pb(II) adsorption by Pannonibacter phragmitetus BB. (1) Intracellular and extracellular Cr(VI) reduction system; (2) Cr(VI) and Cr(III) transportation system; (3) antioxidant and DNA repair systems; (4) cell surface adsorption system of Pb(II), including lipoproteins and sugar-related proteins related to EPS and precipitation by phosphate; (5) Pb(II) transport system; (6) extracellular complexation of Pb(II) with siderophores; (7) anti-transport system of Na+ /H+ . Reprinted from Ref. [87] Copyright 2019, with permission from Elsevier

lipoprotein LipA (WP_082205773.1) located in the outer membrane of the cell acted as an anchoring site for Pb(II) and was upregulated 2.074- and 1.777-fold in the single and binary systems with Pb(II), respectively (Table 2.16 and Fig. 2.45) [106]. The –COO– and –OH functional groups in the lipoprotein have been identified as potential sites for metal ion complexation [107]. Furthermore, proteins involved in sugar metabolism helped EPS attach to Pb(II) on the cell surface. The expression levels of sugar phosphate isomerase (WP_094463772.1) to isomerize phosphosugar and monosaccharides increased 3.009- and 1.86-fold in the single and binary systems with Pb(II), respectively. Similar to this, sugar ABC transporters (ALV30000.1, WP_ 094462776.1, and ALV28671.1) considerably increased (from 2.766 to 5.143 times) in the two Pb(II) treatment systems [104]. The findings showed that Pb(II) precipitation and binding may be significantly influenced by phosphate, lipoprotein, and EPS. Additionally, in the Pb(II) rapid adsorption phases, the Na+ /H+ antiporter (WP_ 094462046.1) was also upregulated, increasing cellular proton uptake and producing an extracellular alkaline microenvironment that could help strain BB precipitate and adsorb Pb(II) [108, 109].

2.5 Molecular Mechanisms of Bacterial Reduction of Cr(VI)

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Table 2.16 The proteins of strain BB involved in the detoxification and adsorption of Pb(II) Categories

Protein_ID

Function-description

Code Fold change (COG) Pb-RA Pb_ VS PF Cr-RA VS PF

Phosphate metabolism and transport

ALV28088.1

Phosphodiesterase

T

1.377

1.327

KND19286.1 Phosphate ABC transporter substrate-binding protein

P

1.565

–

KND17366.1 Phosphonate metabolism transcriptional regulator PhnF

K

1.51

0.809

SUB00018.1

P

–

10

0.423

0.661

Phosphate ABC transporter permease subunit PstC

WP_ ABC-type phosphate transport P 019963025.1 system, periplasmic component Lipoprotein-related WP_ Lipoprotein LipA proteins 082205773.1

M

2.074

1.777

ALV29382.1

M

1.22

1.551

WP_ Lipoprotein-anchoring 094463859.1 transpeptidase ErfK/SrfK

M

1.968

3.474

WP_ Predicted periplasmic 094463598.1 lipoprotein

S

1.541

1.659

1.343

1.627

G

3.009

1.86

WP_ Sugar ABC transporter 094462776.1 substrate-binding protein

G

3.85

5.143

ALV30000.1

Sugar ABC transporter substrate-binding protein

G

4.148

2.509

ALV28671.1

Sugar ABC transporter substrate-binding protein

G

3.565

2.766

ALV25722.1

Co2+ /Mg2+ efflux protein ApaG P

ALV29669.1

Lipoprotein component of the BamABCDE complex

Outer membrane protein OmpA M and related peptidoglycan-associated (lipo)proteins

Sugar transport and WP_ Sugar phosphate isomerase/ metabolism 094463772.1 epimerase

Ion transporters

1.705

–

WP_ Copper chaperone PCu(A)C 094462254.1

P

1.44

–

KND21304.1 Zinc ABC transporter substrate-binding protein ZnuA

P

–

1.437

WP_ Zinc/cadmium/ 094463396.1 lead-transporting P-type ATPase

P

–

1.328

WP_ Manganese-binding lipoprotein 119396709.1 MntA

P

–

1.334 (continued)

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Table 2.16 (continued) Categories

Protein_ID

Function-description

Siderophore

WP_ Siderophore-interacting protein 094462672.1

Code Fold change (COG) Pb-RA Pb_ VS PF Cr-RA VS PF 2.256

2.775

WP_ Periplasmic protein TonB, links M 123194343.1 inner and outer membranes

2.242

2.899

WP_ Siderophore synthetase 094462677.1 component

P

1.652

2.638

WP_ Siderophore synthetase 094462675.1 component

P

1.512

2.313

WP_ Efflux RND transporter 122520134.1 periplasmic adaptor subunit

MV

1.341

1.4

Heat shock proteins KND20743.1 Heat shock protein HslJ

P

O

1.557

1.628

ALV26592.1

Heat-shock protein Hsp20

O

1.487

1.311

ALV27224.1

Molecular chaperone IbpA, HSP20 family

O

1.269

1.267

SUA99928.1

Peroxiredoxin

O

1.383

1.288

WP_ Thiol oxidoreductase 094463599.1

R

1.463

1.442

WP_ Thioredoxin-dependent thiol 094461857.1 peroxidase

O

2.016

-

KND21409.1 Antibiotic biosynthesis monooxygenase

H

3.533

3.541

Divalent-cation tolerance protein

WP_ Divalent-cation tolerance 094461987.1 protein CutA

P

1.302

1.252

Na+ /H+ antiporter

WP_ Na+ /H+ antiporter 094462046.1

O

1.203

1.301

WP_ Sodium-translocating 123193986.1 pyrophosphatase

C

2.614

1.694

Oxidative stress proteins

COG categories are shown in Table 2.15 Reprinted from Ref. [87] Copyright 2019, with permission from Elsevier

Due to their potential roles in the influx or efflux of Pb(II), metal cation pumps such as the Co(II)/Mg(II) efflux protein ApaG (ALV25722.1) and the zinc ABC transporter substrate-binding protein ZnuA (KND21304.1) were also activated by Pb(II) (Table 2.16 and Fig. 2.45) [110]. Additionally, in response to intracellular Pb(II)induced oxidative stress, the expression levels of heat-shock proteins (HSPs), peroxiredoxin, and thiol oxidoreductase were elevated 1.267–1.628 times (Table 2.16)

2.5 Molecular Mechanisms of Bacterial Reduction of Cr(VI)

117

[104]. Additionally, CutA, a protein that tolerates divalent cations, was increased in response to Pb(II)’s harmful effects [111]. Previous research also demonstrated that siderophores were delivered extracellularly by the protein RND to form stable complexes with metals like Cd(II), Pb(II), and Zn(II), which can decrease the mobility of metal ions and lessen their toxicity to cells [105, 112]. The transmission of energy in this method required the TonB system [105, 112]. Here, all of the siderophore synthetase components (WP_094462677.1 and WP_094462675.1) as well as the periplasmic protein TonB (WP_123194343.1), efflux RND transporter periplasmic adaptor subunit (WP_122520134.1), and the single Pb(II) system were significantly upregulated. Overall, the strain BB eliminates Pb(II) via adsorption with proteins associated to sugar and lipoproteins as well as precipitation with phosphate and OH. In order to relieve Pb(II) toxicity, the strain BB additionally makes use of several Pb(II) transporters, antioxidant proteins, and extracellular complexation with siderophores.

2.5.3.4

Bacterial Dynamic Protein Networks for Bioremoval of Cr(VI) and Pb(II)

Understanding the functional and evolutionary characteristics of proteins at the system level is made possible by protein–protein interaction (PPI) networks. Additionally, they offer perceptions into certain interaction partners, which can help to focus on proteins’ functions [113]. At three different phases (Pb + Cr-RG, Pb + Cr-RR, and Pb + Cr-RA), the dynamic networks of differentially expressed proteins (DEPs) in the combined Cr(VI) and Pb(II) system were examined. The results are shown in Fig. 2.46. Using the STRING database, a network structure comprising 80 nodes (proteins) and 166 edges (interactions) was created among the DEPs. Of these, Pb(II) adsorption (red), Cr(VI) reduction (green), and pH variation (orange) were each carried out by 29, 48, and 3 DEPs, respectively. Superoxide dismutase (SOD) (SUB00261.1) and type I glyceraldehyde-3-phosphate dehydrogenase (SUB02547.1) had the highest levels of connection in the network. In order to resist and detoxify Cr(VI) and Pb(II), the antioxidative stress system was crucial. The entire bioremoval process resulted in an upregulation of its DEPs’ expression patterns (Table 2.17). One of the key proteins in the PPI network, SOD interacted with 23 proteins directly, including cytochromes (3), dehydrogenases (5), oxidoreductases (2), and Fe(III) ABC transporter (1), as well as Pb(II) adsorption and transport (2), siderophore release (2), oxidative stress defense-related proteins (6), and DNA repair (2) (Fig. 2.46). The strain BB was used to resist and detoxify Cr(VI) and Pb (II) using two separate methods. SOD may have induced cytochrome

118

2 Mechanisms of Cr(VI) Reduction by Microorganisms

Fig. 2.46 The protein–protein interaction network involved in Cr(VI) reduction, Pb(II) adsorption, and detoxification of the two heavy metals (red nodes represent the proteins involved in adsorption and detoxification of Pb(II); green nodes represent the proteins involved in tolerance and reduction of Cr(VI); orange nodes represent the proteins involved in pH variation; the histograms in the node represent the expression of proteins at different time periods: the rapid growth (blue), rapid Cr(VI) reduction (violet), and rapid Pb(II) adsorption (red) stages, respectively. Reprinted from Ref. [87] Copyright 2019, with permission from Elsevier

c (WP 094462148.1), aldehyde dehydrogenase (WP 094462901.1), and NAD(P)Hdependent oxidoreductase (WP 094461821.1) to reduce intracellular and extracellular Cr(VI) [114]. However, the major methods of reducing the oxidative stress by Pb(II) were adsorption, transportation, and complexation of Pb(II) caused by SOD in strain BB. The amount of cellular oxidative stress by Pb(II) was lessened by the metal’s efflux from cells via the transport protein MntA (WP 119396709.1) [110, 115]. A barrier for Pb(II) was the functional groups of the cell surface that adsorbs Pb(II) [84]. Furthermore, Pb(II) was prevented from entering the cells by forming complexes with siderophores whose releasing proteins (WP 123194343.1 and WP 094462672.1) were highly interconnected with SOD in the network [105]. Also, to lessen oxidative stress damage by both Cr(VI) and Pb(II), SOD directly controls 2 DNA repair proteins (KND16770.1 and WP 094461991.1), 6 downstream oxidative stress defense-related proteins (including peroxiredoxin and Hsp20), and 6 other

2.5 Molecular Mechanisms of Bacterial Reduction of Cr(VI)

119

proteins [116]. These proteins’ dynamic expression profiles were synchronistically linked to SOD (Table 2.17 and Fig. 2.46). Major DEPs connected to Cr(VI) reduction were developed a Cr(VI)-dependent PPI network during the phases of rapid growth (RG) and Cr(VI) reduction (RR), but their expression was decreased during the phase of fast Pb(II) adsorption (RA) (Table 2.17 and Fig. 2.46). A synchronously dynamic subnetwork was created by the direct or indirect connections between the oxidoreductases (WP_123194584.1, KND18397.1, and WP_094463232.1) and the sulfate ABC transporters (WP_ 094461934.1, ALV28416.1, and WP_094461933.1) (Fig. 2.46). According to our research, the strain BB coordinated Cr(VI) transport and intracellular Cr(VI) reduction in response to the culture medium’s fluctuating Cr(VI) concentration, as shown in Fig. 2.41b. Additionally, a subnetwork of DEPs implicated in the extracellular adsorption of Pb(II), and these proteins were most abundantly expressed during the fast periods of Pb(II) adsorption (Fig. 2.46). The siderophore release system (WP_ 122520134.1, WP_094462672.1, WP_094462677.1, and WP_123194343.1) and the phosphonate metabolism and transport proteins (KND17366.1, KND19286.1, SUB00018.1, and WP 019,963,025.1) were connected by the outer membrane protein OmpA (ALV29669.1) (Fig. 2.46). These findings suggested that Pb(II) adsorption and cellular toxicity reduction may require cooperation between the phosphonate metabolism and transport system and siderophore release mechanism. Na+ /H+ antiporters (WP_094462046.1 and ALV29663.1) of strain BB maintained a fairly high expression level throughout the RG, RR, and RA periods because they were involved in both Cr(VI) reduction and Pb(II) adsorption (Table 2.17). Through the lipoprotein-anchoring transpeptidase ErfK/SrfK (WP_ 094463859.1) in the PPI network (Fig. 2.46), they were connected to both lipoproteins (WP 082,205,773.1 and ALV29382.1) and Cr(VI)-reducing related proteins (WP_122520798.1, ALV30032.1, WP_094463055.1, and WP_094463054.1). The Na+ /H+ antiporter, shown in Fig. 2.45, is a channel for the efflux of alkali cations like Na + and the concomitant inflow of outside protons, which can elevate the pH levels of the extracellular milieu and boost ATP generation for cellular energy metabolism [109]. NAD(P)H is produced during energy metabolism and acts as an electron donor for intracellular Cr(VI) reduction through the action of chromium reductases (Fig. 2.45; 25, 115). By using cytochrome c and dehydrogenases (like SUB02547.1) with Cr(VI) as the electron acceptor, respectively, protons and electrons produced by energy metabolism were also transported to the outside of the cell (Fig. 2.45) [114]. Lipoprotein transport and anchoring on the cell membrane benefited Na+ /H+ antiporters as well, and these processes may have caused Pb(II) to react with –OH and –COOH groups (Figs. 2.42 and 2.45) [117]. Furthermore, the extracellular microenvironment’s alkaline conditions caused Pb(II) to precipitate

120

2 Mechanisms of Cr(VI) Reduction by Microorganisms

Table 2.17 The dynamic expression profiles of the proteins involved in the sequential reduction of Cr(VI) and adsorption of Pb(II) in different periods Protein ID

Description

Fold change Pb_ Pb_ Pb_ Cr-RG-VS-OP Cr-RR-VS-OP Cr-RA-VS-OP

WP_ FAD-dependent oxidoreductase 123194584.1

3.542

2.199

1.4

KND18397.1 NADPH-dependent oxidoreductase

1.655

1.477

1.4

WP_ NAD(P)H-dependent 094461821.1 oxidoreductase

1.299

1.188

0.9

WP_ Nitrate reductase subunit beta 122520798.1

1.13

1.103

0.8

ALV30032.1 Nitrite reductase, copper-containing

0.634

1.078

0.7

WP_ Nitrate reductase molybdenum 094463054.1 cofactor assembly chaperone

1.85

1.872

0.5

WP_ Nitrate reductase subunit alpha 094463055.1

0.918

0.859

0.6

WP_ FMN-dependent 094463232.1 NADH-azoreductase

1.373

1.258

0.9

KND17003.1 SDR family NAD(P)-dependent oxidoreductase

1.037

0.992

0.8

ALV30022.1 SDR family oxidoreductase

2.804

1.442

1.3

ALV29613.1 SDR family oxidoreductase

1.427

1.254

1.5

SUB03166.1 Predicted unusual protein kinase 0.84 regulating ubiquinone biosynthesis, AarF/ABC1/UbiB family

0.793

0.6

WP_ Cytochrome C 094462148.1

1.428

1.278

1.2

KND20305.1 Cytochrome c oxidase subunit II 1.268

1.127

1.4

SUB00345.1 Cytochrome d terminal oxidase subunit I

1.104

1.119

1

WP_ Aldehyde dehydrogenase 094462901.1

1.555

1.649

1.6

ALV28703.1 Aldehyde dehydrogenase

0.747

0.784

0.9

WP_ PQQ-dependent sugar 094462030.1 dehydrogenase

1.767

1.622

2

KND16436.1 Predicted oxidoreductase (related to aryl-alcohol dehydrogenase)

1.286

1.1

0.9

WP_ NAD-glutamate dehydrogenase 094462520.1

1.24

1.036

0.9 (continued)

2.5 Molecular Mechanisms of Bacterial Reduction of Cr(VI)

121

Table 2.17 (continued) Protein ID

Description

Fold change Pb_ Pb_ Pb_ Cr-RG-VS-OP Cr-RR-VS-OP Cr-RA-VS-OP

WP_ Homoserine dehydrogenase 094461834.1

1.452

1.319

1.1

WP_ Acyl-CoA dehydrogenase 123193734.1

1.326

1.405

1.7

WP_ 3-hydroxyacyl-CoA 094462237.1 dehydrogenase

1.223

1.207

1.3

WP_ Xanthine dehydrogenase 094461787.1 molybdopterin binding subunit

1.334

1.126

1

WP_ Xanthine dehydrogenase small 094461788.1 subunit

1.28

1.12

1.1

SUB02547.1 type I glyceraldehyde-3-phosphate dehydrogenase

0.787

0.846

0.9

WP_ Glucose-6-phosphate 123194688.1 dehydrogenase

1.058

0.915

0.9

WP_ Saccharopine dehydrogenase 123195680.1 family protein

0.839

0.754

0.6

KND21113.1 UDP-glucose/GDP-mannose dehydrogenase family protein

1.184

0.983

0.8

WP_ 4-hydroxythreonine-4-phosphate 1.518 094461795.1 dehydrogenase PdxA

1.33

1.4

ALV28416.1 Sulfate ABC transporter permease subunit CysT

3.734

2.351

1

WP_ Sulfate ABC transporter 094461934.1 substrate-binding protein

3.705

2.534

2.5

WP_ Sulfate/molybdate ABC 094461933.1 transporter ATP-binding protein

2.223

1.527

1

WP_ Chromate efflux transporter 119396589.1

4.526

3.591

4.5

KND18421.1 ABC transporter substrate-binding protein

1.955

1.91

2.9

WP_ ABC transporter 122519206.1 substrate-binding protein

1.498

1.886

2.7

WP_ Iron ABC transporter 026355339.1 substrate-binding protein

1.28

1.641

1.9

WP_ Iron ABC transporter 094463595.1 substrate-binding protein

1.869

1.957

3.3 (continued)

122

2 Mechanisms of Cr(VI) Reduction by Microorganisms

Table 2.17 (continued) Protein ID

Description

Fold change Pb_ Pb_ Pb_ Cr-RG-VS-OP Cr-RR-VS-OP Cr-RA-VS-OP

WP_ Iron ABC transporter 094462483.1 substrate-binding protein

1.119

1.285

1.6

WP_ Fe(3+) ABC transporter 094462377.1 substrate-binding protein

1.248

1.405

1.9

SUB00261.1 MULTISPECIES: superoxide dismutase

1.36

1.245

1.4

WP_ Sulfurtransferase 094463327.1

2.788

2.5

2.3

WP_ Sulfurtransferase 094462468.1

3.16

2.624

4.6

ALV25717.1 Glutathione S-transferase

2.246

1.869

1.4

KND20786.1 Glutathione S-transferase family 1.177 protein

1.362

3.9

ALV29495.1 Glutathione-disulfide reductase

1.424

1.137

1.1

WP_ Glutathione import ATP-binding 0.948 094463372.1 protein

0.808

0.7

WP_ Glutamate–cysteine ligase 122519374.1

1.129

0.935

0.9

KND17998.1 Catechol 2,3-dioxygenase or other lactoylglutathione lyase family enzyme

1.198

1.546

8.5

WP_ Catechol 2,3-dioxygenase or 122519650.1 other lactoylglutathione lyase family enzyme

1.399

1.385

1.4

SUA99239.1 Class 1b ribonucleoside-diphosphate reductase subunit beta

1.897

1.217

0.8

WP_ DNA recombination protein 123193516.1 RmuC

1.737

1.477

1.7

WP_ Recombinase RecA 094461991.1

1.442

1.316

1.3

KND19751.1 Holliday junction branch migration DNA helicase RuvB

1.279

1.318

1.2

KND16770.1 Excinuclease ABC subunit UvrB 1.162

1.105

1.3

ALV27274.1 Crossover junction endodeoxyribonuclease RuvC

1.517

1.593

1.1

ALV28088.1 Phosphodiesterase

1.096

0.99

0.9 (continued)

2.5 Molecular Mechanisms of Bacterial Reduction of Cr(VI)

123

Table 2.17 (continued) Protein ID

Description

Fold change Pb_ Pb_ Pb_ Cr-RG-VS-OP Cr-RR-VS-OP Cr-RA-VS-OP

KND19286.1 Phosphate ABC transporter substrate-binding protein

1.089

1.152

2

KND17366.1 Phosphonate metabolism transcriptional regulator PhnF

0.955

1.32

1.1

SUB00018.1 Phosphate ABC transporter permease subunit PstC

10

5.487

10

WP_ ABC-type phosphate transport 019963025.1 system, periplasmic component

1.063

1.193

1.9

WP_ Lipoprotein LipA 082205773.1

0.708

0.708

2

ALV29382.1 Lipoprotein component of the BamABCDE complex

0.667

0.59

0.8

WP_ Lipoprotein-anchoring 094463859.1 transpeptidase ErfK/SrfK

1.053

1.062

3.2

WP_ Predicted periplasmic 094463598.1 lipoprotein

0.905

1.205

1.9

ALV29669.1 Outer membrane protein OmpA and related peptidoglycan-associated (lipo)proteins

1.84

1.386

1.9

WP_ Sugar phosphate isomerase/ 094463772.1 epimerase

0.913

1.224

2.6

WP_ Sugar ABC transporter 094462776.1 substrate-binding protein

0.648

1.076

6.1

ALV30000.1 Sugar ABC transporter substrate-binding protein

0.755

1.21

3.2

ALV28671.1 Sugar ABC transporter substrate-binding protein

1.093

1.335

2.9

ALV25722.1 Co2+ /Mg2+ efflux protein ApaG

0.955

1.193

1.9

WP_ Copper chaperone PCu(A)C 094462254.1

1.194

1.492

3.5

KND21304.1 Zinc ABC transporter substrate-binding protein ZnuA

0.772

1.179

1.5

WP_ Zinc/cadmium/lead-transporting 1.358 094463396.1 P-type ATPase

1.184

1

WP_ Manganese-binding lipoprotein 119396709.1 MntA

1.587

5.2

1.108

(continued)

124

2 Mechanisms of Cr(VI) Reduction by Microorganisms

Table 2.17 (continued) Protein ID

Description

Fold change Pb_ Pb_ Pb_ Cr-RG-VS-OP Cr-RR-VS-OP Cr-RA-VS-OP

WP_ Siderophore-interacting protein 094462672.1

1.08

1.468

2.9

WP_ Periplasmic protein TonB, links 123194343.1 inner and outer membranes

0.882

1.11

1.9

WP_ Siderophore synthetase 094462677.1 component

1.447

1.692

2.1

WP_ Siderophore synthetase 094462675.1 component

1.339

1.651

1.9

WP_ Efflux RND transporter 122520134.1 periplasmic adaptor subunit

0.864

0.97

1.5

KND20743.1 Heat shock protein HslJ

1.146

1.176

2.4

ALV26592.1 Heat-shock protein Hsp20

1.212

1.216

1.1

ALV27224.1 Molecular chaperone IbpA, HSP20 family

1.329

1.08

1

SUA99928.1 Peroxiredoxin

1.254

1.245

1.7

WP_ Thiol oxidoreductase 094463599.1

0.979

1.454

1.9

WP_ Thioredoxin-dependent thiol 094461857.1 peroxidase

0.806

0.778

0.7

KND21409.1 Antibiotic biosynthesis monooxygenase

1.06

1.51

3.3

WP_ Divalent-cation tolerance protein 1.116 094461987.1 CutA

1.271

1.7

WP_ Na+ /H+ antiporter NhaA 1 OS 094462046.1

1.414

1.5

1.5

ALV29663.1 Sodium:proton antiporter

1.068

0.866

0.8

WP_ Sodium-translocating 123193986.1 pyrophosphatase

1.121

1.565

3

(Pb_Cr-RG-VS-OP: rapid period of growth; Pb_Cr-RR-VS-OP: rapid reduction period of Cr(VI); Pb_ Cr-RA-VS-OP: rapid adsorption period of Pb(II)) Reprinted from Ref. [87] Copyright 2019, with permission from Elsevier

with free phosphate and OH− (Figs. 2.42e, f). Thus, the strain BB’s intracellular and extracellular Cr(VI) reduction and Pb(II) adsorption were both facilitated by Na+ / H+ antiporters.

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Chapter 3

Microbial Remediation of Chromium-Polluted Soil

Conventional techniques used for remediating Cr(VI) contaminated soils typically involve physico-chemical extraction, landfill disposal, stabilization/solidification, flushing, phytoremediation and soil washing. However, these techniques often need significant amounts of chemical reagents and energy, which can lead to secondary pollution. As a result, bioremediation has become an increasingly popular option for remediating chromium-contaminated soils due to its environmentally friendly nature [1]. This strategy involves converting Cr(VI) into Cr(III) that is less toxic and less mobile, which is then immobilized in the soil. In our prior research, we isolated a novel strain known as Pannonibacter sp., which demonstrates high Cr(VI) reduction capabilities [2]. To determine the Cr(VI) microbial reduction effectiveness, optimize remediation conditions, and assess stability, evaluating the remediation effectiveness of Cr(VI) in soils is of utmost importance.

3.1 The Microbial Remediation Efficacy of Soil Contaminated with Chromium Utilizing microbial methods for the remediation of chromium-contaminated soil, a bioreactor (Model HZ–9212S) was utilized and placed in a bath full of water with a constant temperature of 25 °C, as depicted in Fig. 3.1. The soil sample was obtained from a disposal site for chromite ore processing residue (COPR) located in Hunan Province, China, and consisted of 0–50 cm soil sample. The bacteria Pannonibacter sp. existed in the original soil samples.

© The Author(s), under exclusive license to Springer Nature Singapore Pte Ltd. 2023 W. Yang et al., Remediation of Chromium-Contaminated Soil: Theory and Practice, Environmental Science and Engineering, https://doi.org/10.1007/978-981-99-5463-6_3

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Fig. 3.1 Schematic diagram of the bioreactor. Reprinted from Ref. [3], Copyright 2019, with permission from the Nonferrous Metals Society of China and Elsevier

3.1.1 Microbial Remediation of Cr(VI) in Soil Figure 3.2 showed that four different treatments were investigated to determine the Cr(VI) reduction capabilities of indigenous microorganisms. When culture medium was replaced by deionized water, little change in Cr(VI) concentration was observed in Treatments II and IV. However, the addition of the culture medium to the soil samples resulted in a significant promotion of Cr(VI) reduction in the presence of microorganisms (Treatment III, non-sterilized sample). It is worth noting that Cr(VI) was rapidly decreased after 18 h, with the concentration dropping from the initial 1521.9 to 199.2 mg kg−1 at 66 h. These results suggested that the presence of the culture medium was necessary to stimulate Cr(VI) reduction by microorganisms. In Treatment I, after sterilization, the concentration of Cr(VI) decreased from 1370.3 to 1008.9 mg kg−1 within 66 h, indicating that the culture medium played a dominant role in the reduction of Cr(VI). Figure 3.3 demonstrates that the microorganisms in the original soil sample can be classified into 10 genera. The dominant genus was Exiguobacterium sp., accounting for 42% of the microbial community, followed by Delftia sp. at 19.33%. Additionally, a significant proportion of microorganisms were classified as uncertain, representing 16.67% of the total. Pannonibacter sp. was found to constitute 2.67% of the microorganisms present in the original soil samples. By providing exogenous nutrients, promoting the growth and reproduction of microorganisms, it is possible to significantly enhance the synergy between microorganisms and soil components, thereby achieving efficient reduction of Cr(VI).

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135

Fig. 3.2 The variations in Cr(VI) concentration were evaluated under different experimental conditions, including sterilized sample + sterilized medium (Treatment I), sterilized sample + deionized water (Treatment II), non-sterilized sample + sterilized medium (Treatment III), and non-sterilized sample + deionized water (Treatment IV). Reprinted from Ref. [3], Copyright 2019, with permission from the Nonferrous Metals Society of China and Elsevier

Fig. 3.3 Species of microorganisms in original soil. Reprinted from Ref. [3], Copyright 2019, with permission from the Nonferrous Metals Society of China and Elsevier

3.1.2 Kinetics of Cr(VI) Microbial Reduction As illustrated in Fig. 3.4, The kinetic curve of the indigenous microorganisms for Cr(VI) reduction was studied with the addition of extra culture medium and under aerobic conditions. Zero-order, first-order, and second-order mathematical models were used to analyze the reduction of Cr(VI) under acidic to neutral pH conditions. The reduction process of Cr(VI) was observed to occur in two distinct phases

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Fig. 3.4 Reduction kinetic characteristics of Cr(VI) reduction in soil. Reprinted from Ref. [3], Copyright 2019, with permission from the Nonferrous Metals Society of China and Elsevier

(Fig. 3.4). During Phase A, the activity of microorganisms facilitated the initial reduction of Cr(VI). Subsequently, the presence of culture medium substances in the soil further enhanced the indigenous microorganisms’ ability to reduce Cr(VI). Based on the kinetic data obtained from Phase A, it can be concluded that the Cr(VI) reduction process can be effectively described by an exponential kinetic model as follows  C = −5.56 exp

t 7.33

 + 15338.88, R 2 = 0.995

(3.1)

where C is the concentration of soluble Cr(VI) (mg kg−1 ), and t is time (h). Due to the variations in factors affecting Cr(VI) removal and differences in reduction conditions such as pH, the current study’s fitted results of the kinetic model are inconsistent with the previously reported pseudo-first-order kinetics of amorphous FeS2 [4], S(IV) under acidic conditions [5], bacterial metabolic kinetics of zero-valent iron in groundwater [6], and sulfate-reducing bacteria [7]. Moreover, there is currently no reported kinetic model for the reduction of Cr(VI) specifically in chromium-contaminated soil. The reduction of Cr(VI) in this context is a complex process that involves the interaction of microorganisms with the unique characteristics of the soil system. In Phase B, Cr(VI) reduction was fitted to the linear equation kinetics model as follows: C = 1157.65 − 14.43t, R 2 = 0.999

(3.2)

Consistent with the previous studies which reported reduction kinetics of Cr(VI) with metallic iron [8] and bivalent manganese [9], the results showed that Phase B in Fig. 3.4 was interestingly in agreement with the degradation of phenol by Sphingomonas sp. GY2B-embedded beads at 30 °C and pH 7.0 [10, 11]. This indicates that

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137

the reduction of Cr(VI) remained consistent at a rate of 14.43 mg (kg h)−1 throughout Phase B. This steady reduction rate was attributed to the combined effects of various substances in the soil system.

3.1.3 Changes in pH and ϕ h During Cr(VI) Microbial Reduction The valence state of Cr is strongly influenced by pH and ϕ h values present in the aqueous soil solution. When it comes to environmental stability, Cr(III) derivatives are generally more stable than Cr(VI). Moreover, they will form stable complexes with both organic and inorganic ligands primarily. Under neutral pH conditions, chromium (III) is inclined to bond with OH− ions and create precipitates [12]. The reduction process of Cr(VI) results in a noticeable change in pH. Notably, in Treatments II and IV, there was no discernible variation in soil pH. However, within a span of 12 h, the pH experienced a slight decrease in I and III. Subsequently, the pH exhibited a rapid decline before stabilizing at 8.0 (Fig. 3.5). Soil pH is influenced by factors such as carbon dioxide concentration, organic acids, and inorganic acids. As evidenced by the data presented in Fig. 3.5, the pH values for II and IV align with the aforementioned factors. The results imply that pH plays a vital role in the bioremediation of soil. It is worth noting that the indigenous microorganisms tend to thrive under alkaline conditions with a pH range of 9.0– 10.0, which further highlights the significance of pH in soil remediation [13]. Other factors, such as soil organic matter and culture medium, can also affect the activity of indigenous microorganisms. Additionally, in Treatment III, Cr(VI) reduction was observed to be positively correlated with the alteration in pH values, which helped maintain a favorable metabolic environment for the indigenous microorganisms. Fig. 3.5 Change in pH values in the reduction process of Cr(VI) under different treatments. Reprinted from Ref. [3], Copyright 2019, with permission from the Nonferrous Metals Society of China and Elsevier

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3 Microbial Remediation of Chromium-Polluted Soil

Fig. 3.6 Change in ϕ h in the reduction process of Cr(VI) under different treatments. Reprinted from Ref. [3], Copyright 2019, with permission from the Nonferrous Metals Society of China and Elsevier

For Treatments I, II, and IV, the ϕ h value only exhibited a slight increase, and the Cr(VI) content remained at a high concentration (Fig. 3.6). During these experiments, the soil system maintained a high oxidation–reduction state, which was solely dependent on soil components (Treatments II and IV) and exogenous medium (Treatment I). In Treatment III, the ϕ h value exhibited an initial rise within the initial 12 h, followed by a continuous decrease. The increase in the soil system’s oxidation–reduction potential was linked to the dissolution of significant ions in the soil. These results demonstrated a distinct correlation between the fluctuation in ϕ h and the reduction of Cr(VI). The combined effect of Cr(VI) reduction and ϕ h alteration played a vital role in stabilizing the soil bioremediation system. The ϕ h value in the soil system remained stable until 42 h, which suggested that factors other than the Cr(VI) reduction might have influenced the system after this point.

3.1.4 Changes in Concentrations of Fe2+ , Mn2+ , SO4 2− and NO3 − Impact of predominant ions present in the soil on the reduction of Cr(VI) was found to be multifaceted. To gain a deeper understanding of the sequential reduction of Cr(VI) in the soil system, samples obtained from Treatment III were employed to analyze the levels of Fe2+ , Mn2+ , SO4 2− , and NO3 − throughout the reduction process of Cr(VI). During the initial 30 h, the Fe2+ content exhibited a slight decrease, but it then rapidly increased and remained stable at 4.4 mg kg−1 (Fig. 3.7). Meanwhile, Mn2+ was not significant during the first 30 h. However, it subsequently experienced a sharp rise and maintained a stable concentration of approximately 0.278 mg kg−1 .

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139

The concentration trends of Fe2+ and Mn2+ exhibited almost identical changes. Additionally, certain intermediate products of the reaction and complexes formed with major ions (such as Mn2+ and Fe2+ ) in the soil were confirmed to enhance the rate of Cr(VI) reduction during the reduction stage. After 30 h, there was a significant decrease in the concentration of Cr(VI) in Treatment III, accompanied by a decline in ϕ h and pH values. This suggested that Fe2+ and Mn2+ had an additional impact on the reduction of Cr(VI). Furthermore, according to the standard electrode potential, the reduction of Fe3+ and Mn4+ occurred subsequent to the reduction of Cr(VI). The reduction of Cr(VI) facilitated the dissolution of Fe2+ and Mn2+ in the bioremediation system. The concentration of SO4 2− exhibited only a slight decrease during the course of Cr(VI) reduction, suggesting that SO4 2− had minimal impact on Cr(VI) reduction (Fig. 3.7). In the first 12 h, there was negligible alteration in the content of NO3 − . However, after 12 h, it rapidly declined and became undetectable at 18 h. Fig. 3.7 Concentration changes of Fe2+ and Mn2+ (a) and SO4 2− and NO3 − (b) in the bioremediation process of Cr(VI). Reprinted from Ref. [3], Copyright 2019, with permission from the Nonferrous Metals Society of China and Elsevier

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3 Microbial Remediation of Chromium-Polluted Soil

As previously stated, there was no significant change observed in the concentration of Cr(VI) in the soil during the first 18 h (Fig. 3.2). Only when the NO3 − concentration reached its minimum level was a notable decrease in the concentration of Cr(VI) observed. This indicates that the reduction of NO3 − occurred before the reduction of Cr(VI). Table 3.1 displays the alteration in the content of major ions in the soil during the process of Cr(VI) reduction, coinciding with the decrease in ϕ h values. It was noted that NO3 − experienced rapid reduction once the ϕ h value dropped below 185.6 mV, reaching complete reduction at a ϕ h value of 160.3 mV. However, the concentrations of Mn2+ and Fe2+ showed no significant changes when the ϕ h value exceeded 125.3 mV. However, the concentrations of Mn2+ and Fe2+ increased rapidly with the decrease in ϕ h values. The change in SO4 2− concentration during the reduction process was not evident. The hypothesis proposed that high-valent irons underwent reduction to Fe(II) during Cr(VI) reduction, followed by the transfer of electrons to Cr(VI), leading to its reduction to Cr(III). The reduction potentials of these major ions in soil were inconsistent with their standard reduction potentials and were influenced by the complex interplay of soil composition (such as pH value and organic matter) and the comprehensive effect of the Cr(VI) bioremediation system. Therefore, the apparent reduction sequence of major ions in the chromium-contaminated soil system was NO3 − > Mn4+ > Fe3+ > Cr(VI) > SO4 2− . Consequently, the reduction sequence of bioremediation by the indigenous microorganisms played a crucial role in the chromium-polluted soil. Table 3.1 Concentration changes of major ions in the chromium-contaminated soil system ϕ h /mV

Concentration of major ions/(mg kg−1 ) Cr(VI)

Fe2+

Mn2+

SO4 2−

NO3 −

198.6

1521.9

3.16

0.061

812.6

13.1

189.3

1527.3

3.30

0.091

918.8

13.6

185.6

1528.0

3.33

0.102

881.7

13.1

160.3

1460.8

3.22

0.083

818.7

0

144.3

1376.0

2.75

0.061

869.4

0

125.3

1239.2

2.92

0.064

819.4

0

109.0

777.2

3.74

0.113

759.1

0

59.0

547.7

4.13

0.224

816.1

0

56.6

461.6

4.44

0.310

843.7

0

46.3

389.9

3.86

0.168

813.5

0

47.6

292.6

4.39

0.278

813.7

0

48.0

199.2

4.40

0.284

820.6

0

Reprinted from Ref. [3], Copyright 2019, with permission from the Nonferrous Metals Society of China and Elsevier

3.2 Optimization of Cr(VI) Bioremediation in Polluted Soil

141

3.2 Optimization of Cr(VI) Bioremediation in Polluted Soil The evaluation of factors that influence the bioremediation process of Cr(VI)contaminated soils was conducted using column recycle leaching (Fig. 3.8). The entire structure included a PVC reservoir, organic glass columns, a pump, leachate collectors, and PVC tubes. The leaching column consists of multiple sampling ports located at depths of 20, 40, 60, and 80 cm, respectively. An external water source is heated to a specific temperature and introduced into the lateral space of the inner column to maintain a constant temperature during the bacterial remediation experiment. Below the contaminated soil, approximately 2 cm of large stones are placed as a filtering medium with a height of approximately 5 cm. The leachate is pumped back into the top of the leaching column and sprayed onto the soil. Soil samples are collected daily from each sampling port for analysis of Cr(VI) content. Distilled water is added to the system daily to compensate for water loss due to evaporation. The experiments were conducted at 25 ± 3 °C and an initial pH of 10.0.

Soil

Sampling ports

Reservoir

Leachate collector

Pump

Fig. 3.8 Schematic diagram of column recycling leaching. Reprinted from Ref. [14], Copyright 2013, Central South University Press and Springer-Verlag GmbH Germany, part of Springer Nature

142

3 Microbial Remediation of Chromium-Polluted Soil

3.2.1 Effect of Particle Size To investigate the influence of soil particle size on Cr(VI) reduction, experiments were carried out using the sterilized nutrient medium and soils with different particle sizes (4 cm). Figure 3.9 illustrates the outcomes obtained, demonstrating that soil particle size exerted a notable influence on the leaching and reduction processes of Cr(VI). The study found that native bacteria possess capacity to eliminate Cr(VI). Following 6 days of remediation, the concentration of water-soluble Cr(VI) in soils decreased from 1520.54 to 2.05 mg kg−1 in the 4 cm size fraction, the water-soluble Cr(VI) content in soil after treatment was 172.60 mg kg−1 , and the leaching rate of water-soluble Cr(VI) was 88.65%. This observation could be attributed to small size of the soils polluted by the slag containing Cr, which facilitates the dissociation of Cr(VI) and exposes Cr(VI) to the solid–liquid interface, thereby increasing the contact area and promoting the reaction process. Additionally, The rate of water-soluble Cr(VI) removal in the 1–2 cm fraction was comparable to that of the particle size fraction less than 1 cm. Consequently, the soils within the 0–2 cm depth range would be suitable for bioremediation applications and warrant further experimentation. It was noted that the rate of leaching for Cr(VI) exhibited fluctuations across distinct time intervals for each individual curve. For instance, considering the curve

Fig. 3.9 Removal of water soluble Cr(VI) for different particle sizes. Reprinted from Ref. [14], Copyright 2013, Central South University Press and Springer-Verlag GmbH Germany, part of Springer Nature

3.2 Optimization of Cr(VI) Bioremediation in Polluted Soil

143

Table 3.2 Leaching toxicity of the remediated soils with different particle sizes Leaching Original soil toxicity (mg L−1 ) Cr(VI)

Soils after bioremediation (mg L−1 ) > CPS ~ FeS > FeS2 , and in slightly acidic soil, Na2 S >> FeS > CPS ~ FeS2 , the order was Na2 S >> FeS > CPS ~ FeS2 . A soluble sulfide called calcium polysulfide (CPS) can be effective in a wide range of pH environments and can maintain a highly reducing environment over long periods of time [65, 66]. Extremely contaminated soil with a total chromium (VI) concentration of 18,770 mg kg−1 was stabilized using CPS by Kanchinadham et al. [67]. The findings showed that after the application of CPS, the Cr(VI) and total Cr contents in the TCLP leachates were below detectable limits (0.003 mg L−1 ). When 3% CPS was added as a reductive agent together with mechanochemical treatment, Yuan et al. were able to greatly reduce the leaching concentration of hexavalent chromium from 115 mg L−1 to 0.51 mg L−1 . Natural iron oxides in the soil may make the use of CPS to clean up Cr(VI)-contaminated soil easier [19]. According to

A planting site in Beijing, China

Tamil Nadu Chromates and Chemicals Limited, located at Ranipet, Tamilnadu, India

An electroplating workshop, Measuring Tool Co., Ltd., located in Jingjiang, China

Cr(VI)-spiked soil

Shuitou town, China

CaS5

CaS5

FeSO4 Na2 HSO3 Na2 S Na2 S2 O3

Long duration −NaS2 SO3

S0

5%(w/w)

Dosage

Total Cr(VI): 12.6–42.5 mg kg−1 PBE Cr(VI): 1.1–4.6 mg kg−1

Leached Cr(VI):50 mg L−1

Total Cr: 2239.59 mg kg−1 Total Cr(VI): 1915.41 mg kg−1

Total Cr(VI): 18,770 mg kg−1 total Cr: 44,615 mg kg−1

4.0 mg g−1

1.67%(w/w)

5%(w/w) 5%(w/w) 5%(w/w) 5%(w/w)

3 times molar stoichiometric ratio of C(VI)

Leached Cr(VI):115 mg 3%(w/w) L−1

An abandoned chromic salt Total Cr(VI): factory in Ji’nan, 13,806.4 mg kg−1 Shandong, China leached Cr(VI): 663.98 mg L−1

Na2 S

Initial concentration

Sources of contaminated soil

Amendments

Table 4.3 Typical S-based materials for remediation of Cr(VI)-contaminated soil

[19]

[24]

References

(continued)

PBE Cr(VI) dropped to less [18] than 0.4 mg kg−1

The relative Cr(VI) removal [25] efficiency is 98.25%, obviously higher than 18.01% of bare NaS2 SO3

these four chemicals can [76] reduce Cr6+ Soil to below 5 mg kg−1 , with a removal rate of more than 99%

Both Cr(VI) and total Cr [67] concentrations were below detectable limits (0.003 mg L−1 ) in the TCLP leachates

The Cr(VI) leaching concentration decreased to 0.51 mg L−1

The Cr(VI) leaching concentration decreased to 0.84 mg L−1

Main performances

4.1 Chemical Reductive Materials for Cr(VI) 185

Cr(VI)-spiked soil

A spent industrial site for dumping chromium slag in Chongqing, China

Cr(VI)-spiked soil

Cr(VI)-spiked soil

Cr(VI)-spiked soil

CMC-mFeS

CMC-nFeS

CMC-nFeS

FeS2

FeS2

0.1 g L−1 (soil to solution ratio of 1 g:1 mL)

104.0 mg L−1

Molar ratio of FeS to Cr(VI):1.5:1

Dosage

Total Cr:284.5 mg kg−1 total Cr(VI):106.2 mg kg−1

[14]

[70]

[15]

References

Decreased the equilibrium [75] water leachable Cr(VI) by > 99.0% at pH 6.0 and by > 70.0% at pH 9.0

The reducing capacity of CMC-nFeS was 54.68–198.74 mg Cr(VI) g−1 FeS

The TCLP-leached Cr(VI) decreased to an acceptable level of 46.8–80.7 μg L−1 after treatment

Approximately 98% of Cr(VI) in soil was reduced in 3 d and Cr(VI) concentration decreased to 16 mg kg−1

Main performances

Stoichiometric ratio of 2, 4, Cr(VI) concentration [74] 8 decreased to 61.0, 19.8, and 2.38 mg kg−1 , respectively

The equilibrium Cr(VI) 0.48 g L−1 (soil to solution concentration: 5.21 mg ratio of 2 g:35 mL) L−1 (pH ~ 9.0) and 4.65 mg L−1 (pH ~ 6.0)

Total Cr(VI): 56.01–502.21 mg kg−1

Total Cr(VI): 204.84 mg kg−1 TCLP-leached Cr(VI):4.58 mg L−1

Total Cr(VI): 1407 mg kg−1 total Cr: 2089 mg kg−1

Initial concentration

Reprinted from Yang et al. [26] Copyright 2021, with permission from Elsevier

Sources of contaminated soil

Amendments

Table 4.3 (continued)

186 4 Mechanism of Chemical Reduction of Cr(VI)

4.1 Chemical Reductive Materials for Cr(VI)

187

Zhang et al. findings’ the augmentation of Cr(VI) removal rose from 0 to 9 g kg−1 with the concentrations of goethite or hematite [68]. By acting as a catalyst, natural iron oxides in soil might reduce Cr(VI) utilizing CPS and significantly lower the apparent activation energy of the reaction between Cr(VI) and S(-II). Although the reduction of iron oxides by S(-II) accelerates the conversion of electrons to adsorbed Cr(VI), the reduction of Cr(VI) by CPS is made easier by the Fe(II) species adsorbed on the surface of goethite. Additionally, the adsorbed Fe(II) species had a lower redox potential than common Fe(II) species. To treat soils contaminated with hexavalent chromium, sodium thiosulfate (Na2 S2 SO3 ), a typical sulfide reducing agent, is utilized with ideal reducing power. Due to its high solubility, it tends to dissolve in water and disappear, which results in short duration and inefficient use. Here, He et al. created a long-lasting Na2 S2 SO3 complex supported by albite to remediate Cr(VI)-contaminated soils [25]. The outcomes showed that this combination outperformed bare Na2 S2 SO3 in terms of reduction performance when cleaning up Cr(VI)-contaminated soil. The time and effectiveness of Cr reduction by Na2 S2 SO3 were improved when sodium thiosulfate was effectively combined with concavite through H-bonds and other physical interactions (VI) [25]. Effectively converting Cr(VI) to Cr(III), soluble sulfur compounds typically produce less hazardous and mobile Cr(III) hydroxides. However, Cr(OH)3 is not sufficiently stable as the primary reduction product and can be chelated by low molecular weight organic molecules and hydrogen ions before being further reoxidized to Cr (VI). Furthermore, recent research showed that the addition of CPS directly raised the pH of the soil from 6 to 12, and that even after a year, the soil remained alkaline (pH > 9), which is unsuitable for plant growth. In conclusion, the wide pH range of both sodium sulfide and calcium polysulfide makes them both promising reductive agents. However, the remediation procedure involving those sulfides will result in the production of hazardous hydrogen sulfide gas; as a result, the dose needs to be tightly controlled. Future research should pay greater attention to how the presence of calcium ions in the application of calcium polysulfide would alter the solubility of chromate. Also, more research is required before sodium thiosulfate is applied widely because the removal effects and application circumstances are yet unknown.

4.1.3.2

Elemental Sulfur for the Remediation of Cr(VI)-Contaminated Soil

In agricultural soils, elemental sulfur (S0 ) is commonly used to provide nutrients to promote plant growth. Due to its high redox activity, it may be able to convert polluted soil containing Cr(VI) to Cr (III). After adding 1.5 t ha of S0 , the DTPAextractable chromium in chromium-contaminated soil decreased from 393 mg kg−1 to less than 0.005 mg kg−1 after 10 weeks [69]. Shi et al. discovered that S0 might lessen the bioavailability and toxicity of Cr(VI) in soil when used to treat soil that had been contaminated by a leather tannery. After the addition of 4.0 mg g−1 S0 , the

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4 Mechanism of Chemical Reduction of Cr(VI)

phosphate buffer extraction (PBE) Cr(VI) levels of the three soil samples decreased to below the 0.4 mg kg1 limit for Cr(VI) in agricultural soil use according to the Canadian Soil Quality Guidelines (CSQG) [18]. It has been demonstrated that the addition of S0 enhanced the amount of soluble SO4 2− while lowering the amount of Cr (VI).

4.1.3.3

Iron Sulfides for Remediating Cr(VI)-Contaminated Soil

Due to the presence of both Fe(II) and S species as reducing agents, iron sulfide has a relatively high reducing capacity. Mackinawite and pyrite are the two primary minerals that contain natural sulfides. Yet, it is discovered that under neutral or alkaline circumstances, their reduction rates of Cr(VI) are modest. The manufactured FeS and FeS2 particles are smaller and have a higher specific surface area than natural minerals, which increases their reactivity to Cr(VI) reduction. In addition, compared to soluble sulfides, FeS and FeS2 not only gradually release Fe2+ and S2− or S2 2− ions to reduce Cr(VI), but also adsorb Cr(VI) through surface reactions. Cr(VI) can be successfully immobilized in soil by ferrous sulfide (FeS), which is produced from sulfide and ferrous solutions. However, one of the main problems with the use of FeS particles is aggregation, which reduces the specific surface area and the efficacy of contaminant removal. As a result, different stabilizers have been suggested in recent years as a way to stop particle aggregation. Li et al. created microscale FeS particles that were CMC-stabilized and used them to treat soil that had been contaminated with Cr(VI) [14]. In a 1.5:1 molar ratio, the concentration of Cr(VI) fell from 1407 to 16 mg kg−1 during the course of three days. Recently, nanoscale FeS particles stabilized with CMC were created and used to treat saturated soil contaminated with hexavalent chromium [70]. After elution with 45 PV of FeS NP solution, the Cr(VI) concentration in the filtrate was lower than the Class I value of 0.005 mg L−1 of the groundwater quality standard (China, GB/T14848-93), and the amount of Cr(VI) leached by TCLP decreased from 4.58 mg L−1 to 46.8– 80.7 g L−1 . Moreover, a fluid CMC-nFeS substance demonstrated significant Cr(VI) removal effectiveness [14]. The reduction and fixation of Cr(VI) was best when the pH of the soil was between 5.86 and 8.55. Then, Cr(VI) was converted to safe forms of Cr(OH)3 and Fe0.75 Cr0.25 OOH. According to reports, 90 days after applying FeS, there was no change in the pH of the soil or secondary oxidation of Cr(VI) [14]. The benefits of chitosan are that it is biodegradable, non-toxic and has a large number of primary amine (–NH2 ) and hydroxyl (OH) groups, making it a possible stabilizer to stop the aggregation of nanoscale particles. The creation of MC-FeS composites with chitosan stabilization, which increased the ability of FeS magnets to remove Cr(VI), was successful [71]. The strong affinity of chitosan for Cr(VI) and Fe(II) adsorbed on the surface of MC-FeS composites could be the reason. The MC-FeS composite also demonstrated great aging stability and worked effectively in a long-term reaction system for the elimination of Cr(VI). Particle aggregation can be successfully avoided, and the addition of a stabilizer and supporting material at the same time can improve the physical stability

4.1 Chemical Reductive Materials for Cr(VI)

189

and removal effectiveness of the particles. To immobilize Cr(VI)-contaminated soil, Lyu et al. prepared CMC-FeS@biochar using CMC and biochar as stabilizer and support material, respectively. They confirmed that the composite had higher removal capacity and affinity for Cr(VI) compared to bare FeS and plain biochar, which could be attributed to the synergistic effect due to the interaction between the individual components [72]. However, soil pH was crucial in the reduction of Cr(VI) by FeS particles. The removal effectiveness of Cr(VI) was better than 92% when the pH ranged from 3.80 to 8.55, but it decreased to 87.67% when the pH rose to 9.82 [14]. This may be due to the fact that in strongly alkaline soils, a large amount of OH competes with Cr(VI) for the active site, forming Fe(OH)x on the surface of the particles, further preventing them from coming into contact [73]. A schematic of the removal of Cr(VI) by CMC-nFeS in neutral soil may be found in Fig. 4.4 [14]. Fe2+ , the main reductant, interacted with Cr(VI) in neutral soils (pH = 6.65) to form Cr(OH)3 and Fe(OH)3 . Sulfur and Cr(OH)3 were also created when a little amount of S2− interacted with Cr(VI). Moreover, Crx Fe(1−x) OH was created, which helped to immobilize Cr (VI). It is advantageous to maintain a reductive soil environment because FeS2 theoretically gives 15 electrons when it oxidizes to Fe3+ and SO4 2− . With a framboidal shape and a higher specific surface area of 4.15 m2 /g than nanopyrite, Li et al. produced amorphous FeS2 (50–200 nm) [73]. With a dosage of 0.6 g L−1 , FeS2 demonstrated high efficiency for the reduction of Cr(VI), removing 5.41 mM Cr(VI) in 24 h. In their further investigation, they found that amorphous FeS2 (stoichiometry 4) greatly reduced the leachability and bioaccessibility of Cr(VI) in the soil with an initial total Cr(VI) content of 106.2 mg/kg throughout the 30-day remediation period [74]. Additionally, they used column experiments using groundwater and acid rain as influents to evaluate the long-term stability of Cr(VI)-remediated soil and discovered

Fig. 4.4 Schematic diagram of Cr(VI) removal by CMC-nFeS in neutral soil. Reprinted from Yang et al. [26] Copyright 2021, with permission from Elsevier

190

4 Mechanism of Chemical Reduction of Cr(VI)

that stabilization of Cr(VI) in contaminated soil was successful for 250 PV (equal to maintenance for 1575 days). The use of sodium acetate as a buffering agent in the production process has recently created FeS2 particles with a more crystalline structure [75]. The formed pyrite successfully immobilizes Cr(VI) in both water and Chinese loess models, as it appears to have a more regular cubic morphology, larger specific surface area (7.42 m2 g−1 ) and more consistent dimensions. Leachable Cr(VI) was reduced by > 99.0% at pH 6.0 and by > 70.0% at pH 9.0 when a soil containing Cr(VI) was treated with 0.48 g L−1 (as Fe) of FeS2 . The leachable Cr(VI) was almost completely immobilized when the dose was doubled. FeS2 has a high reducing capacity, high product stability, and minimal adverse effects on soil pH. However, during remediation, the Cr(III) and Fe(III) hydroxyl complexes precipitated on the surface of FeS2 may prevent subsequent interaction with Cr(VI) and weaken its reducing ability. The concentration of Cr(VI) in these soils is relatively low, but the use of synthetic FeS2 for remediation of Cr(VI)-contaminated soils is infrequent. In general, iron sulfide is more capable of reducing Cr(VI) than Fe(II)-containing reductants and produces more stable reduction products than soluble sulfides (Fe– Cr coprecipitates as Crx Fe1-x (OH)3 ); as a result, they could be useful substitutes for reducing Cr(VI). However, based on the data now available, they are only appropriate for the remediation of soil contaminated with Cr(VI) at low concentrations. Thus, it will soon be important to identify ways to optimize their decreasing capacity.

4.1.3.4

Mechanism of S-based Amendments for the Remediation of Cr(VI)-Contaminated Soil

The following is an expression for how Cr(VI) and sulfide ions react [16]: S2− + H2 O → HS− + H+ + 2OH− 2− − − 8CrO2− 4 + 3HS + 17H2 O → 8Cr(OH)3 + 3SO4 + 13OH

(4.16) (4.17)

The following describes how calcium polysulfide, which has the chemical formula CaS5 , reacts with Cr(VI) [68]: + 0 2CrO2− 4 + 3CaS5 + 10H → 2Cr(OH)3 + 15S

(4.18)

2− + S0 + 2CrO2− 4 + 2H2 O + 2H → 2Cr(OH)3 + SO4

(4.19)

Due to the reduction of Cr(VI) by CPS, elemental sulfur is produced, which can further reduce Cr(VI) to Cr(III). As a result, CPS is highly effective and capable of reducing Cr(VI). Goethite and hematite, which are common iron-containing minerals in soil, might encourage the removal of Cr(VI) by CPS, which can be stated as follows: S(−II) + 2Fe(III) − (hydr)oxides → 2Fe(II) + S(0)

(4.20)

4.1 Chemical Reductive Materials for Cr(VI)

3Fe(II) + Cr(VI) → 3Fe(III) + Cr(III)

191

(4.21)

By reacting with Cr(VI), the Fe(III)-generated Fe(II) was reoxidized into Fe(III), then reduced once more by S-2, creating a redox cycle and extending the soil’s reducing atmosphere. The exothermic reaction of Eq. (4.19), as previously indicated, provides the theoretical foundation for elemental sulfur to reduce Cr(VI) to Cr(III). Cr(VI) can be reduced into Cr by the structural Fe(II) and S(-II) or S2 (-II) of Fe(II) sulfides (III). The ensuing equations display the potential pathways of Cr(VI) reduction by FeS [70, 77, 78]: FeS + H+ → Fe2+ + HS−

(4.22)

+ 3+ 3Fe2+ + HCrO2− + Cr3+ + 4H2 O 4 + 7H → 3Fe

(4.23)

− + 3+ 8HCrO− + 3SO2− 4 + 3HS + 29H → 8Cr 4 + 20H2 O

(4.24)

− + 3+ − 2HCrO− + SO2− 4 + HS + 7H → 2Cr 3 + 5H2 O

(4.25)

Equations (4.22)–(4.24) explain the potential reduction pathway of the absorbed Cr(VI) when FeS is insufficient and SO4 2− is a by-product of sulfide oxidation, but in the case of excess FeS, the potential reduction reactions follow Eqs. (4.22)–(4.25). The pH then increases as the hydrogen ions are consumed while these processes continue. As a result, Fe(III), Cr(III) and hydroxyl ions are produced along with Cr(OH)3 , Fe(OH)3 , Crx Fe1-x (OH)3 or Crx Fe1-x OOH products. The following can be used to express the general reactions in Cr(VI)-contaminated soil restoration using FeS2 [73, 74, 79]: 3+ FeS2 + CrO2− + Fe3+ + 2S0 + 8OH− 4 + 2H2 O → Cr 3+ − 3FeS2 + 7CrO2− + 3Fe3+ + 3S2 O2− 4 + 19H2 O → 7Cr 3 + 38OH

(4.26) (4.27)

S0 , S2 O3 2− and S2− are the main by-products of S2 2− oxidation and can maintain the reducing environment of Cr(VI) for a longer period of time. precipitates or co-precipitates of Cr(OH)3 , Fe(OH)3 , Crx Fe1–x (OH)3 or Crx Fe1–x OOH develop with the formation of Fe(III), Cr(III) and OH.Soil contaminated with hexavalent chromium has become a serious environmental problem worldwide. In recent years, an increasing number of papers have focused on the use of various techniques to address hexavalent chromium contamination in soils. Among the physical, chemical and biological treatment methods, the reduction fixation method for remediation of Cr(VI)-contaminated soil has been widely used both at home and abroad because of its fast remediation speed, high efficiency and simple operation. This work

192

4 Mechanism of Chemical Reduction of Cr(VI)

reviews three classes of reductive materials for efficient immobilization of Cr(VI)contaminated soil: organic materials, Fe-bearing reductants, and S-based compounds. Theoretical support for upcoming real-world applications is provided by detailed understanding of the usage of these reductive materials for the treatment of Cr(VI)contaminated soil. Future research should take into account the information gaps and difficulties that still exist in remediation techniques for Cr(VI)-contaminated soils using chemical reduction immobilization. (1) Due to increased remediation standards and the complexity of soil hexavalent chromium contamination, it is difficult for existing reducing materials to meet remediation requirements. For example, the reducing capacity of these materials may still not be the best choice for cleaning up source areas or “hot spots”. Depending on the characteristics of the particular Cr(VI)-contaminated soil, it is necessary to create new reducing materials. These materials could be slowrelease materials that continuously provide electron donors to maintain longterm Cr(VI) reduction capacity, multifunctional materials with high reactivity and selectivity, etc. (2) It often takes a fair amount of time and effort to successfully remediate soil that has been contaminated with Cr(VI) using classic reductive material applications. Therefore, the use of easy-to-use auxiliary techniques (e.g., mechanical ball milling, microwave and ultrasonic) in the remediation of Cr(VI)-contaminated soils by reducing materials would be considered as a potential approach to significantly reduce remediation time and improve remediation efficiency. (3) It is important to consider how reducing materials affect soil properties. The physicochemical and microbiological properties of soils may be altered by the addition of various elements, which can also have an impact on soil properties. As a result, it is crucial to carry out further study and use reductive materials in soil remediation. (4) At present, most studies have focused on the fixation effect of reducing materials, and less attention has been paid to the stability of reduced trivalent chromium in remediated soils. Because of the intricate nature of the soil system, Cr(III) may reoxidize to Cr(VI) and move once more in response to changes in the environment. It is now crucial and important to assess the long-term stability of remedial soil.

4.2 Enhanced Cr(VI) Removal by an “In Situ Synthesized” Iron-Based Bimetal Material For the sequestration of contaminants, iron-based compounds with redox activities are frequently utilized [80, 81]. In particular, zero-valent iron (ZVI) is of considerable interest due to its powerful reducing ability and high reactivity [82, 83]. Unfortunately, there are still difficulties with implementing ZVI [84]. During ZVI corrosion,

4.2 Enhanced Cr(VI) Removal by an “In Situ Synthesized” Iron-Based …

193

The pristine ZVI is easily passivated due to the release of iron ions and the generation of surface oxide layers. This will further prevent corrosion of ZVI and electron transport between ZVI and the target impurity. Then, the reactivity of ZVI decreases rapidly as passivation proceeds [85]. Several techniques have been created to date to activate ZVI prior to actual application [77, 86–88]. The most frequently used methods include the deposition of secondary metals (such as platinum, silver, nickel or copper) on the surface of ZVI to create bimetallic nanoparticles [48, 56, 89]. The addition of a second metal along with ZVI creates an electrochemical pair that may accelerate electron transfer and release Fe2+ to enhance the reduction of Cr(VI). However, the iron-based bimetallic approach has several significant issues that need to be addressed. (1) These ironbased bimetals are typically used to sequester contaminants from pollution caused by a single heavy metal. Cationic heavy metals (e.g. Cu2+ , Co2+ , Ni2+ , etc.) and anionic Cr(VI) usually coexist with natural water and waste due to mining, electroplating and metallurgical industrial activities. (2) Exogenous metals will leach from the produced bimetallic pellets after the ZVI is used up, which may cause another environmental problem [90]. (3) Bimetallic nanoparticles’ practical use is constrained by the high cost of production. (4) It is still unclear how the electrochemical effects in ZVI-based bimetals work mechanically. Here, we mainly focus on the contaminant sequestration of coexisting cationic heavy metals (e.g. Cu2+ , Co2+ , Ni2+ , etc.) and anionic Cr(VI) in electroplating effluent. ZVI is rapidly corroded upon contact with ionic solutions, displacing less reactive metals (e.g. copper, cobalt, etc.) by electron transfer. Coincidentally, the wastewater from electroplating typically contains both of these metal cations. “On-site” synthesis of Fe-based bimetals is very possible through the spontaneous reduction and precipitation of these coexisting cations (including Cu2+ , Co2+ , Ni2+ ) on ZVI. Passivation of ZVI is straightforward using a self-activating technique inspired by the spontaneous synthesis of bimetallic systems, without the use of exogenous metal additions. Most importantly, “in situ production” of ironbased bimetals may be effective in reducing Cr(VI) and removing any co-occurring cationic metals from the plating wastewater. The use of “in situ synthesis” of iron-based bimetals in complex ionic solutions will be limited by the limited adsorption capacity and selectivity of ZVI for the target heavy metal cations [87]. Therefore, improving the interaction between ZVI and these coexisting cations will help to achieve this goal. Thus, achieving this goal will contribute to a successful program of enhancing the contact between ZVI and these coexisting cations. It has been demonstrated that hydroxyapatite (HAP, Ca10 (PO4 )6 (OH)2 ) has reactive Ca2+ , PO4 3− and OH− groups with strong affinity and selectivity for cationic heavy metals, especially for divalent metal cations such as Cu2+ , Co2+ and Ni2+ [91, 92]. To sequester the coexisting Cr6+ , Cu2+ , Co2+ and Ni2+ in electroplating wastewater, ZVI was loaded onto HAP to produce HAP-supported ZVI composites (ZVI/HAP). HAP may enhance the adsorption of coexisting cationic ions in this situation. Then, the adsorbed cationic metals are expected to produce bimetals “in situ” for contaminant sequestration.

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4 Mechanism of Chemical Reduction of Cr(VI)

4.2.1 Activity Toward Cr(VI) Removal According to the XRD pattern, HAP has a highly crystalline phase with a typical hexagonal structure (Fig. 4.5a; JCPDS NO. 09-0432 P63/m (176)) [93]. The distinctive diffraction peaks of HAP were still present after the deposition of ZVI, however they were less intense than before. The diffraction peak of JCPDS NO. 06-new 0696 at 2 = 44.7°, which is indexed to Fe0 [94], indicates the successful production of Fe0 particles on the surface of the HAP [94]. By using TEM and TEM-coupled mapping, the morphology and elemental composition of ZVI/HAP were further investigated. HAP has a distinct nanorod structure with sizes of 20–100 nm and 2–10 nm, respectively (Fig. 4.5b). As shown in Fig. 4.5c, ZVI is present on the surface of HAP in the form of spherical particles with a diameter of 80–100 nm and a dark color. Further evidence that ZVI is uniformly distributed on the HAP surface comes from the Ca, P, O, and Fe elements, which are uniformly distributed on the ZVI/HAP mapping images (Fig. 4.5d). Then, We investigated how the prepared ZVI/HAP affects the removal of Cr(VI). ZVI/HAP can remove 95.05% of Cr(VI) within 24 h, whereas HAP could only remove less than 20% of Cr(VI) (Fig. 4.5e). As a result, ZVI/HAP performed significantly better in removing Cr(VI) than HAP. The elimination of Cr(VI) was fitted using kinetic models, such as the pseudofirst-order and pseudosecond-order kinetic models. The results showed that the removal of Cr(VI) by ZVI/HAP obeyed a pseudo-first-order kinetic model, which was mainly brought about by the chemical reaction (R2 of 0.9875 for the pseudo-first-order kinetic model > R2 of 0.9768 for the pseudo-second-order kinetic model, Fig. 4.6) [95, 96]. A UV–visible spectrophotometer was used to keep track of the solution’s adsorption peaks as Cr(VI) was being removed. Cr2 O7 2− in solution was attributed to peaks at 257 nm and 360 nm (Fig. 4.7) [98, 99]. During the elimination of Cr(VI), the strength of these typical peaks dramatically dropped. Peaks at 276 ± 2 nm and 372 nm, respectively, were redshifted simultaneously and might be attributed to the CrO4 2− signal [99, 100]. The pH of the solution noticeably rose when the protons were used for Cr(VI) reduction (Fig. 4.8). The elimination of Cr(VI) then caused Cr2 O7 2− to undergo pH-dependent conversion to CrO4 2− . By using XRD and XPS, the chemical change in Cr(VI) removal was also examined. Before the removal of Cr(VI), three valence states of Fe species were present on ZVI/HAP, as shown by the high-resolution Fe 2p spectra (Fig. 4.9a).. Fe0 is responsible for a binding energy of 706.2 eV peak [101, 102], which can serve as active sites for the capture of Cr(VI). The peaks of 710.71 and 723.82 eV are connected to the 2p3/2 and 2p1/2 orbitals of Fe(III), respectively. [103], while the peaks at 713.06 and 726.28 eV are related to the binding energies of Fe(II) 2p3/2 and 2p1/2, respectively [104]. The integral area was used to calculate the molar ratio of Fe0 /Fetotal, which is 20.3%. After the reaction, the unique peak of Fe0 at 706.2 eV disappears. With this occurrence, the diffraction peak of Fe0 in the XRD pattern (2θ = 44.7°) similarly disappeared, suggesting that Fe0 may have been involved in the redox reaction during the elimination of Cr(VI). However, the unique peaks of Fe(II) and Fe(III) remain (Fig. 4.9a). the molar ratio of Fe(II)/Fetotal decreases from 46.7 to 43.8% and the molar ratio of Fe(III)/Fetotal

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Fig. 4.5 a XRD patterns of HAP and ZVI/HAP, b TEM image of HAP, c TEM image of ZVI/HAP, d TEM-coupled mapping images of ZVI/HAP, e Cr(VI) removal by HAP and ZVI/HAP. The initial concentrations of ZVI/HAP (or HAP) and Cr(VI) were 2 g L−1 and 100 mg L−1 , respectively. Reprinted from Yang et al. [97] Copyright 2021, with permission from Elsevier

Fig. 4.6 a Plot of ln(qe− qt ) versus time for Cr(VI) removal by ZVI/HAP. b Plot of t/Q versus time for Cr(VI) removal by ZVI/HAP. The initial concentrations of ZVI/HAP and Cr(VI) were 2 g L−1 and 100 mg L−1 , respectively. The initial pH was 5.25, and the temperature was 30 °C. Reprinted from Yang et al. [97] Copyright 2021, with permission from Elsevier

increases from 34 to 56.2%. According to these findings, species Fe0 and Fe(II) give electrons, which are then further converted to Fe(III) [105, 106]. The high-resolution O 1 s spectra (Fig. 4.9b) show that Fe2 O3 and FeO(OH) are the two forms of iron oxide species (OH) present. Thus, the intensity of Fe2 O3 (2θ = 35.65°) and FeO(OH) (2θ = 57.16°) increases in the XRD pattern (Fig. 4.10), confirming the oxidation of Fe0 and Fe(II) to Fe2 O3 and FeO(OH) during the removal of Cr(VI).

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Fig. 4.7 UV–vis absorption of the Cr(VI) solution during the reaction. The initial concentrations of ZVI/HAP and Cr(VI) were 2 g L−1 and 100 mg L−1 , respectively. The initial pH was 5.25, and the temperature was 30 °C. Reprinted from Yang et al. [97] Copyright 2021, with permission from Elsevier

Fig. 4.8 The pH changes as a function of reaction time during Cr(VI) removal. Reprinted from Yang et al. [97] Copyright 2021, with permission from Elsevier

The newly identified Cr 2p peak of 576.97 eV in the investigated spectrum (Fig. 4.11) supports the insoluble Cr deposited on the surface of ZVI/HAP particles as the cause of Cr species transfer [101, 107]. The high-resolution Cr 2p spectrum (Fig. 4.2c) can be subdivided into two peaks, Cr(III) (93%) and Cr(VI) (7%), indicating that most of the Cr(VI) is converted to Cr(III) [109]. All of the incorporated hydroxides (52%) and oxides (41%) for Cr(III) were found. As a result, Cr(VI) is reduced by ZVI/HAP by the transfer of electrons from Fe0 and Fe(II) to Cr(VI) (Eq. 4.28, 4.29). The HSC Chemistry 9 Reaction Module calculates the Gibbs free energy (G0) at various temperatures. The reduction of Cr(VI) by Fe2+ and Fe0 produced negative values of G0 at all temperatures (Fig. 4.12), suggesting that the reduction of Cr(VI) is spontaneous. Furthermore, based on the development of a new peak at 35.5° in the XRD pattern, it is further proposed that Fe0 and Fe(II) of ZVI/ HAP reduce and mineralize Cr(VI) to FeCr2 O4 − a natural mineral (Fig. 4.10). Thus,

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Fig. 4.9 a High-resolution XPS of Fe 2p of ZVI/HAP before (up) and after Cr(VI) removal (down), b high-resolution XPS of O 1 s of ZVI/HAP before (up) and after Cr(VI) removal (down), c Highresolution XPS of Cr 2p of ZVI/HAP after Cr(VI) removal. Reprinted from Yang et al. [97] Copyright 2021, with permission from Elsevier Fig. 4.10 XRD patterns of ZVI/HAP before and after Cr(VI) removal. Reprinted from Yang et al. [97] Copyright 2021, with permission from Elsevier

by loading HAP, the aggregation of ZVI can be reduced and thus the removal of Cr(VI) can be improved. + 3+ Fe0 + CrO2− + Fe3+ + 4H2 O 4 + 8H → Cr

(4.28)

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4 Mechanism of Chemical Reduction of Cr(VI)

Fig. 4.11 XPS survey spectrum of ZVI/HAP after Cr(VI) removal. Reprinted from Yang et al. [97] Copyright 2021, with permission from Elsevier

Fig. 4.12 Gibbs free energy versus temperature for Cr(VI) reduction by Fe0 and Fe2+ . Reprinted from Yang et al. [97] Copyright 2021, with permission from Elsevier

+ 3+ 3Fe2+ + CrO2− + 3Fe3+ + 4H2 O 4 + 8H → Cr

(4.29)

4.2.2 Enhancement of Coexisting Cations on Cr(VI) Removal The selectivity and adsorption capacity of ZVI for heavy metal cations is usually very low. The removal of heavy metal cations by commercial ZVI particles was less than 50% within 24 h (Fig. 4.13).ZVI removed 43.0% of Cu2+ from the coexisting Cr-Cu system, but only 30.4% of Cr(VI) (Fig. 4.14), which was even lower than that in the single Cr system (36.9%). Hence, the coexisting cations did not exhibit any improvement in ZVI’s reactivity for the removal of Cr(VI). According to reports, HAP has

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Fig. 4.13 Metal cation removal by commercial ZVI particles in a single system. The initial concentrations of commercial ZVI and cations (Cu2+ , Co2+ , or Ni2+ ) were 2 g L−1 and 50 mg L−1 , respectively. The initial pH was 5.25, and the temperature was 30 °C. Reprinted from Yang et al. [97] Copyright 2021, with permission from Elsevier

a good affinity for metal cations [93]. Here, pure HAP could almost completely adsorb all of the Cu2+ in 60 min, which was a far faster rate than that of Co2+ and Ni2+ (Fig. 4.15a). Additionally, when all of the metal cations were present, HAP absorbed Cu2+ preferentially (Fig. 4.15b). Hence, in the coexisting Cr–Cu system, HAP with strong Cu2+ adsorption activity will enhance the interaction between ZVI and Cu2+ . As a result, the coexisting Cr-Cu solution’s Cu2+ content substantially dropped. Within 120 min, almost 99.88% of Cu2+ was eliminated (Fig. 4.16). Also, the impact of coexisting metal cations (Cu2+ , Co2+ , and Ni2+ ) on the elimination of Cr(VI) was assessed. In the presence of metal cations, ZVI/HAP totally eliminated all of the Cr(VI), whereas the single Cr system removed 95.05% of the Cr(VI) (Fig. 4.17). As a result, in the coexisting system, ZVI/HAP displayed a somewhat superior performance of Cr(VI) elimination. In contrast, the chromium elimination rate was significantly higher for the coexisting system (87.13–96.53%) than for the single chromium system (68.67%) within 9 h. The corresponding apparent rate constants were 2.8 to 1.6 times higher than those of the single chromium system (4.68 × 10–3 min−1 ), ranging from 7.4 × 10–3 to 12.90 × 10–3 min−1 (Fig. 4.18). Thus, the coexisting metal cations greatly accelerated the elimination of hexavalent chromium. When the apparent rate constant is logarithmically related to the inverse of the corresponding temperature, the activation energy for the removal of hexavalent chromium can be determined using the Arrhenius equation (Eq. 4.30). lnKobs = −Ea /RT + lnA

(4.30)

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Fig. 4.14 Heavy metal removal by commercial ZVI particles. The initial concentrations of commercial ZVI and Cr(VI) were 2 g L−1 and 100 mg L−1 , respectively. The initial concentration of Cu2+ was 50 mg L−1 . The initial pH was 5.25, and the temperature was 30 °C. Reprinted from Yang et al. [97] Copyright 2021, with permission from Elsevier

Fig. 4.15 Metal cation removal by HAP in a single system (a) and in a combination system (b). The initial concentration of cations (Cu2+ , Co2+ , or Ni2+ ) was 50 mg L−1 in a single system. The initial concentrations of cations (Cu2+ , Co2+ , and Ni2+ ) were 50 mg L−1 in the combination system. The dosage of HAP was 2 g L−1 , the initial pH was 5.25, and the temperature was 30 °C. Reprinted from Yang et al. [97] Copyright 2021, with permission from Elsevier

where Kobs is the ideal gas constant (8.314 kJ mol−1 K−1 ), A is the frequency factor, Ea is the activation energy (kJ mol−1 ), R is the chemical reaction rate constant (min−1 ), T is the temperature, and (K). The calculated Ea for Cr(VI) removal in a single chromium system was 68.54 kJ mol−1 , indicating that the removal of Cr(VI) was controlled by reaction rather than by diffusion (Ea < 10–13 kJ mol−1 ) [108]. In the system present together, Ea decreases to 51.44–57.17 kJ mol−1 . Thus, the coexisting cations decrease Ea by 16.6–24.9%, which makes it easier to get rid of Cr(VI). This acceleration efficiency proceeds sequentially from Cu2+ to Co2+ to Ni2+ (Fig. 4.19).

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Fig. 4.16 Cu2+ removal by ZVI/HAP during Cr(VI) removal. The initial concentrations of Cr(VI) and Cu2+ were 100 mg L−1 and 50 mg L−1 , respectively. The initial pH was 5.25, and the temperature

Fig. 4.17 a Cr(VI) removal in the presence of cations by ZVI/HAP, b high-resolution XPS of Cu 2p of ZVI/HAP after Cr(VI) removal. The initial concentrations of Cr(VI) and cations were 100 mg L−1 and 50 mg L−1 , respectively. Reprinted from Yang et al. [97] Copyright 2021, with permission from Elsevier

The sharpest diffraction peak of FeCr2 O4 (2θ = 35.5° in Fig. 4.20) was found in the presence of Cu2+ , which showed that Cu2+ had the biggest effect on the reduction and mineralization of Cr(VI). 50 mg L−1 was the best amount of Cu2+ . When the concentration was higher than 50 mg/L, the increase in Cr(VI) removal was clearly less (Fig. 4.21).

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Fig. 4.18 Plots of ln(qe− qt ) versus time for Cr(VI) removal by ZVI/HAP in the presence of cations. The initial concentrations of Cr(VI) and cations were 100 mg L−1 and 50 mg L−1 , respectively. The dosage of ZVI/HAP was 2 g L−1 , the initial pH was 5.25, and the temperature was 30 °C. Reprinted from Yang et al. [97] Copyright 2021, with permission from Elsevier

4.2.3 In Situ Synthesized Fe-Cu Bimetal During Cr(VI) Removal The amount of Cu2+ decreases rapidly when Cr(VI) is removed from the Cr–Cu system. Moreover, the appearance of a new peak at 934.12 eV in the investigated spectrum (Fig. 4.22) indicates that Cu was added to the surface of ZVI/HAP during this process. In more detail, the deposited Cu is composed of 38.43% Cu0 species (at 932.78 and 952.44 eV), 10.32% Cu+ species (at 932.45 eV) and 51.25% Cu2+ species (at 934.45 and 954.56 eV) (Fig. 4.23) [109]. During the removal of Cr(VI), the above experiments showed that Cu0 formed in situ on the surface of ZVI/HAP via reduction reactions (Eq. 4.31). The negative values of ΔG at all temperatures (~ − 140 kJ mol−1 , 0–100 °C) confirm that the in situ deposition of Cu0 tends to occur on its own. The isoelectric point of ZVI/HAP increases sharply from 1.59 to 5.02 after Cu0 deposition (Fig. 4.24). Over a range of pH values, a high isoelectric point generally results in a more stable material and increases the electrostatic attraction with the anion [101]. Therefore, the introduction of copper effectively changes the charge distribution of ZVI/HAP and alleviates its electronegativity, thus improving the Cr(VI) capture efficiency. Fe0 + Cu2+ → Cu0 + Fe2+

(4.31)

TEM was used to observe the shape of ZVI/HAP after removal of Cr(VI) to find out more about how Cu0 deposition affects Cr(VI) removal (Fig. 4.25). In the single Cr system, some flake deposits adhered to the surface of the ZVI/HAP. The chromium

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Fig. 4.19 Plots of ln(qe− qt ) versus time for Cr(VI) removal by ZVI/HAP at different temperatures: a single Cr system, b coexisting Cr–Cu system, c coexisting Cr–Co system, and d coexisting Cr– Ni system. The initial concentrations of Cr(VI) and cations were 100 mg L−1 and 50 mg L−1 , respectively. The dosage of ZVI/HAP was 2 g L−1 , and the initial pH was 5.25. e Arrhenius plots of Cr(VI) reduction in different systems. Reprinted from Yang et al. [97] Copyright 2021, with permission from Elsevier

in these flake deposits was the same (Fig. 4.25a). This is probably due to the fact that polymeric Cr(III) hydroxides form after Cr(VI) is removed and act as passivation layers preventing Cr(VI) from being removed again. On the other hand, in the Cr-Cu system, the surface of ZVI/HAP shows irregular nanoparticles. When comparing the elemental maps in Fig. 4.25b, the area covered by Cr in the Cr–Cu system is in good

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Fig. 4.20 XRD patterns of ZVI/HAP after Cr(VI) removal in the presence of cations. Reprinted from Yang et al. [97] Copyright 2021, with permission from Elsevier

Fig. 4.21 Effect of Cu2+ concentration on Cr(VI) removal in 2 h. The initial concentration of Cr(VI) was 100 mg L−1 . The dosage of ZVI/HAP was 2 g L−1 , the initial pH was 5.25, and the temperature was 30 °C. Reprinted from Yang et al. [97] Copyright 2021, with permission from Elsevier

agreement with the area covered by Fe. Meanwhile, Cu is still distributed in the same way as Fe. It is believed that the in situ fabricated Fe–Cu bimetal anchors the Cr and leads to the reduction and precipitation of Cr(VI) on the ZVI/HAP surface. This view was supported by the fact that the concentration of Cr(III) gradually increased and then decreased with the removal of Cr(VI) (Fig. 4.26). Then, the structure of the composite oxide nanoparticles was loose and filled with pores (Fig. 4.25b). Thus, the surface area of ZVI/HAP after removal of Cr(VI) (97.77 m2 g1 ) is significantly larger than that of fresh ZVI/HAP (29.03 m2 g−1 ). This rough surface may make it easier for Cr(VI) to be carried out by adsorption and electron transfer. Thus, the above analysis suggests that “in situ synthesis” of Fe-based bimetals can be made during the reduction process, and this bimetallic strategy largely activates the surface passivated ZVI to help remove more Cr(VI).

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Fig. 4.22 XPS survey spectrum of ZVI/HAP after Cr(VI) removal in the coexisting Cr–Cu system. Reprinted from Yang et al. [97] Copyright 2021, with permission from Elsevier

Fig. 4.23 Gibbs free energy versus temperature for each reaction. Reprinted from Yang et al. [97] Copyright 2021, with permission from Elsevier

4.2.4 Electrochemical Behaviors and Contribution of in Situ Synthesized Fe–Cu Bimetal on Cr(VI) Removal It is noteworthy that the rapid removal of Cr(VI) (the removal slope of Cr(VI) is steeper in the Cr–Cu system) occurs simultaneously with the deposition of Cu0 in the first 120 min. Based on how the Fe–Cu bimetal formed, it seems likely that a galvanic couple could form between the Cu0 that was deposited in place and the Fe0 . Then, the galvanic effect could speed up the transfer of electrons from the iron to the chrome, which would make it much easier to get rid of the Cr(VI). Online electrochemical analysis techniques such as linear sweep voltammetry (LSV), Tafel scan, electrochemical impedance spectroscopy (EIS) and open circuit potential were used to investigate the electrochemical behavior of ZVI/HAP in the removal

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Fig. 4.24 Zeta potentials of ZVI/HAP and Cu0 -deposited ZVI/HAP at different pH values. The initial concentration of Cr(VI) was 100 mg L−1 in the single system, and the initial concentrations of Cr(VI) and Cu2+ were 100 mg L−1 and 50 mg L−1 in the Cr–Cu system, respectively. The dosage of ZVI/HAP was 2 g L−1 , and the temperature was 30 °C. Reprinted from Yang et al. [97] Copyright 2021, with permission from Elsevier

Fig. 4.25 a TEM-coupled mapping images of ZVI/HAP in a single Cr system after Cr(VI) removal, b TEM-coupled mapping images of ZVI/HAP in a coexisting Cr-Cu system after Cr(VI) removal. Reprinted from Yang et al. [97] Copyright 2021, with permission from Elsevier

of Cr(VI). This was done to confirm that Fe-Cu bimetals accelerate the transfer of electrons (OCP). LSV was used to study the speed of corrosion, as shown in Fig. 4.27a. When the potential exceeds 0.06 V, spontaneous corrosion causes a large rise in current. Once Fe-Cu bimetals are formed on the surface of ZVI/HAP, electron transfer occurs at a lower potential, surging to a higher current density than that of ZVI/HAP. It is believed that putting down Cu0 accelerates the onset of the electron transfer process. In the Cr-Cu system, Cu0 formed rapidly on the surface of ZVI when ZVI was submerged in Cu2+ solution for 120 min. For the Tafel scan, EIS and OCP tests, the time of soaking ZVI/HAP was different. Figure 4.27b shows that the fresh ZVI/HAP had a free corrosion potential of less than 0.82 V. As the time spent in the Cu2+ solution went up, the corrosion potential kept going down. After 80 min, the corrosion potential was the lowest it could be (0.95 V). This 130 mV

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Fig. 4.26 Change in total Cr, Cr(VI) and Cr(III) in the coexisting Cr–Cu system. The initial concentrations of Cr(VI) and Cu2+ were 100 mg L−1 and 50 mg L−1 in the Cr–Cu system, respectively. The dosage of ZVI/HAP was 2 g L−1 , and the temperature was 30 °C. Reprinted from Yang et al. [97] Copyright 2021, with permission from Elsevier

difference in corrosion potential showed that the Cu0-deposited ZVI/HAP corroded in a very different way, which could help speed up the transfer of electrons during Cr(VI) reduction. EIS was used to differentiate the charge transfer resistance (Rct) of Cu0 -deposited ZVI/HAP. (Fig. 4.28). The Rct of ZVI/HAP decreases sharply during Cu0 deposition, which indicates that Cu0 -deposited ZVI/HAP has better electrical conductivity to help electron movement. With the deposition of Cu0 , the negative OCP of the glassy carbon electrode (GCE) with ZVI/HAP coating (ZVI/HAP-GCE) decreased sharply from 0.73 to 0.78 V (Fig. 4.27c). Since the value of OCP shows the likelihood of a material losing electrons through corrosion, more negative potentials suggest that Cu0 -deposited ZVI/HAP may lose electrons faster than new ZVI/HAP. To further understand how the “in situ synthesized” Cu-Fe bimetal affects the corrosion reactivity of ZVI/HAP, real-time OCP was measured in situ during the corrosion process (Fig. 4.27d). When Cu2+ solution was added to the stable system, the OCP of the ZVI/HAP-GCE electrode increased rapidly. This may be due to spontaneous Fe0 corrosion, in which electrons are donated from Fe0 for the reduction of Cu2+ . Then, the OCP slowly decreases with the accumulation of Cu0 . Once a Cu-Fe bimetal is formed on the surface of ZVI/HAP, the addition of Cr(VI) causes the OCP to rise sharply. Similarly, when a solution of Cr2 O7 2− and Cu2+ was added, the OCP rose a lot. On the other hand, in the absence of Cu0 , the OCP rises only slightly. This difference may be due to the enhanced electron transfer from Fe to Cr by the deposited ZVI/HAP with Cu0 . Thus, the Cu-Fe bimetal made in situ accelerated the electron transfer and helped the reduction of Cr(VI).Then, it was looked at how the facilitated electron transfers affected the removal of Cr(VI). By corrosion, the core of ZVI can be made Fe(II) in the form of Fe/Fe2 O3 [107]. This Fe(II) can change Cr(VI) to Cr(III), and the rest of the reduction comes from the direct transfer of electrons from the conduction band of Fe0 to Cr(VI). Here, 0.1 g of 1,10-phenanthroline, which is a Fe(II) quencher [112], was added to the reaction system to find out how much Fe(II) helped get rid of Cr(VI) in the Cu–Fe bimetal ZVI/HAP system. The removal of Cr(VI) was stopped when 1,10-phenanthroline was added to both the single Cr and

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Fig. 4.27 a LSV of ZVI/HAP and Cu0 -deposited ZVI/HAP electrodes. b Tafel scans of Cu0 deposited ZVI/HAP with different Cu2+ soaking times. c OCPs of Cu0 -deposited ZVI/HAP electrodes with different Cu2+ soaking times (inset: OCP curves of Cu0 -deposited ZVI/HAP with different Cu2+ soaking times). d OCP curves of ZVI/HAP electrodes. Reprinted from Yang et al. [97] Copyright 2021, with permission from Elsevier

the Cr–Cu systems (Fig. 4.29). In the monochromium system, the removal of Cr(VI) was slowed down by 57.54%. This means that Fe(II) released by the corrosion of Fe0 is responsible for 57.54% of the reduction of Cr(VI). In the case of the Cr-Cu system where it was already present, it reached 79.54%, indicating that the addition of Cu2+ made the contribution of Fe(II) to the removal of Cr(VI) rise. This result showed that Cr(VI) was taken out of the Cr–Cu system by Fe(II) that was released when Fe0 corroded. Thus, Fe-Cu bimetals on ZVI/HAP accelerate the electron transfer of Fe0 corrosion and release more Fe(II), which contributes to the reduction of Cr(VI).

4.2.5 Enhanced Cr(VI) Removal Mechanism by in Situ Synthesized Fe–Cu Bimetal Based on the above results and discussion, a possible mechanism is proposed to explain how the in situ fabricated Fe-Cu bimetal can help remove Cr(VI) (Fig. 4.30). In general, Cr(VI) was removed through three processes: adsorption, reduction, and

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Fig. 4.28 Nyquist plots of ZVI/HAP with different Cu2+ soaking times. a 20 min, b 30 min, c 40 min, d 80 min, e 120 min, f 360 min. Reprinted from Yang et al. [97] Copyright 2021, with permission from Elsevier

Fig. 4.29 The inhibitory efficiency of 1,10-phenanthroline on Cr(VI) removal. The initial concentrations of Cr(VI) and cations were 100 mg L−1 and 50 mg L−1 , respectively. The dosage of ZVI/ HAP was 2 g L−1 , the initial pH was 5.25, and the temperature was 30 °C. The dosage of 1,10phenanthroline was 0.1 g. Reprinted from Yang et al. [97] Copyright 2021, with permission from Elsevier

co-precipitation. So, the rate-limiting step is reduction, which determines how well Cr(VI) is removed. In the single Cr system, Fe0 directly transfers multiple electrons from its conduction band to Cr(VI) and slowly corrodes in the aqueous environment, releasing Fe(II) for further reduction of Cr(VI), which results in a low Cr(VI) removal rate. In the Cr-Cu system, there are three redox reactions going on simultaneously. At

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4 Mechanism of Chemical Reduction of Cr(VI)

Fig. 4.30 Scheme of enhanced Cr(VI) removal mechanism of in situ synthesized Fe–Cu bimetal. Reprinted from Yang et al. [97] Copyright 2021, with permission from Elsevier

the beginning of the Cr(VI) removal process, HAP helps Cu2+ to stick to the surface of ZVI/HAP. Then, Cu2+ acquires electrons from the rusted Fe0 and puts them on the surface of ZVI as Cu0 . Secondly, both Fe0 and Fe(II) dissolved from ZVI make the Cr less stable. Here, microscopic galvanic couples were made between Cu0 that was deposited in place and Fe0 . Cu0 was the inert electrode. In each electrochemical cell, electrons move rapidly from the anode (Fe0 ) to the cathode (Cu0 ). This rapid movement of electrons allows the reaction to occur more quickly [113]. In general, the large number of microscopic electric couples overcomes the self-inhibition of electron transfer in the reduction reaction. This accelerated the release of Fe(II). At the same time, the in situ fabricated Fe–Cu bimetal immobilized Cr and led to the reduction and precipitation of Cr(VI), forming a loose porous structure on the ZVI/ HAP surface, making more Cr(VI) more readily available for uptake and reduction. With the development of electroplating, mining and metallurgical industries, wastewater usually contains more than one metal ion, such as Cu2+ , Co2+ , Ni2+ , etc. Therefore, it is useful to remove multiple metal ions simultaneously in the process of cleaning water. Here, ZVI/HAP was also used in a polycation complex system of Cr2 O7 2− , Cu2+ , Co2+ and Ni2+ . Spontaneous reduction and deposition of Co2+ and Ni2+ exhibited negative values of ΔG at all temperatures (~ −30 kJ mol−1 , 0–100 °C, Figure S16). Therefore, the combined effect of multiple cations contributes to the effectiveness of Cr(VI) removal (Fig. 4.31). In the presence of other cations, the Cr(VI) removal Ea decreased by 16.6% and more than 90% of Cr(VI) disappeared within 180 min (Fig. 4.32). More importantly, almost all the co-existing Cu2+ , Co2+ and Ni2+ were removed at the same time. Considering the initial concentration of heavy metals, the amount of material used and the length of treatment time, ZVI/ HAP in this study showed good results in the removal of Cr(VI). This iron-based bimetal can be used “in situ” and can be used to clean up combined contaminants.

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Fig. 4.31 a Plots of ln(qe− qt ) versus time for Cr(VI) removal by ZVI/HAP at different temperatures in the coexisting Cr–Cu–Ni–Co system. The initial concentrations of Cr(VI) and cations were 100 mg L−1 and 50 mg L−1 , respectively. The dosage of ZVI/HAP was 2 g L−1 , and the initial pH was 5.25. b Arrhenius plots of Cr(VI) reduction in different systems. Reprinted from Yang et al. [97] Copyright 2021, with permission from Elsevier

Fig. 4.32 Removal of Cr(VI) and metal cations by ZVI/HAP in coexisting Cr-Cu-Ni-Co system. The initial concentrations of Cr(VI) and cations were 100 mg L−1 and 50 mg L−1 , respectively. The dosage of ZVI/HAP was 2 g L−1 , and the initial pH was 5.25. Reprinted from Yang et al. [97] Copyright 2021, with permission from Elsevier

4.3 Interaction Between Pyrite and Zerovalent Iron that Has a Higher Ability to Reduce Through Fe(II) Regeneration One of the most commonly employed materials in pollutant sequestration is nanoscale zero valent iron (ZVI), which has great environment compatibility and a moderately resilience. [85, 97, 110]. Unfortunately, considerable surface passivation and nonselective reduction severely limit its applicability (i.e., H2 O reduction).

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4 Mechanism of Chemical Reduction of Cr(VI)

Its lowering capability is depleted by this indiscriminate reaction, which also shortens its reactive lifetime [111, 112]. Sulfidated ZVI (SZVI) has recently come to light as a viable strategy to increase ZVI’s reactivity and selectivity [113–115]. The Fe0 core and FeSx shell are formed as a result of this strategy. To reduce indiscriminate H2 evolution and the H2 O reduction reaction, the incorporation of sulfur may enable electron transport from the Fe0 core to the impurities and restrict the adsorption location of atomic hydrogen [116–119]. However, the main electron donor for SZVI’s reductive elimination of contaminants, Fe(II) [120], still has a rather poor capacity for regeneration [120–122]. The mechanism of SZVI’s increased responsiveness has remained a mystery [123, 124]. Generally speaking, the removal of pollutants could be considerably aided by the regeneration of Fe(II) from Fe3+ by Fe0 corrosion [125–127]. However, the precipitation of secondary FeSx as well as the dense base layer of FeSx were detrimental to the rate of corrosion of the underlying Fe0 core [127–129]. The surface FeSx layer from SZVI substantially inhibits the regeneration of Fe(II) [129]. Also, its practical applicability is restricted by the limited pH range that is usable (only acidic circumstances). Due to the composition of iron (hydrogen) oxides, the electronic delivery from ZVI to the target pollutants and the release of Fe(II) would be impeded when the pH is above 5.00 [129]. The Fe0 core-FeSx shell configuration thus appears to have the potential to reduce the SZVI’s usage efficiency. The aforementioned restriction of SZVI seems to be overcome if we build a structure using reductant FeSx as the primary component and fine-grained ZVI evenly distributed throughout.Minerals like pyrite (FeS2 ) are common in tailings or the crust [130, 131]. It has a considerable reducing ability and is an efficient pH buffer thanks to the composition of reductive groups, Fe(II), and S2 2− . Pyrite is a more hydrophobic substance in terms of electron selectivity and has a lower affinity for H because of its greater S:H ratio. Moreover, pyrite and ZVI exhibit a collaborative impact on decreasing to release additional Fe because FeS2 can combine with Fe3+ , which is created via pollutant reduction by ZVI, to produce Fe2+ (II) [132, 133]. FeS2 is the perfect material for sulfurizing and dispersing ZVI in order to sequester contaminants. Nevertheless, natural pyrite has l low-reducing activity in neutral or alkaline settings due to its large grain size, solid crystal construction, and severe shallow passivation [134, 135]. Amorphous FeS2 , on the other hand, is far more effective and has a larger specific surface area [136]. SZVI is prepared using two basic techniques: ball milling and chemical coprecipitation. Mechanical ball milling (BM), in contrast to chemical coprecipitation, is distinguished by effectiveness, economy, and simplicity of use. Large particle fragmentation, amorphization, and the creation of composites with evenly distributed amounts of particular elements have all been made possible [137–139]. Thus, it is probable that the BM method is effective in activating natural pyrite and sulfidating ZVI.

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213

4.3.1 Incorporation of Sulfur into Fe0 Mechanical ball milling has recently emerged as a cost-effective technique for converting micron-scale zero-valent iron to nanoscale iron [102]. In contrast, because of its ductility and the interaction between its particles, pyrite might still be micronlevel after 24 h of ball mills [140]. ZVI particles in this case showed an uneven spheres on the surface of pyrite following BM to produce ZVI-loaded pyrite (FeS2 / ZVI) (Fig. 4.33). With rising FeS2 /ZVI molar ratios, the particle size of the readymade FeS2 /ZVI reduced, and the surface is rougher. The findings revealed that certain surface areas increased when FeS2 /ZVI molar ratios rose (Table 4.4). With a FeS2 / ZVI molar ratio of 9:1, the highest specific surface area was 10.9 m2 g−1 . As a result, numerous nanoscale ZVI particles clung to pyrite surface during the BM process to increase the specific surface area. By providing more reactive sites, this enhanced specific surface area could significantly increase their reactivity for the elimination of Cr(VI). Comparing the O content of FeS2 /ZVI to severe passivation of the pure pyrite (85.28%), 3.67% was found (Fig. 4.34). BM successfully prevented the materials’ further passivation as a result. Characterization of the FeS2 /as-prepared ZVI’s characteristics. Figure 4.33 displays the XRD spectra of FeS2 /ZVI at various molar ratios. Throughout the milling process, no new phases emerged. Fe0 was given credit for the diffraction peaks at 44.6° and 64.9°. The pattern also revealed the typical FeS2 2 peaks at 28.5°, 33.1°,

Fig. 4.33 SEM images of FeS2 /ZVI with molar ratios of 5:5 (a), 7:3 (b), and 9:1 (c); (d) XRD patterns of FeS2 /ZVI with different molar ratios (10:0–0:10) and (e) lattice constants of the Fe BCC structure according to XRD Rietveld refinement. Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022

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4 Mechanism of Chemical Reduction of Cr(VI)

Table 4.4 Specific surface area, pore diameter, and pore volume of FeS2 /ZVI with different molar ratios FeS2 /ZVI molar ratio 10:0 9:1

Specific surface area (m2 g−1 ) 8.19 10.9

Pore diameter (nm)

Pore volume (cm3 g−1 )

11.4

0.0283

8.22

0.0259

7:3

7.04

9.92

0.0204

5:5

7.45

8.92

0.0208

3:7

7.78

8.33

0.0166

1:9

5.12

6.26

0.00666

0:10

1.07

5.55

0.00188

Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022

Fig. 4.34 SEM–EDS images of FeS2 /ZVI (molar ratio 9:1). Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022

37.1°, 40.8°, 47.4°, 56.3°, 59.0°, 61.7°, 64.3°, 76.6°, and 78.9°. The peak breadth of Fe0 (44.6°) grew as the FeS2 /ZVI molar ratio increased, while its intensity decreased. When the FeS2 /ZVI molar ratio reached 9:1, the diffraction peak at 64.9° vanished in the interim. These findings suggested that Fe0 ’s crystalline phase may have been damaged during the milling procedure. The lattice constant of Fe0 in FeS2 /ZVI at various molar ratios was calculated using contour fitting and refinement (Rietveld) in order to better investigate the fluctuation in the Fe0 crystalline phase. Figure 4.33e displays the Fe (110) peak’s expanded profile. The peak diffraction of the Fe(110) crystal plane in FeS2 /ZVI redshifted to 44.7° in comparison to pure ZVI, and the grid spacing of Fe(110) was substantially smaller. It could be explained by the inclusion of S during milling into the body-centered cubic grid of iFe. Due to its high electron conductivity, FeS2 /ZVI with a 9:1 moore’s ratio has the widest grid spacing, which may make it easier for electrons to go from the iron core to the hexavalent chromium oxide ion [118]. At the same time, the (311) crystal plane with lattice spacing of 0.16 nm in highresolution TEM images corresponded to FeS2 in FeS2 /ZVI (Fig. 4.35a–c, Fig. 4.36). It is important to note that some lattice dislocation and deformation can be observed at the edges of the crystal plane. This disordered alignment of lattice stripes suggested

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Fig. 4.35 HRTEM image (a), fast Fourier transform (FFT) patterns (b and d), and inverse FFT images (c and e) of the FeS2 /ZVI sample. Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022

that the atoms left their balanced posture to a substable state with high reactivity, thus increasing the free energy of FeS2 /ZVI. As indicated in the S 2p XPS spectra (Fig. 4.37), four forms of sulfur speciation were found on FeS2 /ZVI, including SO4 2− at 170.03 and 168.95 eV, S0 at 165.01 eV, Sn 2− at 164.03 eV, and S2 2− at 162.87 eV. The sulfur types in FeS2 can only classified as S2− and S2 [127, 132, 142]. The newly found diffraction peaks may be connected to the fact that the FeS2 supplementary encouraged the formation of an amorphous FeSx layer from Fe0 [142]. It consequently appeared as a cystic formation in the FeS2 /ZVI TEM pictures (Fig. 4.38). The FeSx layer is an effective electron conductor that favors electron transport from Fe0 to Cr(VI) at the surface and has a favorable affinity for anions that makes it easeful to concentrate Cr(VI) anions on the FeS2 /ZVI surface [143]. A Mössbauer study was carried out to more clearly depict the structural characteristics of Fe in FeS2 /ZVI. The 57Fe Mössbauer spectrum revealed two doublets and one sextet, indicating that there are three iron components in the solid phase (Fig. 4.39). The doublets’ characteristics, I.S. = 0.32 mm s−1 and Q.S. = 0.59 mm s−1 for double-layered 1 and I.S. = 1.24 mm s−1 and Q.S. = 2.73 mm s−1 for doublet 2, respectively, were in agreement with Fe2+ and Fe3+ (Table 4.5). Fe0 was credited with producing the I.S., Q.S., Hhf, and sextet 1 (− 0.02 mm s−1 , − 0.05 mm s−1 , and 329.48 KOe, respectively). Fe0 , Fe2+ , and Fe3+ were found to account for 13.1, 61.2, and 25.7% of the total Fe, respectively, by fitting the peak area in the Mössbauer spectrum. Also, the peaks at 1090 cm1 in the FTIR spectra (Fig. 4.39b) were attributed to the hydroxyl groups’ bend swing (Fe–OH). The intensity of the Fe-OH obviously increased after the milling procedure. In general, the Fe-OH bond is in charge of creating internal sphere surface complexes and has a propensity to interact with HCrO4 − species [144, 145]. It is possible that the Iron (hydrogen) oxide produced by erosion will enlarge the surface area and facilitate Cr Adsorption, reduction and sedimentation (VI).

216

4 Mechanism of Chemical Reduction of Cr(VI) 100

0.16 nm

80 60 40 20 0 -20 -40 -60 -80 100 0.0

0.5

1.0

1.5

2.0

2.5

3.0

3.5

Fig. 4.36 D-spacing intensity profiles of the different regions of the FeS2 /ZVI sample. Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022

Fig. 4.37 S 2p XPS spectra of FeS2 /ZVI (molar ratio 9:1). Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022

4.3.2 Removal Activity of FeS2 /ZVI Toward Cr(VI) ZVI and pyrite are both frequently utilized for the deletion of contaminants, while they have an additive impact on the Cr(VI) reduction [132]. It is possible for hybrid materials with various FeS2 /ZVI molar ratios to react differently with Cr(VI). ZVI and FeS2 only eliminated 11.2% and 26.8% of Cr(VI) after 3 h, respectively. FeS2 / ZVI could deleted 14.9–99.9% of Cr(VI), with a higher FeS2/ZVI moore’s ratio increasing the efficacy of Cr(VI) removal (Fig. 4.40). For a FeS2 /ZVI molar ratio of 9, the greatest Cr(VI) removing was 99.9%. The reduction in Cr(VI) removal to 92.71% and 79.48%, respectively, was caused by further raising the FeS2 /ZVI

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Fig. 4.38 TEM images of FeS2 /ZVI (molar ratio 9:1). Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022

Fig. 4.39 a 57 Fe Mossbauer spectra of FeS2 /ZVI, b FTIR spectra of FeS2 , ZVI and FeS2 /ZVI before and after reaction with Cr(VI); (Particle dosage: 1 g L−1 , [Cr(VI)]0 = 50.0 mg L−1 , pH0 = 5.20, air atmosphere). Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022 Table 4.5 The 57 Fe Mossbauer parametersa of FeS2 /ZVI Component

H (KOe)

IS (mm s−1 )

QS (mm s−1 )

G/2 (mm s−1 )

Fe%

Identified state of iron

Doublet 1

0.32

0.59

0.16

61.2

Fe2+

Doublet 2

1.24

2.73

0.15

25.7

Fe3+

− 0.02

0.05

0.16

13.1

Fe0

Sextet 1

329.48

Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022 a IS: isomer shift. QS: quadrupole splitting. H: hyperfine field. G/2: half-linear

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4 Mechanism of Chemical Reduction of Cr(VI)

moore’s ratio to 10:1 and 11:1 (Table 4.6), which could be connected to the material gathering with increased pyrite loading [140]. Thus, the ideal molar ratio of FeS2 /ZVI was determined to be 9:1. Surprisingly, only 29.1% of Cr(VI) was eliminated when the ZVI particles were combined with FeS2 identical moore’s ratio of 9:1 without the BM (Fig. 4.40b). These findings suggested that the ZVI and FeS2 in FeS2 /ZVI may have synergistic effects on the elimination of Cr(VI). The material’s reactivity could be significantly impacted by the BM time. In Fig. 4.40c, the influence of BM duration on the removing of Cr(VI) by FeS2 /ZVI with a 9:1 moore’s ratio is depicted. With the BM time extended, the removing capability of Cr(VI) greatly enhanced. As the BM time exceeded 4 h, Cr(VI) removal stopped increasing, indicating that too much BM time may cause broken particles to reagglomerate and prevent the regeneration of active sites [137]. Hence, 4 h was the ideal BM time for FeS2 /ZVI.

Fig. 4.40 Influence of FeS2 /ZVI (a) molar ratio and (c) ball milling time on Cr(VI) removal; (b) efficiency of Cr(VI) removal by different materials (“FeS2 /ZVI” means the preparation of FeS2 / ZVI through ball milling. “Mixed FeS2 -ZVI” means the mixture of FeS2 and ZVI without ball milling. “BM-ZVI” means the preparation of ZVI through ball milling. “BM-FeS2 ” means the preparation of FeS2 through ball milling.); (d) The concentration of total Cr, Cr(VI) and Cr(III) during Cr(VI) removal by FeS2 /ZVI; (e) Effect of pH on Cr(VI) removal by FeS2 /ZVI and (f) the changes in pH during the removal of Cr(VI). (Particle dosage: 1 g L−1 , [Cr(VI)]0 = 50.0 mg L−1 , pH0 = 5.20, air atmosphere). Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022

Table 4.6 Cr(VI) removal by FeS2 /ZVI at different FeS2 /ZVI molar ratios Molar ratio (FeS2 :ZVI)

9:1

10:1

11:1

10:0

Cr(VI) removal rate (%)

99.9

92.71

79.48

26.8

Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022

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219

The first step in the process is to determine your targeted attendees. The second step is to determine your targeted attendees (Fig. 4.40). Then, the Cr(III) concentration, which was determined by subtracting the Cr(VI) concentration from the overall chromium concentration, grew with time. After Cr(VI) reduction, the FeS2 / ZVI system had a greater Cr(III) concentration than the combined FeS2 -ZVI system (Fig. 4.41). Meanwhile, combined FeS2 -ZVI’s surface was covered in stacked, thick precipitates (Fig. 4.42), which prevented mass and electron transport. On the surface of FeS2 /ZVI, regular and loose nanoparticles are rather evenly scattered. This porous structure might promote Cr(VI) elimination by facilitating future iron corrosion. This discovery is in line with other research showing that an increase in surface area caused by a sparse and porous structure promotes electron transport to sustain longterm reactivity [146]. T Table 4.7 lists the pseudosecond-order chemical reactions that were used to determine the speed constant (K), balanced adsorption capacity (Qe ), and R2 values. The investigation revealed that both mixed FeS2 -ZVI and FeS2 / ZVI followed pseudosecond-order kinetics in the elimination of Cr(VI) (Fig. 4.43). According to this finding, the removal of Cr(VI) was primarily a chemical absorptionreduction processes, and the reaction’s rate control steps was a chemical procedure based mostly on an electron transfer reaction. The combined FeS2 -ZVI had an equilibrium adsorption capacity of the secondary kinetic fit (Qe) of 11.95 mg g−1 . The Qe of FeS2 /ZVI, in contrast, added by 4.4 times to 52.08 mg g−1 (Table 4.7), which was significantly greater than that of the majority of known SZVI materials [45, 120, 128, 147]. In general, ZVI-based materials show a strong pH dependence [148]. An ironmediated mechanism that uses a proton-consuming action is more effective in removing Cr(VI) under acidic circumstances [149, 150]. A pH range of 3.04–10.4

Fig. 4.41 The concentrations of total Cr, Cr(VI) and Cr(III) during Cr(VI) removal by mixed FeS2 ZVI. (Particle dosage: 1 g L−1 , [Cr(VI)]0 = 50.0 mg L−1 , pH0 = 5.20, air atmosphere). Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022

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4 Mechanism of Chemical Reduction of Cr(VI)

Fig. 4.42 SEM images of a mixed FeS2 -ZVI and b FeS2 /ZVI after Cr(VI) removal. (Particle dosage: 1 g L−1 , [Cr(VI)]0 = 50.0 mg L−1 , pH0 = 5.20, air atmosphere). Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022

Table 4.7 Linear regression analysis of the kinetics of Cr(VI) removal by mixed FeS2 -ZVI and FeS2 /ZVI FeS2 -ZVI

FeS2 /Fe0

k 1 (min−1 )

0.01

0.0254

Qe (mg g−1 )

13.92

49.93

R2

0.9355

0.9405

Parameter Pseudo first-order

Pseudo second-order

k 2 (g

mg−1

min−1 )

0.006544

0.002137

Qe (mg g−1 )

11.95

52.08

R2

0.9939

0.9986

Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022

Fig. 4.43 a Cr(VI) removal by FeS2 /ZVI and mixed FeS2 -ZVI, the fitting plots of b pseudofirstorder kinetic modeling and c pseudosecond-order modeling by FeS2 /ZVI and mixed FeS2 -ZVI. (Dosage: 1 g L−1 , [Cr(VI)]0 = 50.0 mg L−1 , pH0 = 5.20, air atmosphere). Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022

was tested to see how it affected the removing of Cr(VI) by FeS2 /ZVI (moore’s ratio: 9:1, BM time: 4 h) (Fig. 4.40e). The findings demonstrated that throughout a wide pH range, starting pH had little effect on the elimination of Cr(VI) (3.04–9.20). During two hours, all 50 mg L−1 of Cr(VI) had been eliminated. FeS2 /ZVI also still reduces 86.2% of Cr(VI), which is better than the majority of SZVI, as previously reported,

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Table 4.8 Intrinsic properties of FeS2 /ZVI (molar ratio 9:1) Name

FeS2 /ZVI

Content (mg g−1 ) Fe

S

770

175

Specific surface area (m2 g−1 ) 10.9

S speciation (%)

Lattice constant (Å)

SO4 2−

S0

Sn 2−

S2 2−

50.0

9.80

12.3

27.9

2.861

Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022

even under severely non-acidic environment (pH 10.4). Additionally, a pH meter was used to track the pH during the elimination of Cr(VI). The pH in the ZVI system gradually rose, as depicted in Fig. 4.40f. However, after the addition of FeS2 , FeS2 / ZVI, and combined FeS2 -ZVI, the pH dropped significantly. The pH was then kept at a low level (below 3.8). Pyrite was able to reduce the rise in pH and demonstrated beneficial pH modification as Cr(VI) removal progressed due to H+ creation in the reaction of pyrite with Fe3+ . The pH decrease in FeS2 /ZVI was the greatest. As a result, the prolonged low pH prevented the deposition of (hydr)oxides to prevent the passivation of ZVI’s surface, and ZVI was more likely to corrode for the release of Fe2+ , which would facilitate the removal of Cr(VI). FeS2 /ZVI is therefore pHindependent and exhibits great activity in the removal of Cr(VI) over a broad pH range. In order to identify the factors that contributed to each batch of synthesized SZVI’s distinct features and performance, it is also crucial to measure and report the intrinsic attributes of per batch [151]. Table 4.8 displays the intrinsic qualities of the ideal FeS2 /ZVI.

4.3.3 Mechanism of Cr(VI) Removal by FeS2 /ZVI Using the use of XRD and XPS, the chemical conversion in Cr(VI) reduction was examined in order to better understand the removing process of Cr(VI) by FeS2 / ZVI. After Cr(VI) reduction in the ZVI system, the peaks of diffraction of Fe0 at 44.6° and 64.9° persisted (Fig. 4.44a). On the other hand, in the FeS2 /ZVI system, the typical Fe0 peaks vanished and the strength of the FeS2 diffraction peaks dramatically reduced. This could be explained by the ZVI in FeS2 /ZVI having been activated during the BM process to promote corrosion. The high-resolution Cr 2p peak reveals the presence of chromium in mixed hydroxide (66.5%), oxide (22.2%) and ferrochrome (hyer) oxide(11.3%). (Fig. 4.44b). Generally speaking, ZVI-based materials can change Cr(VI) to Cr(III) through an e-transfer directly from the Fe0 core to the Fe2+ ion, followed by reduction. The H+ ions were utilized for Cr(VI) reduction, and then the steadily rising pH would inhibit the response. This explains why the removing of Cr(VI) from ZVI is facilitated by acidsic environments. In this instance, FeS2 was added to a ZVI-based polymer. Particularly, the peak binding energy is 720.07 eV is assigned to Fe0 in the Fe 2p XPS spectra of FeS2 / ZVI (Fig. 4.44c). Whereas the peaks at 710.94 and 724.55 eV filed under the signal

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4 Mechanism of Chemical Reduction of Cr(VI)

Fig. 4.44 a XRD spectra of FeS2 , ZVI and FeS2 /ZVI before and after Cr(VI) removal; b Cr 2p XPS spectra of FeS2 /ZVI after Cr(VI) removal; c Fe 2p XPS spectra of FeS2 /ZVI before and after Cr(VI) removal. (Particle dosage: 1 g L−1 , [Cr(VI)]0 = 50.0 mg L−1 , pH0 = 5.20, air atmosphere). Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022

of Fe(III)-O 2p3/2 as well as 2p1/2, respectively, the peaks at 714.50 and 728.14 eV were ascribed to the 2p3/2 as well as 2p1/2 tracks from Fe(II)-O, respectively [157, 158]. Fe(II)-S was the source of another peak at 707.3 eV. Integral areas calculations revealed that 9.60% of the Fe species in FeS2 /ZVI were Fe0 , 24.7% were Fe(II)O, 48.1% were Fe(III)-O, and 17.6% were Fe(II)-S. In general, Fe0 and FeS2 can interact with Fe3+ to produce Fe2+ (Eqs. 4.34 and 4.35), which expedites the elimination of Cr(VI). On the surface of FeS2 /ZVI, an Fe2+ → Fe3+ → Fe2+ cycle was thus present. We measured the amounts of Fe3+ and SO4 2− in the reaction system to clarify the electronic donors for Cr(VI) reduction (Fig. 4.45). Both the FeS2 /ZVI system and the mixed FeS2 /ZVI system had a significant rise in the content of Fe3+ , which was mostly caused by the oxidation of Fe0 , Fe2+ , and FeS2 (Eqs. 4.32, 4.33, and 4.36). The mixed FeS2 -ZVI system’s Fe3+ content subsequently reached equilibrium (Fig. 4.45a). The FeS2 /ZVI system’s Fe3+ concentration might be raised by continually reducing Cr(VI) with the regenerated Fe(II), in contrast. The Fe2+ → Fe3+ → Fe2+ cycle on the surface of FeS2 /ZVI was verified by the gradual rise in Fe3+ concentration and the increased concentration of SO4 2− . It’s also important to note that the reaction between Fe0 and FeS2 produces the + H ions. The reaction equilibrium of Cr(VI) reduction (Eq. 4.36) may move to the right as a result of the rise in H+ . As a result, the addition of FeS2 improved the pH adaptability of ZVI-based materials. The molar ratio of Fe0 dropped to 9.40% after response. While the moore’s ratio of Fe(III) climbed to 59.3%, the moore’s ratio of Fe(II) (containing Fe(II)-O and Fe(II)-S) declined to 18.5% and 12.8%, respectively. According to these results, Fe0 and Fe2+ supplied electrons for the reduction of Cr(VI) and then changed into Fe3+ . + 3+ 2Fe0 + Cr2 O2− + 2Fe3+ + 7H2 O 7 + 14H → 2Cr

(4.32)

+ 3+ 6Fe2+ + Cr2 O2− + 6Fe3+ + 7H2 O 7 + 14H → 2Cr

(4.33)

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Fig. 4.45 The concentrations of a Fe3+ and b SO4 2− in the mixed FeS2 -ZVI system and FeS2 / ZVI system. (Particle dosage: 1 g L−1 , [Cr(VI)]0 = 50.0 mg L−1 , pH0 = 5.20, air atmosphere). Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022

2Fe0 + Fe3+ → 3Fe2+

(4.34)

+ FeS2 + 14Fe3+ + 8H2 O → 15Fe2+ + 2SO2− 4 + 16H

(4.35)

+ 3+ 3+ 2FeS2 + 5Cr2 O2− + 4SO2− + 19H2 O 7 + 38H → 2Fe 4 + 10Cr

(1 − x) Fe3+ + xCr3+ + 2H2 O → Crx Fe1−x OOH + 3H+

(4.36) (4.37)

The dedication of the iron species to the electron transport in the reduction of Cr(VI) was then examined. Fe2+ is strongly complexed, and 1,10-phenanthroline can prevent Fe2+ from being reduced to Cr(VI), protecting Fe2+ during Cr(VI) removal. In order to investigate the function of Fe2+ species in Cr(VI) removal experiments, 1,10-phenanthroline was introduced to the reaction as a Fe2+ quencher. In a nutshell, the reaction with 1,10-phenanthroline present was carried out at pH 3.93 in NaAcHAc buffer solution. 1,10-Phenanthroline was added in excess to totally protect Fe2+ . When 1,10-phenanthroline was present, the elimination of Cr(VI) was considerably inhibited (Fig. 4.46). Only around 10% of Cr(VI) could be deleted by mixed FeS2 ZVI, which indicated that 64.1% of Cr(VI) reduction was caused by Fe2+ corrosion from Fe0 . The FeS2 /ZVI system’s scenario dropped to 61%. FeS2 /ZVI could still delete 39% of Cr(VI), albeit the activated sulfide may diminish this amount. The iron content that was discharged from the iron-based substance was also found. In the first five minutes in the pure water system, the Fe2+ released from the mixed FeS2 ZVI system peaked at 32.46 mg L−1 and subsequently started to decline (Fig. 4.47). Yet within 30 min, the Fe2+ released from FeS2 /ZVI grew gradually but steadily, and 126 mg L−1 of Fe2+ was received, which was about 4 times that of mixed FeS2 -ZVI. FeS2/ZVI emitted nearly 2.5 times as much Fe2+ overall. FeS2 /ZVI thus inclination

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4 Mechanism of Chemical Reduction of Cr(VI)

to discharge more Fe2+ for Cr(VI) reduction than mixed FeS2 -ZVI. In general, the Fe2+ produced when iron corrodes to form iron (hydrogen) oxide may be discharged into a solution or exist as surface-bound Fe2+ . The same quantity of posted Fe2+ could only delete roughly 49% of the total Cr(VI) under the same circumstances if the number of Fe2+ released in Cr(VI)-containing solution equivalent to those discharging in Cr(VI)-free solution (126 mg L−1 ) (Fig. 4.48). Consequently, 11% of the reduction in Cr(VI) during Cr(VI) removal was due to surface-bound Fe2+ . The mechanism of Cr(VI) elimination was suggested in light of the findings and discussion above. The primary steps in the removing of Cr(VI) by FeS2 /ZVI were coprecipitation, reduction, and adsorption. At the initial stage, the added sulfur significantly reduced ZVI’s surface oxidation during BM and encouraged the development of an indeterminate FeSx stratum. Then, the FeSx stratum of FeS2 /ZVI exhibits good Fig. 4.46 Cr(VI) removal by mixed FeS2 -ZVI and FeS2 /ZVI in the absence or presence of 1,10-phenanthroline. (Particle dosage: 1 g L−1 , [Cr(VI)]0 = 50.0 mg L−1 , NaAc-HAc buffer solution, pH0 = 3.93, air atmosphere). Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022

Fig. 4.47 Dissolution of Fe from mixed FeS2 -ZVI and FeS2 /ZVI in a pure water system. (Particle dosage: 1 g L−1 , air atmosphere). Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022

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225

Fig. 4.48 Effect of different amounts of Fe2+ added on Cr(VI) removal. ([Cr(VI)]0 = 50.0 mg L−1 , pH0 = 5.20, air atmosphere). Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022

rapport to allow the adsorption of Cr(VI) ions onto surface of the material. There were several redox methods used to decrease the concentrated Cr(VI): (1) The reduced sulfur species in pyrite immediately helps to the reduction via Eq. 4.36; (2) ZVI directly transferred multiple electrons coming out of its conduction band (Eq. 4.2) and slowly eroded to discharge Fe(II) (Eq. 4.33); and (3) ZVI and pyrite progressively corrode to give Fe2+ for Cr(VI) reduction (Eqs. 4.34 and 4.35). In this instance, the added sulfur encouraged the growth of extremely active iron on the surface of FeS2 /ZVI. Consequently, 39% of Cr(VI) was removed via direct reduction using ZVI and pyrite (VI). For the decrease of Cr(VI), the released and surface-bound Fe2+ were responsible for 50% and 11%, respectively. Cr(III) hydroxides, oxides, and Crx Fe1-x OOH were the final products of the Cr3+ ions’ precipitation (Eq. 4.37).

4.3.4 Sulfur Speciation Affects the Electron Transfer of FeS2 /ZVI Generally speaking, Sulfur specimens influences the ZVI material activity. Here, FeS2 /ZVI, FeS/ZVI, and S/ZVI were produced using several sulfur precursors, including S2 2− , S2− , and S0 . S2 2− , S2− , and S0 sulfur precursors’ effects on ZVI materials for Cr(VI) Sealed were assessed. The characteristic crystal structure of pyrite and ZVI was preserved in FeS2 /ZVI (Fig. 4.49). The large diffraction peaks in FeS/ ZVI are a sign of a disorganized structure. The appearance of the usual Fe2 O3 (2 = 35.6°, 62.5°) diffraction peak suggests significant passivation during the BM process.

226

4 Mechanism of Chemical Reduction of Cr(VI)

With the incorporation of S to the body-centered cubic lattice of Fe, the diffraction peak of the Fe(110) crystal plane in FeS2/ ZVI, FeS/ZVI, and S/ZVI redshifted noticeably. The highest lattice spacing was found in FeS2 /ZVI, which would have low drag and make electronic transmission for Cr(VI) quarantine easier (Fig. 4.50). Electrochemical techniques were used to study the electrochemical characteristics in order to further verify the role that electronic migration plays in the FeS2 /ZVI system. The corrosion potential varies with sulfur speciation, as seen in Fig. 4.51a. In FeS2 /ZVI, the lowest corrosive potential (0.27 V) was attained. In line with this,

Fig. 4.49 XRD patterns of FeS2 /ZVI, FeS/ZVI and S/ZVI. Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022 Fig. 4.50 Lattice constants of the Fe BCC structure according to XRD Rietveld refinement. Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022

4.3 Interaction Between Pyrite and Zerovalent Iron that Has a Higher Ability …

227

the combined FeS2 -ZVI had a very low free corrosion potential of 0.07 V. This 340 mV difference in corrosion potential indicated that FeS2 /ZVI exhibited drastically different corrosive behavior, which would be helpful for accelerating the electron flow for Cr(VI) Sealed. EIS was used to quantify the iron-based materials’ electron-transport resistance (Ret) (Fig. 4.51b). The introduction of FeS2 to ZVI resulted in the smallest Ret, demonstrating that the FeS2 /ZVI as synthesized had greater conductivity to improve electron transmission. The highest corrosive current (0.6511 A) was realized in this situation (Fig. 4.51c). Moreover, compared to mixed FeS2 -ZVI, FeS/ZVI, and S/ZVI coatings, the negative OCP of the FeS2 /ZVI coating was significantly lower (Fig. 4.51d). Because FeS2 /ZVI often has a higher OCP value and a higher corrosion potential for transferring electrons, it may release electrons more quickly than other iron-based materials. Evaluation was done on the enhanced electron transport’s contribution to the reduction of Cr(VI). There were 35.2°, 51.7°, and 105.7° water contact angles for FeS2 /ZVI, FeS/ZVI, and S/ZVI, respectively (Fig. 4.52). These findings revealed that the ability of FeS2 /ZVI to release Fe2+ from a solution was enhanced by its superior hydrophilicity and increased adhesion force. The Fe2+ concentration was monitored

Fig. 4.51 Electrochemical characterization of FeS2 /ZVI, FeS/ZVI, mixed FeS2 -ZVI and S/ZVI; a Tafel corrosion scans; b electrochemical impedance spectroscopic analyses; c the calculated corrosion current; d open circuit chronopotentiograms. Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022

228

4 Mechanism of Chemical Reduction of Cr(VI)

as these ferrous-based products were submerged in a system of pure water (Fig. 4.53). Fe2+ was released in greater quantities from FeS2 /ZVI than from ZVI or FeS/ZVI, respectively, by a factor of 3.55 and 1.75, which may have a remarkable impact on the decrease of Cr(VI) and subsequent conversion to Fe3+ . In this investigation, the produced Fe3+ might interact with either activated ZVI or FeS2 to produce Fe2+ once more. The Fe3+ Liquid was injected into purified water system to further confirm the reactivity of the Fe2+ regeneration. 68.8 mg L−1 of Fe2+ were produced by FeS2 / ZVI with the addition of Fe3+ . Just 24.48 mg L−1 and 19.93 mg L−1 , respectively, of the scenarios were found in ZVI and FeS/ZVI. Hence, compared to FeS/ZVI and S/ZVI, respectively, FeS2 /ZVI had a Cr(VI) removal efficiency that was 1.52 and 12.39 times greater (Fig. 4.54). A proposed mechanism to explain the enhancement of sulfur incorporation on Cr(VI) elimination was mentioned above. The material directly transferred multiple electrons to decrease Cr(VI) in the case of Cr(VI) reduction by ZVI, however this reduction was inefficient since the material slowly corroded to release Fe2+ for additional Cr(VI) removing. The ability of iron sulfides to transport electrons is superior to that of iron oxides [152], so adding sulfur significantly increased both the transfer of electrons from ZVI to Cr(VI) and the corrosion process of ZVI. Meanwhile, a

Fig. 4.52 Water contact angle images of FeS2/ZVI, FeS/ZVI, and S/ZVI. Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022 Fig. 4.53 Contribution of different S precursors to Fe2+ release and recycling regeneration. (Fe3+ : 20 mg L−1 , particle dosage: 1 g L−1 , air atmosphere). Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022

4.3 Interaction Between Pyrite and Zerovalent Iron that Has a Higher Ability …

229

Fig. 4.54 The influence of S precursors on hydrophobicity and Cr(VI) removal. (Particle dosage: 1 g L−1 , [Cr(VI)]0 = 50.0 mg L−1 , pH0 = 5.20, air atmosphere). Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022

logical theory may be put out that a galvanic couple could be established between FeS2 (+ 0.35 V) and Fe0 (− 0.44 V) because of the substantial potential difference between the two [153]. The anode in each galvanic cell was made of FeS2 . The quick electron passage across the anode and cathode significantly speeds up the release of Fe2+ . Integratedly, the self-inhibition of electron transmission was overcome by numerous tiny galvanic couples, accelerating the reduction of Cr(VI). Later, Fe0 and FeS2 reacted with Fe3+ to create Fe2+ and H+ , which encouraged reduction of Cr(VI) as well as reased the material’s scope of suitable pHs (Fig. 4.55).

Fig. 4.55 Scheme of Enhanced Cr(VI) removal mechanism of FeS2 /ZVI. Reprinted from Li et al. [141] by permission of The Royal Society of Chemistry 2022

230

4 Mechanism of Chemical Reduction of Cr(VI)

4.4 Synergistic Cr(VI) Reduction and Organic Pollutant Oxidative Degradation With the expansion of tanning, galvanizing, dyeing and other sectors, highly hazardous Cr(VI) accompanied by organically-based contaminants (such as phenol) has been commonly detected in actual environments [154–156]. With their manufactured toxicity, such compounded contamination produces more substantial harm to the environment and humans [157]. The majority of researchers focus on organic pollutant degradation and Cr(VI) reduction independently. Unfortunately, it is difficult to find an efficient way to handling Cr(VI) and organically-based contaminants at the same time. The conversion of Cr(VI) to Cr(III) is a critical step in the remediation of Cr(VI) pollution [158, 159]. Advanced oxidation processes (AOPs) have been shown to be successful at removal enol [160]. As a result, the technique of reducing Cr(VI) to Cr(III) and oxidatively degrading organics is extremely beneficial to the remediation of combined pollution [161]. A few research have examined combined Cr(VI) reduction and oxidation of organic pollutants by Fe-based materials using the Fenton response system as well as reactive persulfates method.. However, the technique combining Cr(VI) reduction and phenol oxidation has encountered various difficulties. (1) The oxidation degradation of phenol (such as the production of ·OH radicals) is likely to result in the oxidization of Cr(III) into Cr(VI); (2) With an excess of Cr(VI), phenol oxidation will be reduced to satisfy the demand of Cr(VI) reduction. (3) The removal of Cr(VI) and organics is greatly reliant on solution pH, with alkaline pH resulting in a lower removal rate [162]. The causes could be that the systems’ oxdation reduction buffer ability are insufficient, the oxidation selectivity is low, and the materials (such as Fe0 and Fe3 O4 ) were blunted by product sedimentation. Is it possible to design a constant and effective system for synchronous Cr(VI) reduction and oxidative degraders of phenol with high redox buffering capacity and oxidative selectivity. As it oxidizes to Fe3+ and SO4 2− , pyrite (FeS2 ) provides 15 electrons, allowing a reductive environment to be maintained and the pH increase in the reaction media to be suppressed [127]. FeS2 can be thought of as a oxiation reduction buffer that prevents Cr(III) oxidation [122]. Surprisingly, FeS2 can operate as an acivating agent of Fe0 sulphidation, increasing its hyrhobic and reactive properties [115], In the FeS2 / Fe0 system, more ferrous ions are produced by reacting FeS2 with Fe3+ to drive the Fe2+ /Fe3+ cycles [127]. Similarly, the sulphidation Fe0 with a pyrite-like construction is more hydrophobic, making it easier to delete hydrophobic impurities, and the increased Fe2+ /Fe3+ cycles boost PS activation efficiency throughout The Process of Oxidation. PS-activated oxidation is reported with selectivity and reactivity with organic pollutants in a complex environment [163, 164]. As a result, it is plausible to speculate that FeS2 /Fe0 in conjunction with PS could offer a robust and efficient system for simultaneous reduction of hexavalent chromium and oxidative degradation of phenol.

4.4 Synergistic Cr(VI) Reduction and Organic Pollutant Oxidative Degradation

231

4.4.1 Simultaneous Cr(VI) Reduction and Phenol Oxidation In the FeS2 /Fe0 + PS + Cr(VI) + phenol system, the maximum Cr(VI) removal rate was 100% in 90 min (Fig. 4.56). When compared to the phenol-containing systems, the rates of Cr(VI) removal in the phenol-free systems dropped dramatically (77.85% in the FeS2 /Fe0 + PS + Cr(VI) system and 51.79% in the FeS2 /Fe0 + Cr(VI) system). This finding suggested that degradation of phenol may have aided Cr(VI) reduction. Likewise, Joining Cr(VI) accelerated complete phenol elimination, and the the response time time was reduced from 30 to 10 min (Fig. 4.56b). As a result, these findings indicated a significant synergistic effects between the reduction of Cr(VI) and the oxidation of phenolin the FeS2 /Fe0 + PS system. It is generally understood that Cr(VI) reduction and heterogeneity AOPs are extremely depends on pH. with rapid improvement in removal efficiency declining as (a)

(b)

(c)

(d)

Fig. 4.56 Removal efficiencies of Cr(VI) (a) and phenol (b) in the various systems and effects of initial pH on Cr(VI) reduction (c) and phenol degradation (d) in the FeS2 /Fe0 + PS system (20 mg/ L Cr(VI), 10 mg/L phenol, 0.5 g/L FeS2 /Fe0 and 2 mM PS). Reprinted from Yang et al. [165] Copyright 2021, with permission from Elsevier

232

4 Mechanism of Chemical Reduction of Cr(VI)

liquid pH rises. The effect of starting pH on the rates of Cr(VI) and phenol elimination in the FeS2 /Fe0 + PS system was investigated (Fig. 4.56c and d). Using an initial pH range of 3.0 to 9.0, 100% of Cr(VI) and phenol removing rates were achieved in 120 min and 20 min, respectively. According to the XRD data, FeS2 /Fe0 was stable both pre- and post-reaction (Fig. 4.57). According to zeta potential tests, the catalyst possessed a positive charge in the pH range of 3 to 9, which was advantageous for absorbing negatively charged chromate and decreasing Cr(VI) (Fig. 4.58). When compared to typical Fenton processes, the FeS2 /Fe0 + PS system had a broader usable pH range and higher removal efficiency. Surveillance the pH alteration during the procedure revealed that the pH significantly fell after 3 min (Figs. 4.59 and 4.60). This might be explained by the oxidation of FeS2 to Fe3+ and the breakdown of persulfate, which resulted in a drop in solution pH values PS system [127]. Fig. 4.57 XRD of FeS2 /Fe0 before and after the reaction. Reprinted from Yang et al. [165] Copyright 2021, with permission from Elsevier

Fig. 4.58 Determination of zeta potential in a water system. Reprinted from Yang et al. [165] Copyright 2021, with permission from Elsevier

4.4 Synergistic Cr(VI) Reduction and Organic Pollutant Oxidative Degradation

233

Fig. 4.59 Final pH values after reaction in the solutions with different initial pH conditions. Reprinted from Yang et al. [165] Copyright 2021, with permission from Elsevier

Fig. 4.60 Changes in pH during the reaction of different systems. Reprinted from Yang et al. [165] Copyright 2021, with permission from Elsevier

4.4.2 The Evolution of Cr(VI) in the FeS2 /Fe0 + PS System In the response system including both Cr(VI) and phenol, Fe0 in FeS2 /Fe0 not only serves as a catalyst for PS revitalization o oxidize phenol, but it also plays a significant role in Cr(VI) reduction [164]. Recent research has demonstrated that raising the Cr(VI) concentration can diminish the clearance rates of Cr(VI) and organic contaminants by Fe0 -based Fenton or Fenton-like systems by producing more Crx Fe1−x (OH)3 and/or Crx Fe1-x OOH on the surface of Fe0 [173, 174]. As demonstrated in Fig. 4.61a, the clearance rates of Cr(VI) increased with increasing Cr(VI) to phenol ratios (4:1, 6:1 and 8:1). XPS also revealed that Cr(VI) was primarily reduced to Cr(III), with intense Cr(III) information appearing at 587.26 eV and 577.83 eV after the reactions (Fig. 4.62). Nevertheless, with varied Cr(VI) to phenol ratios, FeS2 /Fe0 still had 100% phenol oxidation degradation efficiency in 10 min. The results showed

234

4 Mechanism of Chemical Reduction of Cr(VI)

that there was no significant passivation impact in the FeS2 /Fe0 + PS system with Cr(VI) as well as phenol. On the surface of FeS2 /Fe0 , Cr(VI) was converted to soluble Cr(III) instead of water-insoluble form such as Crx Fe1-x (OH)3 . This was matched with the XPS data, which showed that the Cr(III) sediment contents on FeS2 /Fe0 reduced from 3.47 at.% to 0 in the PS and phenol systems (Fig. 4.63). Furthermore, the FTIR spectra (Fig. 4.61c) revealed the formation of a Fe–O–Cr band (618 cm1) in the FeS2 /Fe0 + Cr(VI) and FeS2 /Fe0 + Cr(VI) + phenol systems but not in the FeS2 /Fe0 + PS + phenol + Cr(VI) systems [166]. Furthermore, with Cr(VI) decrease, the peak of adsorption related to phenol decomposition products changed from 289 to 294 nm (the distinctive peak at 343 nm) (Fig. 4.61d). The generated Cr(III) is likely to form Cr(III)-organic complexes with the products of phenol degradation in the FeS2 /Fe0 + PS system [167]. (a)

(b)

(c)

(d)

Fig. 4.61 Removal efficiencies of Cr(VI) at different concentration ratios of Cr(VI) to phenol in the FeS2 /Fe0 + PS system (a); removal efficiencies of Cr(VI) and total Cr in the FeS2 /Fe0 + PS system with phenol (b); FTIR spectra of FeS2 /Fe0 in the different systems (c); UV–vis spectra of the FeS2 /Fe0 + PS system with phenol and Cr(VI) (d) (20 mg/L Cr(VI), 10 mg/L phenol, 0.5 g/L FeS2 /Fe0 and 2 mM PS). Reprinted from Yang et al. [165] Copyright 2021, with permission from Elsevier

4.4 Synergistic Cr(VI) Reduction and Organic Pollutant Oxidative Degradation

235

Fig. 4.62 The Cr species on the catalyst after reaction. Reprinted from Yang et al. [165] Copyright 2021, with permission from Elsevier

Fig. 4.63 XPS survey scan of the initial and reacted FeS2 /Fe0 in the different systems. Reprinted from Yang et al. [165] Copyright 2021, with permission from Elsevier

The mid-level Cr(V) could be formed during the reduction of Cr(VI) to Cr(III) [158], ESR also conducted a qualitative investigation into it (Fig. 4.64). In the FeS2 / Fe0 + PS + phenol + Cr(VI) and FeS2 /Fe0 + phenol + Cr(VI) systems, two ESR signals (g-value = 1.9661(1) and 1.9719 (2)) for Cr(V) were found. However, in the FeS2 /Fe0 + phenol + Cr(VI) system, the signal strength rose with increasing reaction time, but it fell to zero after 60 min in the FeS2 /Fe0 + PS system with Cr(VI) and phenol. When comparing the removal efficiency of Cr(VI) and phenol in the two systems, it was safe to conclude that the middle Cr(V) is transformed to Cr(III).more quickly in the FeS2 /Fe0 + PS + phenol + Cr(VI) system, and that degradation of

236 (a)

4 Mechanism of Chemical Reduction of Cr(VI) (b)

Fig. 4.64 ESR signals of Cr(V) in the FeS2 /Fe0 + Cr(VI) + phenol system (a) and FeS2 /Fe0 + Cr(VI) + phenol + PS system (b). Reprinted from Yang et al. [165] Copyright 2021, with permission from Elsevier

phenol expedited this transition [17, 168]. As a result, the degradation of phenol is favorable for the reduction of Cr(VI).

4.4.3 The Roles of Surface-Bound Fe2+ In the original FeS2 /Fe0 sample, Clear observation of the crystal structure of pyrite iron disulfide (Fig. 4.65A1 and B1). The fast Fourier transform (FFT) types (Fig. 4.65A2 and B2) and reversed FFT picture successfully identified the crystallographic planes of FeS2 (Fig. 4.65A3 and B3). The (110) and (210) planes of FeS2 can be indexed by lattice fringes with d-spacings of 0.31 and 0.24 nm, respectively [169]. Several amorphous phases were clearly detected on the surface of the FeS2 / Fe0 catalog, indicating that crystalline FeS2 was converted to amorphous phases throughout the reaction (Fig. 4.65C). As a result, the Fe2+ ions were easily dissolved from the amorphous phases and may contribute to Cr(VI) reduction and PS activation. 1,10-phenanthroline, which can operate as a Fe2+ removal by chelating both water-based and surface combined Fe2+ ions, was used to test the involvement of Fe2+ ions in oxidation reduction conversion [117], was incorporated into the FeS2 /Fe0 + PS system. It is worth noting that 1,10-phenanthroline drastically reduced Cr(VI) efficiency of removal to 16%, confirming the role of Fe2+ ions in Cr(VI) reduction (Fig. 4.66a). At the same time, the phenol remove efficiency gradually declined with decreasing Fe2+ ions (Fig. 4.66b). Our findings revealed that Fe2+ ions were the primary active species responsible for the removal of Cr(VI) and phenol. The Fe2+ solutions was analyzed in comparison with the FeS2 /Fe0 + PS system to identify which Fe2+ species, in the form of aqueous or surface bound, play a role in these

4.4 Synergistic Cr(VI) Reduction and Organic Pollutant Oxidative Degradation

237

processes at concentrations (40 and 200 mg/L) were greater than the Fe2+ contents in the FeS2 /Fe0 + PS system (38.75 mg/L) (Fig. 4.67). And yet, the reductionrate of Cr(VI) by purifiedaqueous Fe2+ was 42% (Fig. 4.66c), and the 100% removal rate of the FeS2 /Fe0 + PS system is much higher than this. In the meantime, purified aqueous Fe2+ had comparatively modest degradation of phenol activity (Fig. 4.66d). This phenomenon is the possibility that the free radicals formed(such as SO4 •− ) in the system have reacted with the excess of Fe2+ in solution. [170]. As a result, surface-bound Fe2+ was likely the most important reactionary species in the removal of Cr(VI) and phenol. The dissolved form was found to have a lower Fe2+ ratio activity than the iron sulfide bound. Furthermore, Solid minerals such as Fe0 and pyrite absorb a portion of the recurring Fe2+ . It has been reported that combining Fe2+ and Fe0 could speed up redox reactions. One possible explanation was that Fe2+ could accelerate the corrosion of Fe0 , which was favourable to rapid electronic delivery [121, 171]. However, when compared to systems without Fe0 , aqueous Fe2+ with Fe0 did not noticeably enhanced the remove efficiency of Cr(VI) and phenol (Fig. 4.68). This result revealed that the Fe2+ /Fe0 loop could not be liable for the FeS2 /Fe0 + PS system’s greater ability to delete Cr(VI) and phenol concurrently. XPS reveals the oxidized state of Fe on the surface of FeS2 /Fe0 (Fig. 4.69a). The peaks at 707.43 and 711.27 eV Corresponding to Fe 2p3/2 of Fe2+ and Fe3+ , respectively, and are pyrite peaks. The peak at 720.0 eV was thought to be due to

Fig. 4.65 HRTEM images of the initial (A1 and B1) and reacted (C1) FeS2 /Fe0 samples; the corresponding Fast Fourier transform (FFT) patterns (A2, B2 and C2) and inverse FFT images (A3, B3 and C3) of the regions (a, b and c); the intensity profiles (A4 and B4) of the d-spacing of the regions (a and b). Reprinted from Yang et al. [165] Copyright 2021, with permission from Elsevier

238

4 Mechanism of Chemical Reduction of Cr(VI) (a)

(b)

(c)

(d)

Fig. 4.66 Influences of 1,10-phenanthroline on Cr(VI) reduction (a) and phenol degradation (b) in the FeS2 /Fe0 + PS system and effects of aqueous Fe2+ on the removal of Cr(VI) (c) and phenol (d) (20 mg/L Cr(VI), 10 mg/L phenol, 0.5 g/L FeS2 /Fe0 and 2 mM PS). Reprinted from Yang et al. [165] Copyright 2021, with permission from Elsevier

Fe 2p1/2 of Fe0 , while the peaks at 724.6 and 728 eV were typical of Fe3+ [172, 173]. The concentrations of Fe2+ fell dramatically after reacting with Cr(VI), from 29.66 to 4.8% (Table 4.9), indicating that Fe2+ was the main reactive species for the reduction of Cr(VI) by FeS2 /Fe0 alone. After reacting with phenol and PS, the Fe0 concentration still remain constant, while the majority of the Fe2+ was transformed to Fe3+ . The findings suggested that Fe2+ on the surface was straightforward engaged in the oxidation of phenol. Nevertheless, a modest drop in Fe2+ concentration (from 29.66 to 21.29%) was found in the FeS2 /Fe0 + PS system used to delete Cr(VI) and phenol. This could be because the synergistic redox conversion of Cr(VI) and phenol increased surface-bound Fe2+ regeneration.

4.4 Synergistic Cr(VI) Reduction and Organic Pollutant Oxidative Degradation

239

Fig. 4.67 Concentrations of Fe2+ in the FeS2 /Fe0 + PS system. Reprinted from Yang et al. [165] Copyright 2021, with permission from Elsevier

4.4.4 The Roles of SO4 2− in the FeS2 /Fe0 + PS System As a result, the high-resolution S 2p spectra (Fig. 4.69b) revealed that S was present as Fe(II)-S (165.4 eV), S2 2− (162.7 eV), Sn 2− (163.95 eV), and SO4 2− (170.7 eV). Because of their low band clearance (Fe3 S4 = 0.0 eV; FeS = 0.10 eV; FeS2 = 0.95 eV), iron oxide phase(Fe2 O3 = 2.2 eV) shows higher electron transport barriers and better reactivity than iron polysulfide. As a result, the produced FeS2 /Fe0 in the current investigation was transmitted to the redox processes. After lowering Cr(VI), the concentrations of S2 2− , Sn 2− , and S0 reduced, indicating that Cr(VI) reduction by FeS2 /Fe0 may be the separate involvement of polysulfides. The concentration of S2 2− reduced from 37.17 to 31.75% after reaction with phenol and PS (Table 4.9). S2 2− may beoxidized during PS activation. Surprisingly, removal of Cr(VI) and phenol in FeS2 /Fe0 + PS system, an obvious drop in SO4 2− level was detected, showing that SO4 2− could contribute in the cooperative Cr(VI) and phenol oxidatio reduction. The SO4 2− balance concentration sustained at around 102 mg/L in the FeS2 /Fe0 + Cr(VI) system as well as rose to 440 mg/L in the FeS2 /Fe0 + Cr(VI) + PS and FeS2 / Fe0 + PS + phenol + Cr(VI) systems (Fig. 4.70). SO4 2− was produced in the various systems through FeS2 oxidation and PS breakdown. Furthermore, the equilibrium time of SO4 2− was reduced by 50% in the FeS2 /Fe0 + PS + phenol + Cr(VI) system compared to the FeS2 /Fe0 + Cr(VI) + PS system, indicating that phenol oxidation increased SO4 2− production. SO4 2− aids in the removal of the passive oxidation layer of ferrous, encouraging Fe0 corrosion. As a result, it can be concluded that the increased formation of SO4 2− during degradation of phenol is one of the main reasons for the promotion of Cr(VI) removing, which was also in agreement with the results of Cr(VI) removal.

240 Fig. 4.68 Removal of Cr(VI) (a) and phenol (b) by the Fe2+ /Fe0 + PS system. Reprinted from Yang et al. [165] Copyright 2021, with permission from Elsevier

4 Mechanism of Chemical Reduction of Cr(VI) (a)

(b)

4.4.5 Reactive Species for Phenol Degradation Quenching studies were carried out to assess the involvement of responsive species in degradation of phenol in the FeS2 /Fe0 + PS system (Fig. 4.71a). Tert-butanol (TBA) is commonly employed as a HO· decavenger [173]. MeOH acted as a decavenger for HO· and SO·− 4 species [174]. p-Benzoquinone (p-BQ) and furfuryl alcohol (FFA) 1 were used as O·− 2 and O2 cavengers, respectively [175, 176]. The phenol elimination efficiency efficiencies were marginally reduced (from 100% to 93.53%) with the adding of 0.2 M TBA, showing that HO· may not be the primary reactive species for phenol removal. Nevertheless, FFA hindered phenol elimination by 85.42%, implying that 1O2 is the major responsible and reactive species for phenol oxidation

4.4 Synergistic Cr(VI) Reduction and Organic Pollutant Oxidative Degradation (a)

241

(b)

Fig. 4.69 XPS spectra of Fe 2p (a) and S 2p (b) of the initial and reacted FeS2 /Fe0 in the different systems. Reprinted from Yang et al. [165] Copyright 2021, with permission from Elsevier Table 4.9 Relative abundances of Fe and S species on the surface of the initial and reacted FeS2 / Fe0 in the different systems Samples

Fe2p

S2p

Fe0 (%) Fe(II) (%) Fe(III) (%) S2 2− (%) Sn 2− (%) S0 (%) SO4 2− (%) FeS2 /Fe0

12.84

29.66

57.49

37.17

21.93

4.46

36.43

FeS2 /Fe0 + Cr(VI)

11.4

4.8

83.82

28.41

17.80

4.17

49.62

FeS2 /Fe0 + 12.79 PS + phenol

8.75

78.45

31.75

22.54

8.89

36.83

FeS2 /Fe0 + 15.16 PS + phenol + Cr(VI)

21.29

63.55

42.55

28.51

5.53

23.40

Reprinted from Yang et al. [165] Copyright 2021, with permission from Elsevier Fig. 4.70 Release of SO4 2− in the different reaction systems. Reprinted from Yang et al. [165] Copyright 2021, with permission from Elsevier

242

4 Mechanism of Chemical Reduction of Cr(VI) (a)

(b)

Fig. 4.71 Phenol degradation rates with different scavengers in the FeS2 /Fe0 + PS system (a); EPR spectra using DMPO and TEMPO as spin-trapping regents for the FeS2 /Fe0 samples at different reaction times (b) (20 mg/L Cr(VI), 10 mg/L phenol, 0.5 g/L FeS2 /Fe0 , 2 mM PS and pH0 = 5). Reprinted from Yang et al. [165] Copyright 2021, with permission from Elsevier

in the FeS2 /Fe0 + PS system. Furthermore, MeOH and p-BQ dramatically reduced phenol degradation efficiency, dropping from 100% to 60.27% and 80.63%, respectively. This meant that SO4 ·− and O2 HO·− were almost certainly involved in phenol elimination. As a result, it was inferred that both militant and nonradical routes were implicated in PS activation by FeS2 /Fe0 , with the former perhaps being the more irrelevant. In order to further determine the formation and revolution of the activating species in the FeS2 /Fe0 + PS system using electron paramagnetic resonance (EPR) spectroscopy (Fig. 4.71b). A robust messages with a 1:1:1 trifecta lettering, typical of DMPO-1 O2 , was found. This identified that 1 O2 was produced and had a significant role in the breakdown of phenol. The messages of the DMPO-HO ·, DMPO-O2 − , and DMPO-· SO4 − additives were still seen, showing that the system generated HO ·, O2 , and·SO4 − . No Cr(III) was reoxidized, which suggests that the nonradical pathway may be stronger selectivity for organic molecules. To test that hypothesis, vary Cr(III) concentrations (20, 50, and 100 mg/L) were introduced to the FeS2 /Fe0 + PS + phenol system (Fig. 4.72). Only a trace amount of Cr(VI) (1.4–2.3 mg/L) was produced in the first 10 min and was quickly removed. In the meantime, Cr(III), whether straightly added or reduced, increased degradation of phenol in the FeS2 / Fe0 + PS system (Fig. 4.73). Therefore, the decreased Cr(III) in the FeS2 /Fe0 + PS system increased phenol degradation.

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243

Fig. 4.72 Production of Cr(VI) with the addition of different concentrations of Cr(III) in the FeS2 /Fe0 + PS system Reprinted from Yang et al. [165] Copyright 2021, with permission from Elsevier

Fig. 4.73 Degradation efficiencies of phenol with the addition of Cr(III) or Cr(VI) in the FeS2 /Fe0 + PS system. Reprinted from Yang et al. [165] Copyright 2021, with permission from Elsevier

4.4.6 Synergistic Redox Conversion Mechanism Figure 4.74 depicts the redox properties of the various systems. Cathode maximum I considered to be FeS2 /Fe0 decrease of surface species (Eqs. (4.38) and (4.39)) [177]. Peak I in the FeS2 /Fe0 + PS + phenol + Cr(VI) system designed to a oxidation reduction capacity than in the FeS2 /Fe0 and FeS2 /Fe0 + Cr(VI) systems, demonstrating that Fe2+ is more probable with PS and phenol on FeS2/Fe0 surfaces. Cathodic peak II at 0.35 V was recognized as the surface decrease of S0 (Eqs. (4.40) and (4.41). However, no clear tendency to change was observed in the three systems,

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4 Mechanism of Chemical Reduction of Cr(VI)

Fig. 4.74 Cyclic voltammograms for FeS2 /Fe0 (T = 25 °C, scan rate = 20 mV/s) in the different systems. Reprinted from Yang et al. [165] Copyright 2021, with permission from Elsevier

indicating that S0 was not the dominant reactionary species. By adding Cr(VI), the Cr(VI)-Cr(III) redox pair was produced, which was disclosed by cathodic peak III at 0.6–0.9 V. Also noteworthy were cathodic peak III deviated to lower the redox potential with additional PS and phenol addition. This data revealed that phenol oxidation was more likely involved in the conversion of Cr(VI) to Cr(III). These findings supported the synergistic effects of Cr(VI) reduction and phenol oxidation. + FeS2 + 11H2 O → Fe(OH)3 + 2SO2− 4 + 19H + 15e

(4.38)

Fe(OH)3 + 3H+ + e → Fe2+ + 3H2 O

(4.39)

FeS2 → Fe2+ + S0 + e

(4.40)

S0 + 2H+ + 2e → H2 S

(4.41)

Figure 4.75 depicts a schematic diagram depiction of the Cr(VI) and phenol co-redox conversion methods by FeS2 /Fe0 and PS. Cr(VI) was reduced to Cr(III) throughout the redox process via the intermediate Cr(V). The reduced Cr(III) product increases the degradation of phenol and may also produce soluble polymers with phenol degradation produits to refrain from FeS2 /Fe0 Superficial passivation. The oxidative degradation of phenol by FeS2 /Fe0 -activated PS, on the other hand, engaged both radical and nonradical routes. The nonradical pathway to produce 1 O2 was

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Fig. 4.75 Schematic diagram of the synergistic redox conversion mechanism of Cr(VI) and phenol removal by FeS2 /Fe0 and PS. Reprinted from Yang et al. [165] Copyright 2021, with permission from Elsevier

the largest donators to phenol degradation, which could explain the poor Cr(III) reoxidation in the system. Furthermore, degradation of phenol combined with rapid PS breakdown accelerated SO4 2− formation, which may encourage Fe0 corrosion to produce the table bound Fe2+ and hence improve reduction of Cr(VI). Furthermore, degradation products of phenol may hasten the transition of Cr(V) to Cr(III) and, therefore, enhance the reduction of Cr(VI) to Cr(III).

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Chapter 5

Chemical Remediation of Chromium-Contaminated Soil

For the purpose of cleaning up soil that has been tainted by chromium, several remediation methods have been invented throughout the course of time. The chemical reduction strategy is one of the common technologies, and it has the advantages of being inexpensive, having wide applicability, and being quite effective in cleaning up contaminated sites. Adding chemical reduction agents is one method for using this strategy, which transforms highly toxic and mobile Cr(VI) into less toxic and less mobile Cr(III). Even while iron-bearing reductants have a beneficial effect on lowering Cr(VI), using them does not come without its share of adverse side effects. These downsides include soil acidification or secondary pollution, costly expenditures, and restricted adaptive pH ranges, which are often below 6 [1, 2]. For instance, ferrous sulfate has been shown to lower the pH of soil, which in turn raises the likelihood of soil acidification [3] which in turn causes an excessive concentration of sulfate to be leached away; In addition, ferrous sulfate is applicable only when the surrounding environment is acidic, and it is much less helpful when the surrounding environment is alkaline. It is simple to aggregate and passivate nano zerovalent iron (nZVI), which lowers its reducing capability and hinders the long-term stability of soil that has been remedied. For these reason, the decontamination of soil that is polluted with Cr(VI) requires the development of an effective chemical reduction strategy that is based on iron-bearing reductants.

5.1 Kinetics, Thermodynamics And Long-Term Effects in Remediation of Cr(VI)-Contaminated Soil Due to its higher Fe(II) regeneration and more comprehensive usable pH range for Cr(VI) removal, zerovalent iron-loaded pyrite (FeS2 /ZVI) shows promise as an effective reductive material for the cleanup of Cr(VI)-polluted soil. Iron with zero valences has been synthesized as FeS2 /ZVI. In addition, there has yet to be any systematic © The Author(s), under exclusive license to Springer Nature Singapore Pte Ltd. 2023 W. Yang et al., Remediation of Chromium-Contaminated Soil: Theory and Practice, Environmental Science and Engineering, https://doi.org/10.1007/978-981-99-5463-6_5

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research on how well-regenerated soil holds up in a complicated setting. Validating the feasibility of remediating Cr(VI)-contaminated soil may depend on answering this question.

5.1.1 Factors Affecting Remediation of Cr(VI)-Contaminated Soil Soil samples containing chromium compounds were collected in Gansu Province, China. 966.1, 448.1, and 439.6 mg kg−1 were the values for total Cr (Crtotal), total Cr(VI) (the total quantity of hexavalent chromium in soil), and water-soluble Cr(VI), respectively. The percentage of Cr(VI) that may be dissolved in water is greater than 98% (Table 5.1). The effectiveness of pristine pyrite, ZVI, and ZVI/FeS2 in the removal of Cr(VI) from contaminated soil was analyzed and compared (Fig. 5.1a). The amount of total Cr(VI) and water-soluble Cr(VI) that was present in the soil decreased as the treatment duration increased from 1 to 7 days, but the removal impact did not alter after the seventh day had passed. It is abundantly clear that the impact of ZVI/FeS2 on the restoration process was superior than that of pure pyrite and ZVI. In only seven days, ZVI/FeS2 was able to achieve a removal effectiveness of more than 99.5% for both total Cr(VI) and water-soluble Cr(VI). The fact that the surface of ZVI was enveloped by numerous oxides and passivated in alkaline circumstances contributed to the fact that all removal efficiencies by pure ZVI were less than 40%. Pyrite had an efficiency of 87.8% in eliminating total Cr(VI); however, there was still up to 54.62 mg kg−1 of residual Cr(VI) in the soil, which is still highly harmful and needs further treatment. In particular, with the more demanding remediation requirement, 5.7 mg kg−1 has been recommended as the Cr(VI) risk screening value for soil contamination of development land in line with GB36600-2018. This value is used to determine whether or not soil pollution poses a risk to human health. This value is used to assess the potential danger posed by the presence of the contaminant. It was tested how the dose of ZVI/FeS2 affected the remediation performance of soil that was polluted with Cr(VI). Once the dose was increased, there was a subsequent drop in the total Cr(VI) content in the soil. Because of the application of a dose of 5%, the total Cr(VI) in the soil was reduced by 99.5%, and now there Table 5.1 Physical and chemical properties of soil samples Property

pH

Cation exchange capability (CEC)

Total organic carbon (TOC)

Electrical conductivity (EC)

Water soluble Cr(VI)

Total Cr(VI)

Total Cr

Value

9.12

10.8 cmolc kg−1

0.88%

240 mS m−1

439.6 mg kg−1

448.1 mg kg−1

966.1 mg kg−1

Reprinted from Min et al. [4] Copyright 2021, with permission from Elsevier

5.1 Kinetics, Thermodynamics And Long-Term Effects in Remediation …

259

Fig. 5.1 Immobilization effect of FeS2 , ZVI and ZVI/FeS2 on Cr(VI)-contaminated soil (a), Effect of ZVI/FeS2 dosage (b), soil–water ratio (c) and pH of soil (d) on the removal of Cr(VI). Reprinted from Min et al. [4] Copyright 2021, with permission from Elsevier

is only 2.33 mg kg−1 of total Cr(VI) left. During the subsequent studies, a dose of 5% ZVI/FeS2 was decided upon, taking into consideration both its efficiency and its cost. The influence of the soil–water ratio on the effectiveness of the cleanup was another factor that was looked at. The ratio of soil to water had minimal effect on the remediation effect of ZVI/FeS2 , and regardless of the ranges of soil to water that were investigated, all of the removal efficiencies of total Cr(VI) in soil were greater than 98%. This was the case regardless of the soil-to-water ratio being evaluated. Because of this, the process of restoring ZVI/FeS2 in soil polluted with Cr(VI) may have a high degree of adaptation to the conditions of the soil. The effectiveness of ZVI/FeS2 in rehabilitating contaminated sites was evaluated using a range of soil pH values from 5.2 to 10.1. It was previously thought that alkaline circumstances would not be suitable for reducing Cr(VI) because these conditions would produce hydroxide precipitates on the surface of iron-based materials. It is important to note that the removal rates of total Cr(VI) by ZVI/FeS2 in this study were over 99% throughout the pH ranges of 5.2 to 10.1, and the concentrations of total Cr(VI) in the soil following ZVI/FeS2 remediation were below 4 mg kg−1 . The results of this study were published in the journal Environmental Science and Technology. The results of this study were published in the journal Environmental Science and Technology.

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5 Chemical Remediation of Chromium-Contaminated Soil

This demonstrated that ZVI/FeS2 exhibited vigorous reduction activity even under alkaline conditions. This may be linked to the formation of hydrogen ions that occurs during the oxidation of FeS2 , which in turn can be attributed to the reduction activity of ZVI/FeS2 . Because the soil at many Cr(VI)-contaminated sites (e.g., chromium chemical sites) is alkaline, the soil pH can be as high as 11 [5], and ZVI/FeS2 with self-adjusting alkali ability has a great deal of potential in the restoration of Cr(VI)contaminated soil. [Cr(VI)]-contaminated soil can be remedied using ZVI/FeS2 with self-adjusting alkali ability. ZVI/FeS2 with self-adjusting alkali ability has a lot of promise in the restoration of Cr(VI)-contaminated soil. This is because the soil at many Cr (VI)-contaminated sites (e.g., chromium chemical sites) is alkaline, and the soil pH can be as high as 11 [5]. ZVI/FeS2 with the ability to self-adjust its alkali level can be used to clean up soil that is polluted with [Cr(VI)].

5.1.2 Kinetics and Thermodynamics The ZVI/FeS2 dosage was reduced to 1%, and a medium consisting of a dilute acetic acid solution was used for this experiment. The purpose of this study was to evaluate the effect that temperature has on the kinetics of removing Cr (VI) from soil. This helped to minimize the time needed for the reaction, which made it possible to analyze the reaction kinetics in a reasonably short length of time (two hours). The connection between the multiphase surface reaction and the reaction that took place between nZVI/ZVI and Cr (VI) may be seen in the fact that the ZVI/FeS2 and Cr (VI) reactions were quite similar. Therefore, in order to fit the reaction kinetics data, the first-order Langmuir-Hinshelwood model (Eq. 5.1), which is sufficient for understanding the majority of surface reaction processes, was utilized [6]. V =−

kbC dC = dt 1 + bC

(5.1)

where k stands for the solid surface reactivity factor, and b stands for the sorption factor in the equation. The equation can be reduced as given below when taking into consideration the extremely low concentration of the reaction substrate: V =−

dC = kbC = kobs C dt

(5.2)

The expression k obs = kb can be found in the equation, and the following reaction may be incorporated into the equation:  ln

Ct C0

 = −kobs t + c

(5.3)

5.1 Kinetics, Thermodynamics And Long-Term Effects in Remediation …

261

Fig. 5.2 Effect of temperature on the removal of Cr(VI) (a) and linear fit relationship of ln(C/C 0 ) and time (b). Reprinted from Min et al. [4] Copyright 2021, with permission from Elsevier

where C 0 and C t represent the initial content of Cr(VI) and the content of Cr(VI) at t min in the soil leachate, respectively, c is a constant, and K obs may be calculated from the slope of the line that connects ln(C t /C 0 ) and t. In this equation, C 0 represents the initial concentration of Cr(VI), and C t represents the Cr(VI) concentration at t min. It was demonstrated that an increase in temperature led to an improvement in the removal efficiency of Cr(VI) (Fig. 5.2). At temperatures of 293 K, 298 K, 303 K, and 308 K, respectively, the reaction rate constants Kobs were found to be 21.9, 22.8, 25.3 and 31.6 min−1 . The reduction in Cr(VI) concentration was facilitated by an increase in temperature. Because of the fact that, as the temperature rises, there is an increase in the activation degree of reactant molecules, this finding may be attributed to the fact that higher temperatures [7]. E a can be obtained from the Arrhenius equation: ln K = −

Ea + ln A0 RT

(5.4)

where E a is the apparent activation energy measured in kilojoules per mole, R is the gas constant measured in 8.314 kJ mol−1 , A0 is the Gibbs free energy, and T is the temperature measured in degrees Kelvin. Ea and lnA0 were calculated to be 17.97 and 3.51 kJ mol−1 , respectively, after the apparent rate constants were fitted to the temperature data in Fig. 5.2. In most cases, a chemical reaction’s activation energy, or Ea, ranged somewhere between 60 and 250 kJ mol−1 . The fact that the value of Ea was so low in our investigation suggested that the reaction reduced Cr(VI) went exceptionally smoothly. We used the following thermodynamic equations to get the thermodynamic parameters, such as the Gibbs free energy ΔG0 , the standard enthalpy ΔH 0 , and the standard entropy ΔS 0 : G 0 = −RT ln K D ln K D = −

H 0 S 0 + RT R

(5.5)

(5.6)

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5 Chemical Remediation of Chromium-Contaminated Soil

The adsorption equilibrium constant is denoted by K D . We may calculate the ΔH 0 and ΔS 0 variation using the 1/T and lnKD relationship. The relevant thermodynamic information is included in Table 5.2. Cr(VI) spontaneously reacted with ZVI/FeS2 since all ΔG0 values were less than 0. The more negative ΔG0 increased with increasing temperature, the more favourable the conditions were for the reaction. This matched the outcomes of fitting the LangmuirHinshelwood kinetic model. Since the enthalpy change during Cr(VI) reduction was 471.41 J mol−1 , this reaction was endothermic. Cr(VI) reduction via ZVI/FeS2 was also shown to increase disorder and degree of freedom at the interface, as measured by a positive value of the standard entropy change (ΔS 0 ) in the system.

5.1.3 Soil pH Variation We measured the soil’s pH before and after ZVI and FeS2 treatment (Fig. 5.3). During the ZVI/FeS2 remediation process, the original soil pH ranged from 5.2 to 10.1 but has since trended towards neutrality. This demonstrates that ZVI/FeS2 has a high degree of self-regulation, which is crucial for rehabilitating soil function. Cr(VI) predominantly persisted as HCrO4 − under acidic conditions (pH 5.2), as predicted, whereas ZVI/FeS2 particles dissociated to create Fe2+ and S2 2− , both of which can lower Cr (VI). Soil pH was raised because H+ (Eq. 5.7–5.10) was lost during the corrosion of ZVI and the oxidation–reduction pathway involving Fe(II), S2 2− , and HCrO4 − . The generation of hydrogen ions during ZVI/FeS2 remediation in alkaline soil with a pH between 9.1 and 10.1 can be attributed to the oxidation of pyrite (Eq. 5.11), the circulation of Fe(III) and the reincarnation of Fe(II) (Eq. 5.12), and the coprecipitation of Cr(III) and Fe(III) (Eq. 5.13): 3Fe0 + 2HCrO4 − + 14H+ → 2Cr3+ + 3Fe2+ + 8H2 O

(5.7)

FeS2 → Fe2+ + S2 2−

(5.8)

3Fe2+ + HCrO4 − + 7H+ → Cr3+ + 3Fe3+ + 4H2 O

(5.9)

3S2 2− + 14HCrO4 − + 50H+ → 14Cr3+ + 6SO4 2− + 32H2 O

(5.10)

2FeS2 + 7O2 + 2H2 O → 2Fe2+ + 4SO4 2− + 4H+

(5.11)

FeS2 + 14Fe3+ + 8H2 O → 15Fe2+ + 2SO4 2− + 16H+

(5.12)

(1 − x)Fe3+ + xCr3+ + 3H2 O → Crx Fe(1−x) (OH)3 + 3H+

(5.13)

22.8

25.3

31.6

303

308

21.94

27.40

30.40

31.65

The half-life periods (min)

0.9944

0.9845

0.9870

0.9886

R2

17.97 (Generally: 60–250 kJ·mol−1 )

Activation energy (kJ·mol−1 )

Reprinted from Min et al. [4] Copyright 2021, with permission from Elsevier

21.9

298

K obs (mmin−1 )

Kinetic parameters

293

Temperature (K)

Table 5.2 Kinetic and thermodynamic parameters under different temperatures

471.41

− 53.3 − 63.7 − 80.9

0.0316

− 56.5

ΔH 0 (J·mol−1 )

ΔG0 (J·mol−1 )

0.0253

0.0228

0.0219

lnK D (min−1 )

Thermodynamic parameters

1.78

ΔS 0 (J·mol−1 )

5.1 Kinetics, Thermodynamics And Long-Term Effects in Remediation … 263

264

5 Chemical Remediation of Chromium-Contaminated Soil

Fig. 5.3 The pH changes in soil during remediation by ZVI/FeS2 . Reprinted from Min et al. [4] Copyright 2021, with permission from Elsevier

5.1.4 Soil Cr Speciation Change At various intervals during the restoration process, a five-stage sequential extraction was used to undertake a study of the speciation of Cr in the soil (Fig. 5.4). In the control sample, the exchangeable fraction of the soil’s Cr was 50.77%, the carbonatebound fraction was 4.20%, the Fe-Mn oxide-bound fraction was 26.62%, the organic matter-bound fraction was 10.53%, and the residual fraction was 7.88%. The transition of Cr into more stable forms was shown by the considerable improvement in both the OX and OM fractions following the addition of ZVI/FeS2 , which led to a significant improvement in the OX fraction. By three days, the EX fraction had almost completely vanished. In general, as the amount of time spent on remediation increased, the OX percentage also climbed, reaching a proportion of 85.03% after thirty days had passed. During the ZVI/FeS2 reduction remediation, the majority of the Cr was found in the speciation of Cr(III)/Fe(III) coprecipitation (Crx Fe1−x (OH)3 ). This enhanced the binding quantity of Fe-Mn oxide, successfully decreasing the mobility of Cr. The chain spherical precipitation layer was wrapped around the solid surface in the scanning electron micrograph (Fig. 5.5) that depicted the results of the interaction between ZVI/FeS2 and Cr(VI). Treatment of Cr(VI)-contaminated soil with ZVI/FeS2 resulted in the creation of Cr(III)/Fe(III) coprecipitation, as confirmed by the related EDS analysis, which revealed that the precipitate was constituted of Fe and Cr, with a Cr to Fe ratio of around 1:5.65. This proved that the precipitate was mostly made up of Fe and Cr.

5.1 Kinetics, Thermodynamics And Long-Term Effects in Remediation …

265

Fig. 5.4 Speciation distribution of Cr at different times. Reprinted from Min et al. [4] Copyright 2021, with permission from Elsevier

Fig. 5.5 SEM–EDS images for reaction products of Cr(VI) and ZVI/FeS2 . Reprinted from Min et al. [4] Copyright 2021, with permission from Elsevier

5.1.5 Long-Term Stability of Cr in Remediated Soil Most soils polluted with hexavalent chromium found in the literature can accomplish Cr(VI) reduction in the short term after treatment with reductive materials. Nevertheless, only a few researchers have focused on the long-term stability of restored soil. It was discovered in recent years that soil that had been polluted with Cr(VI), even after it had been cleaned up, and particularly soil that had been heavily contaminated with alkaline, was prone to Cr(VI) reoccurrence, also known as the yellowing phenomena [8–10]. In theory, there were two ways that Cr(VI) might come back into existence: I, the release of any remaining Cr(VI) the reoxidation of any reduced Cr (III). It is essential to conduct a longer-term stability test on soil treated with ZVI/ FeS2 to remove Cr(VI) contamination. Several different evaluation strategies, such as TCLP, SPLP, and PBET, were utilized to simulate the soil’s stability following

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5 Chemical Remediation of Chromium-Contaminated Soil

remediation in various settings. This was done so that we could gauge how much of a threat the remedied soil would be to wildlife and ecosystems. Objectives, simulated settings, and other details for each of the three methods of evaluation are included in Table 5.3. At pH 5.2, 7.6, 9.1, and 10.1, the original soil had concentrations of TCLPextractable Cr(VI) that were 10.02, 17.8, 18.2, and 19.89 mg L−1 , respectively (Fig. 5.6). During three days of treatment with ZVI/FeS2 , the leaching content of Cr(VI) dropped quickly to less than 0.1 mg L−1 , which was a value that was much lower than the restriction for hazardous waste in the United States, which is 5 mg L−1 . For the subsequent 180 days, not one of the Cr(VI) leaching concentrations increased beyond the highly safe threshold of 0.1 mg L−1 . After being treated with ZVI and FeS2 , all of the soluble Cr(VI) concentrations in the SPLP test were lower than 0.23 mg L−1 , within the acceptable limit. The soil’s bioaccessibility to Cr was measured using the PBET test. A consistent leachable Cr(VI) content of 0.1 mg L−1 was maintained throughout the PBET test period of 180 days. At pH 5.2, 7.6, 9.1, and 10.1, the amount of Cr(VI) bioaccessible in the untreated soils was, correspondingly, 33.03%, 58.02%, and 67.62%, respectively and 62.71%, respectively. The bioaccessibility of Cr(VI) in all samples was reduced to < 0.23% after 180 days, after only three days of remediation. Results showed that soil treated with ZVI/FeS2 for Cr(VI) contamination displayed remarkable long-term stability. These two materials were used to clean up the dirt. Because Cr(VI) was reduced and Cr was immobilized predominantly by the production of Cr(III)/Fe(III) coprecipitation, the leaching concentrations of Crtotal dropped dramatically following ZVI/ FeS2 treatment. Similar to Crtotal , Cr(VI) leaching concentrations in the treated soil followed a trend. There was no obvious sign of Cr(III) reoxidation after 180 days, and the leachate concentrations of Crtotal in TCLP and SPLP remained at deficient Table 5.3 Detailed information on the three evaluation methods Evaluation Method Toxicity Characteristic Leaching Procedure (TCLP)

Synthetic Physiologically Based precipitation leaching Extraction Test (PBET) procedure (SPLP)

Purpose of evaluation

Whether it is a hazardous waste

Stability of leachate under characteristic acid rain conditions

Bioaccessibility of Cr to humans

Simulation environment

Landfill leachate environment

Characteristic acid rain

Human gastric system and small intestine environment

Main extracts

Acetic acid, water

Sulfuric acid, nitric acid, water

Glycine, water

pH of the extract solution

4.93

4.2

2.3

Solid-to-liquid ratio

1:20

1:20

1:100

Reprinted from Min et al. [4] Copyright 2021, with permission from Elsevier

5.2 Remediation of Cr(VI) and Organic Pollutant Cocontaminated Soil

267

Fig. 5.6 Three different extractive toxicities of Cr(VI) and total Cr at different time points after remediation of contaminated soil by ZVI/FeS2 . Reprinted from Min et al. [4] Copyright 2021, with permission from Elsevier

levels. Hence, the remediated soil was stable over the long term under the simulated conditions of landfill leachate and typical acid rain conditions. Slow increases in PBET leachable Crtotal concentrations after 30 days suggest the reduced products may be reabsorbed into the aqueous phase here. Since the leachable Cr(VI) concentration in PBET remained at a surface level for 180 days, the re-released Cr was in the form of Cr(III), which was still non-hazardous due to its low concentration and low toxicity. As a result, ZVI/FeS2 has impressive stability in soil polluted with Cr(VI) after remediation, which may increase its usefulness in remedial engineering.

5.2 Remediation of Cr(VI) and Organic Pollutant Cocontaminated Soil Soil (like soil at a tanning facility) can be polluted with both Cr(VI) and organic contaminants like phenol., and developing a highly efficient approach to eliminate Cr(VI) and organic pollutants simultaneously has received extensive attention. In the FeS2 /Fe0 + PS system, researchers have recently uncovered a surprising synergistic interaction between the reduction of Cr(VI) and the oxidation of phenol [11]. Our findings shed fresh light on the concurrent removal of Cr(VI) and organic contaminants and give new insight into the process. In addition, creating a dependable FeS2 / Fe0 -based system is anticipated to facilitate the cleanup of soil polluted with both Cr(VI) and organic pollutants simultaneously.

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5 Chemical Remediation of Chromium-Contaminated Soil

5.2.1 Factors Affecting the Remediation of Cr(VI) and Organic Pollutants The tested soil contained 480 mg kg−1 total Cr(VI) and 45.02 mg kg−1 phenol (Table 5.4). The concentrations of water-soluble Cr and water-soluble Cr(VI) were 500 mg kg−1 and 439.6 mg kg−1 , respectively. Factors including FeS2 /Fe0 dosage, PS dosage and water–soil ratio on the remediation of Cr(VI) and phenol co-contaminated soil were evaluated. The removal efficiency of water-soluble Cr(VI) was above 88% within 1 day when 1–5% of FeS2 /Fe0 dosage was added to the soil (Table 5.5). When 5% FeS2 / Fe0 was added, the removal efficiency of water-soluble Cr(VI) reached 98%. With the extension of treatment time from 1 to 3 days, the removal efficiency of watersoluble Cr(VI) was further improved and reached above 99% with dosages of 3–5% materials. The potassium persulfate (PS) dosage has little effect on the removal efficiency of Cr(VI) in soil (Fig. 5.7), which maintains a high removal efficiency of approximately 98% with dosages from 4 to 15 mm. However, the PS dosage has a great impact on the degradation efficiency of phenol in the soil. When the PS dosage increases from 4 to 10 mm, the phenol removal efficiency correspondingly increases from 49.98 to 84.26%. When the PS dosage continues to increase to 15 mm, the phenol removal efficiency decreases to 79.20%. This is possibly caused by an excessive dose of oxidant, which leads to a self-quenching effect of excessive free radicals. The removal effectiveness of soil Cr(VI) was maintained at about 98% even when the water-to-soil ratio was changed from 1:1 to 2:1. (Fig. 5.8).However, the removal efficiency of phenol in the soil is greatly affected by the water-to-soil ratio. Phenol elimination effectiveness was just 38.66% in 1 day when the ratio was 1:2. Due to a higher proportion of water to soil, the removal efficiency of phenol also increased. When the water-to-soil ratio increased to 2:1, the removal efficiency of Table 5.4 Physical and chemical properties of soil samples Property

pH

Water soluble Cr(VI)

Total Cr(VI)

Water soluble Cr

Phenol

Value

7.31

439.6 mg kg−1

480 mg kg−1

500 mg kg−1

45.02 mg kg−1

Table 5.5 Effect of FeS2 /Fe0 dosage on the removal of water-soluble Cr(VI) Treatment time

FeS2 /Fe0 dosage (%)

Removal efficiency of water-soluble Cr(VI) (%)

1d

1

88.7

3

97.66

5

98.67

1

92.14

3

99.12

5

99.22

3d

5.2 Remediation of Cr(VI) and Organic Pollutant Cocontaminated Soil

(a)

Cr(VI)

(b) 100

80 60 40

60 40 20

20 0

phenol

80

Removal rate (%)

Removal rate (%)

100

269

4

6

8

10

12

0

15

4

6

8

10

15

12

dosage of PS (mM)

dosage of PS (mM)

Fig. 5.7 Effect of PS dosage on the removal of Cr(VI) (a) and phenol (b) in soil (a)

Cr(VI)

100

(b) 100

phenol

80

Removal rate(%)

Removal rate(%)

80

60

40

40

20

20

0

60

1:2

1:1 water:soil

2:1

0

1:2

1:1 water:soil

2:1

Fig. 5.8 Effect of the water-to-soil ratio on the removal efficiency of Cr(VI) (a) and phenol (b) in soil

phenol reached 86.32%. The reason for this difference may be that reducing the water-to-soil ratio will lead to uneven water/soil mixing so that the oxidant cannot fully contact the pollutants in the soil.

5.2.2 Remediation Performance Comparison of Different Reaction Systems Different reaction systems have different remediation effects on Cr(VI) and phenol (Fig. 5.9). Whether PS was added or not, Cr(VI) removal efficacy neared 98%, with FeS2 /Fe0 being the primary contributor in soil Cr(VI) reduction.. It is also observed that the degradation efficiency of phenol can reach 100% and the removal efficiency of Cr(VI) can reach 63.81% when PS is added alone. Decreased phenol degradation

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5 Chemical Remediation of Chromium-Contaminated Soil

Fig. 5.9 Effects of different reaction systems on the removal of hexavalent chromium and phenol from soil

Cr(VI)

100

phenol

Removal rate(%)

80

60

40

20

0

FeS2/Fe0+PS

FeS2/Fe0

PS

efficiency with the addition of FeS2 /Fe0 can be explained by the fact that FeS2 / Fe0 may induce the quenching of certain free radicals; moreover, the soil system is complex, and the dominant free radicals may transform in the soil (Eq. 5.14–5.16). Hydrogen peroxide can convert Cr(VI) to Cr(III). 2O2 ·− + 2H+ → O2 + H2 O2

(5.14)

3O2 ·− + Cr6+ → Cr3+ + 3O2

(5.15)

+ 3+ 2HCrO− + 3O2 + 8H2 O 4 + 3H2 O2 + 8H → 2Cr

(5.16)

5.2.3 Soil Cr(VI) Speciation The following is a list of the species of Cr that can be found in the original soil samples in the following order: exchangeable fraction makes up 61.02% of the total, followed by the fraction that is bonded to Fe-Mn oxide (13.04%), organic matter (11.61%), carbonate (11.45%), and then the residual fraction (2.9%). The percentage of carbonate-bound and exchangeable chromium in soil falls lower when the reaction time is stretched out (Fig. 5.10). When a reaction had taken place for 21 days, and there was just 1% exchangeable Cr left. The proportion of Fe-Mn oxide-bound Cr(VI) and residual Cr(VI) increased to 65 and 10%, respectively, which indicated that the bioavailability of Cr(VI) in soil after the treatment continued to decline.

5.3 Synergistic Remediation of Cr(VI) and Cationic Metal …

271

Exchangeable Bound to iron and manganese oxides Residual

Fig. 5.10 Changes in Cr speciation in soil at different reaction times

Bound to carbonates Bound to organic matter

100

Percentage (%)

80

60

40

20

0

0

3

7

14

21

Reaction time (days)

5.3 Synergistic Remediation of Cr(VI) and Cationic Metal Co-contaminated Soil Soil contamination often consists of several metal ions, including the anion chromate Cr(VI) and cationic metals like Cu(II), Ni(II), and Co(II). This pollution is primarily the result of businesses like electroplating, tannery, medicines, and metal furnishing [12–15]. Soil polluted with Cr(VI) is often remedied by converting the extremely poisonous and soluble Cr(VI) to the less harmful and insoluble Cr(III) [16, 17]. The immobilization of cationic metals, which can reduce their mobility and bioavailability by adsorption and complexation via the application of amendments, is the main method for cationic metal-contaminated soil remediation [18]. Hence, an efficient method for getting rid of Cr(VI) and cationic metals is to engage in the processes of reduction and immobilization simultaneously. In the context of this in situ systemization method, composing ZVI onto a supporting material may function as both a dispersion matrix for ZVI and an amendment for immobilizing cationic metals [19]. It is a well-known fact that when coexisting cationic metals such as Cu(II), Co(II), Ni(II), and more are present in wastewater and mixed with ZVI, they invariably precipitate onto the surface. This leads to the creation of a bimetal without any requirement for foreign metals. The destruction rate of Cr(VI) can be expedited significantly by the presence of bimetal, which reduces the activation energy required for its elimination. That is to say, and this bimetallic substance can be “in situ generated,” which means it can erase the initial pollution caused by cationic metals and increase Cr(VI) reduction simultaneously. Hydroxyapatite (HAP, Ca10 (PO4 )6 (OH)2 ) is a powerful substance that can significantly reduce the mobility and availability of positively charged metals. This is due to the exceptional ability of the phosphonates in Hydroxyapatite to attach to cationic ions, resulting in a stable metal phosphate

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5 Chemical Remediation of Chromium-Contaminated Soil

precipitate. This is made possible by the chemical formula: hydroxyapatite [19, 20]. The ZVI/HAP composite is unequivocally a bimetal composed of iron, supported by HAP and can be conveniently created on-site by attaching positively charged metals like Cu(II). This fact has been proven beyond doubt.

5.3.1 Factors Affecting the Remediation of Cr–Cu–Ni–Co–Contaminated Soil Field-contaminated soil was collected from an abandoned electroplating plant in Hebei Province (Table 5.6). The effectiveness of nZVI-HAP as a soil remediate for removing Cr(VI)-Cu(II)Ni(II)-Co(II) contamination was evaluated. The effect of the dosage of nZVI-HAP (1, 2, 3, and 5%) on the reduction of total Cr(VI) in the soil is illustrated in Fig. 5.11.As the dosage of nZVI-HAP increased, there was a discernible improvement in the total Cr(VI) remediation effect. The application of 3 or 5% nZVI-HAP can achieve a total Cr(VI) removal rate of more than 99% after 20 days, with residual total Cr(VI) concentrations of 1.49 mg kg−1 (3%) and 1.05 mg kg−1 (5%), respectively. These levels are below the Cr(VI) risk screening value for Class I land, which is set at 3.0 mg kg−1 . The first week of nZVI-HAP treatment was more successful than the following twelve weeks combined. This was caused by the first-stage binding of certain Cr(VI) tightly to soil particles and the immediate consumption of nZVI-HAP active sites. The rates at which water-soluble Cr(VI), Cu(II), Ni(II), and Co(II) are removed from the soil are depicted in Fig. 5.12, which may be found here. In the first 10 days, all of the water-soluble Cr(VI) was eliminated when the dosage of nZVI-HAP was either 3% or 5%. During ten days, the application of a test dosage of nZVI-HAP ranging from one percent to five percent can result in the total elimination of water-soluble Cu(II), Ni(II), and Co(II). Figure 5.13 shows how the amount of nZVI-HAP used affects the amount of Cr(VI), Cu(II), Ni(II), and Co(II) extracted from the soil using DTPA. After 20 days of treatment with 3% nZVI-HAP, DTPA-extractable Cr(VI) was eliminated entirely, and reasonably high removal rates of DTPA-extractable Cu(II), Ni(II), and Co(II) were also achieved. For the subsequent evaluations, a 3% nZVI-HAP was selected since it was regarded as the most valuable and cost-effective choice. Table 5.6 The properties of the soil sample Soil pH 8.37

Total Cr(VI) (mg/kg)

Water-soluble (mg kg−1 )

DTPA-extractable (mg kg−1 )

Cr(VI)

Cu

Co

Ni

Cr(VI)

Cu

Co

Ni

480

446

0.9

0.6

0.5

476.87

4.46

4.79

2.56

Reprinted from Yang et al. [21] Copyright 2023, with permission from Elsevier

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273

Fig. 5.11 Effect of nZVI-HAP dosage on the reduction of total Cr(VI) in the soil. Reprinted from Yang et al. [21] Copyright 2023, with permission from Elsevier

Fig. 5.12 Changes in water-soluble heavy metals in soil with remediation time: a Cr(VI), b Cu, c Co, and d Ni. Reprinted from Yang et al. [21] Copyright 2023, with permission from Elsevier

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Fig. 5.13 Changes in DTPA-extractable heavy metals in soil with remediation time: a Cr(VI), b Cu, c Co, and d Ni. Reprinted from Yang et al. [21] Copyright 2023, with permission from Elsevier

The remediation performance of Cr(VI), Cu(II), Ni(II), and Co(II)-contaminated soil by nZVI-HAP was also investigated as a function of the water-to-soil ratio (Fig. 5.14). [Cr(VI)] stands for chromium VI, whereas [Cu(II)] stands for copper II, [Ni(II)] stands for nickel II, and [Co The efficiency of the removal of Cr(VI) was significantly improved as the ratio of water to soil was made significantly higher. When the water-to-soil ratio was 2:1, and the removal rates were 98.56%, 99.56%, and 99.10%, respectively, the total Cr(VI), water-soluble Cr(VI), and DTPA-extractable Cr(VI) were decreased to 6.9, 1.95, and 4.31 mg kg−1 , respectively. These values represented a reduction of 98.56%, 99.56%, and 99.10%, respectively. It has been hypothesized that water acts as a crucial “medium” in the reaction process that decreases Cr(VI) in soil. Figure 5.14 depicts the impact that the ratio of water to soil has on the immobilization of water-soluble and DTPA-extractable copper, nickel, and cobalt found in soil. It was discovered that the proportion of water to soil substantially impacted the effectiveness of the nZVI-HAP restoration process. Both the cationic metals that could be extracted using water and those that could be extracted using DTPA were virtually entirely immobilized when the water-to-soil ratio was set to 2:1. As a result, the water-to-soil ratio of 2:1 was shown to be the most effective in the remediation of soil contaminated with Cr(VI)-Cu(II)-Ni(II)-Co(II) by nZVIHAP. This ratio was utilized in further studies.

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Fig. 5.14 Changes in total Cr(VI), water-soluble and DTPA-extractable Cr(VI), Cu, Co and Ni with the water-to soil ratio: a Cr(VI), b Cu, Co and Ni. Reprinted from Yang et al. [21] Copyright 2023, with permission from Elsevier

5.3.2 Sequential Extraction, Followed by an Evaluation of Risk Reduction The operationally defined speciation of all of the metals that may be found in soil was investigated in the circumstances consisting of a water-to-soil ratio of 2:1 and a dose of 3% ZVI/HAP (Fig. 5.15a). During the course of the treatment, there was a substantial reduction in the amounts of total Cr(VI), water-soluble Cr(VI), and DTPAextractable Cr(VI). The concentrations of water-soluble Cr(VI), DTPA-extractable Cr(VI), and total Cr(VI) all decreased to their respective values of 0 mg kg−1 , 0.43 mg kg−1 , and 1.49 mg kg−1 during the course of 20 days. The elimination effectiveness reached more than 99%, which means that the biological toxicity of Cr(VI) in polluted soil can be significantly reduced, as can the availability of plants. In addition, Fig. 5.16 illustrates the speciation of soil Cr in untreated virgin soil and soil treated with ZVI/HAP. The Cr speciation of the virgin soil contained a fraction of EX that was 40.97%, a fraction of CB that was 12.30%, an OX fraction that was 33.07%, an OM fraction that was 6.80%, and an RS fraction that was 6.86%. With the application of ZVI/HAP, the EX fraction dropped significantly, and after 7 days, it almost completely vanished. Throughout the remediation process, there was a rise in the OX, OM, and RS fractions, which suggests that the accessible species of Cr were converted into reasonably stable forms. It is important to note that the OX portion of Cr increased as the treatment went on, eventually reaching 58.32% after 20 days. The danger level of Cr in contaminated soil moved from being highly detrimental with a RAC value of 53.3% to being a medium risk level with a RAC value of 13.4%. This change occurred because the concentration of Cr in the soil decreased. (Table 5.7). With a clearance rate above 99%, the active Cr(VI) content was only calculated to be 0.40 mg kg−1 , corresponding to the previous statement. This behavior is most likely caused by the deposition of Cr(III) on the surface of ZVI/HAP, which forms

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the Fe-Cr compound. This compound improved the quantity bound to Fe-Mn oxide, substantially inhibiting the mobility of Cr. Also explored was how the remediation affected the current levels of copper, nickel, and cobalt (Fig. 5.15). After applying ZVI/HAP, the amount of watersoluble and DTPA-extractable copper, nickel, and cobalt in the soil would gradually decrease with increasing treatment time. Within ten days, all the water-soluble copper, nickel, and cobalt had been completely adsorbed, and the concentrations of DTPAextractable copper, nickel, and cobalt had dropped to 0.11, 0.87, and 0 mg kg−1 .

Fig. 5.15 a Changes of total Cr(VI), water-soluble Cr(VI) and DTPA-extractable Cr(VI) in soil after remediation. b Change of water-soluble state and DTPA-extractable state of the co-existing metal cations in soil after remediation. The ZVI/HAP dosage was 3%, and the water-to-soil ratio was 2:1. Reprinted from Yang et al. [21] Copyright 2023, with permission from Elsevier Fig. 5.16 The fraction distribution of Cr in soil before and after remediation. Reprinted from Yang et al. [21] Copyright 2023, with permission from Elsevier

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Table 5.7 Risk assessment codes (RACs) of heavy metals in soil samples (The ZVI/HAP dosage was 3%, and the water-to-soil ratio was 2:1) Sample

RAC (%) Cr

Cu

Co

Ni

Control soil

53.28

14.60

23.91

26.24

Remediated soil after 20 days

13.43

3.41

10.00

9.60

Reprinted from Yang et al. [21] Copyright 2023, with permission from Elsevier

According to these findings, ZVI/HAP has the capability to efficiently adsorb the water-soluble and DTPA-extractable cationic metals that were present in the soil. Additionally, ZVI/HAP was able to do so without releasing any of the adsorbates into the environment. The precipitation of phosphate was the primary factor that led to the high adsorption efficiency of the cationic metals. Phosphate precipitation occurred when cationic metals in the soil that were unstable or leachable interacted with phosphate on ZVI/HAP. Figure 5.17 depicts the changes in the species of copper, nickel, and cobalt as a function of the treatment period. The application of ZVI/HAP may transition the unstable state (EX and CB) of Cu, Ni, and Co into a more stable state, which will lead to a decrease in the amount of unstable state as the treatment time progresses. The extraction efficiency (EX) and concentration bias (CB) of copper in virgin soil were 9.71% and 4.89%, respectively. After twenty days, the EX fraction was reduced to 0.49%, while the CB fraction was reduced to 2.93% thanks to the inclusion of ZVI/HAP. Likewise, the RAC values for Cu, Co, and Ni in polluted soil were 14.6%, 23.9%, and 26.2%, respectively (Table 5.7). These values correspond to risk levels that are medium in severity. The RAC values for the soil were 3.4%, 10.0%, and 9.6% for copper, cobalt, and nickel, respectively, after 20 days of remediation, indicating that the soil posed a low danger. As a consequence of this, ZVI/HAP has the potential to reduce the mobility and bioavailability of Cu, Ni, and Co, so producing the remediation impact that was intended. This finding was consistent with our theory, which claimed that the presence of soluble cationic metals bound onto ZVI/HAP would make the adsorption of cationic metals in unstable states easier, hence transforming those states into ones that are relatively more stable. This result was consistent with our hypothesis. In addition, we utilized the RAC risk assessment approach to quantify the shifts that occurred in the amount of active heavy metals present. The results indicated that when ZVI/HAP was added, the active concentration of copper and nickel dropped to 0.10 mg kg−1 , 0.85 mg kg−1 , and 0.01 mg kg−1 , respectively. These findings are comparable to the results obtained from the DTPA-exchangeable condition. The removal rates each reached their respective peaks of 97.78, 82.37, and 99.62%. In general, ZVI/HAP can simultaneously adsorb Cu, Ni, and Co in polluted soil while also performing reductively effective remediation of Cr(VI).

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5 Chemical Remediation of Chromium-Contaminated Soil

Fig. 5.17 The fraction distributions of Cu, Co and Ni in soil before and after remediation: a Cu, b Co, and c Ni. Reprinted from Yang et al. [21] Copyright 2023, with permission from Elsevier

5.4 Chemical-Microwave Combined Remediation of Strongly Alkaline High Cr(VI)-Contaminated Soil Chemical reduction is now the technology most widely employed for remediating soil polluted with Cr(VI). This is because chemical reduction requires less expensive treatment and has a broader range of applications: However, it is important to note that there are significant challenges associated with utilizing this technology [22, 23]. Firstly, residual concentrations of Cr(VI) can reach up to 60 mg kg−1 even after the remediation process, necessitating additional cleanup efforts. Secondly, the remediation process is quite time-consuming and can take over 15 days to complete. Thirdly, the long-term stability of the process is low due to Cr(III) deoxidation in complex soil systems. The highly alkaline, high Cr(VI)-contaminated soil (SAHCR) is particularly concerning, with a soil pH above 10.0 and a Cr(VI) content of 3000– 10,000 mg kg−1 , posing a significant threat to the environment and public health. Therefore, the cleanup process is incredibly challenging and must be approached with great care and attention. It is difficult for CrO4 2− to compete with OH− for anionic sites in SAHCR soil because the soil is prone to deprotonation [24]. Because of this, in contrast to other soils contaminated with Cr(VI), CrO4 2− is challenging to adsorb to the soil surface and is typically thought to be imprisoned inside the structure or pores of soil minerals [25, 26]. In addition, in order to meet the more demanding remediation criteria (a Cr(VI) risk screening value of 33.0 mg kg−1 for Class I land), we need to comply with this standard., [27]the Cr(VI) removal rate for SAHCR soil should be greater than 99.9% if the Cr(VI) concentration is higher than 3000 mg kg−1 . This indicates that the restoration of SAHCR soil requires materials with a high reduction capacity and a quick reaction rate. However, reduction materials that are typically employed, such as materials that contain iron, suffer from restricted reduction capacity due to their surface passivation and poor ability to donate electrons [1, 25, 28]. In addition,

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279

the typical reduction kinetics of Cr(VI) in the soil is characterized by the fact that it demonstrates a two-stage process: a quick reduction of Cr(VI) in the first three days, followed by a much slower reduction that can last for up to many months [29, 30]. As a result, it is complicated to remove 99.9% of the Cr(VI) in a short period; in particular, some of the Cr(VI) is found in the deep pores of the soil or is closely bound with the soil particles. It has the potential to act as a reduction material in the process of SAHCR soil remediation because zerovalent iron-loaded pyrite (FeS2 /ZVI) displays more excellent reduction capability and robust adaptability under alkaline soils. As a consequence, it has the ability to serve as a reduction material [4]. The following inquiry concerns how the response rate of Cr(VI) in the soil can be increased. Due to the molecular-level heating that it produces, microwave irradiation has recently garnered much attention in soil remediation, particularly in the remediation of organically polluted soil [31, 32]. Nonetheless, there has been a limited amount of research conducted on the use of microwave irradiation in cleaning up soil polluted with heavy metals. It has been observed that the reduction of Cr(VI) is an endothermic process that occurs spontaneously [33]. In this context, the combination of microwave irradiation and fast heating is beneficial for accelerating the reduction of Cr(VI) in soil. The idea behind microwave irradiation is that the material being heated is a dielectric material with a high dielectric constant and can effectively absorb microwave energy and convert it into heat energy [34–37]. This is the foundation upon which microwave irradiation is based. The presence of magnetic characteristics in FeS2 / ZVI enables it to function as an absorber of microwave irradiation, making it possible to achieve quick and selective heating when the material is subjected to microwaves. It is necessary for the “hot spots” on the surface of the FeS2 /ZVI material to be activated by microwave radiation in order to hasten the interaction between Cr(VI) and FeS2 /ZVI. It is also important to note that subjecting heavy metals to microwave irradiation can increase the crystallinity and size of the metals, as well as fix them on a transition metal lattice that has excellent microwave absorption capabilities [15, 38]. This contributes to the metals’ increased degree of stability. In a similar vein, we may predict that stable forms of Cr(III), such as chromite (FeCr2 O4 ), will be generated during the reduction of Cr(VI) by FeS2 /ZVI in the presence of microwave irradiation. This is due to the fact that chromite is an example of a compound that contains both Cr(III) and FeCr2 O4 . Because of this, we will be able to tackle the problem of poor long-term stability of remediated Cr(VI) soil. This problem arises when unstable Cr(III) forms, such as Cr(OH)3 , are the reduction result. This will allow us to fix the problem. Therefore, in order to accomplish quick and efficient soil remediation along with long-term stabilization for SAHCR, we suggest a method that combines microwave irradiation with FeS2 /ZVI in order to act as both a reduction material and a microwave-absorbing material. This will allow us to achieve our rapid and effective soil remediation goals.

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5 Chemical Remediation of Chromium-Contaminated Soil

5.4.1 Comparison of Remediation Effect by Various Amendments An old industrial site in Jinan, Shandong Province, China, contaminated with chromium chemicals was used to collect soil samples. The overall concentration of chromium in the soil, denoted by Cr(Crtotal ), is 18,959 mg kg−1 , whereas the concentration of chromium in the soil, denoted by Cr(VI), is 3900.8 mg kg−1 . The pH of the soil is 10.5, which indicates that it is a highly alkaline soil (Table 5.8). In the first step of this research project, the effectiveness of pyrite, ZVI, FeSO4 , and FeS2 /ZVI in restoring SAHCR soil was evaluated (Fig. 5.19). The use of many reduction materials resulted in a gradual drop in the soil’s Cr(VI) concentration throughout treatment. The fact that the effectiveness of Cr(VI) removal was discovered to be in the sequence of FeS2 /ZVI > FeSO4 > FeS2 > ZVI should not come as a surprise to anybody. Following a treatment period of thirty days, the Cr(VI) removal efficiencies of FeS2 /ZVI, FeSO4 , FeS2 , and ZVI with a dose of five percent were, respectively, 98.95, 95.46, 63.29, and 33.09%. The Cr(VI) concentrations that remained after the reaction were, in order, 40.9, 177.14, 1432.11, and 2610.18 mg kg−1 , all of which were much greater than the limit value of 3 mg kg−1 for Cr(VI). In addition, even though the dosage of reduction materials was increased to 10%, their repair performance was not a discernible improvement. According to these findings, it appears it will be nearly impossible for the SAHCR soil to meet the Cr(VI) remediation standard by relying solely on reduction materials. This is primarily because a single restoration material’s reduction capacity is restricted when exposed to high levels of alkalinity. Changes in the content of Cr(VI) in the soil can be seen in Fig. 5.19b, which depicts the effects of microwave irradiation on various reduction materials. After only ten minutes of exposure to microwaves, the remediation efficiencies had already reached or even significantly exceeded those of the reduction material alone for thirty days (Fig. 5.18c). It has been demonstrated that subjecting SAHCR soil to microwave irradiation can both hasten the reduction of Cr(VI) and bring about a more significant decrease in its concentration. After being exposed to microwaves for ten minutes, the Cr(VI) concentration decreased from 3900.8 mg kg−1 to 2.38 mg kg−1 . Additionally, the enhancing influence of ZVI and FeS2 /ZVI is more noticeable. Specifically, the enhancing impact of FeS2 /ZVI is more striking than its counterpart. Even after 15 min of treatment, the lack of microwave irradiation resulted in a Cr(VI) removal rate of less than 40% for the FeS2 /ZVI system. This rate was much lower than the 99% attained by the combination of microwaves and FeS2 /ZVI, which was significantly higher. It is plausible that the high magnetism of ZVI and FeS2 /ZVI has anything to do with this event, and there is a good chance that it does. In general, magnetic materials coupled with other components of various microwave loss processes are favorable to remarkable microwave absorption performance [40]. This is because magnetic materials are good at attracting and holding onto microwaves. The saturation magnetization intensity (Ms) of FeS2 /ZVI (3.68 emu g−1 ) was more than 40 times that of FeS2 (0.09 emu g−1 ), confirming that the strong magnetism could be

10.05

Value

38.9 cmolc ·kg−1

Cation exchange capability (CEC)

0.66%

Total organic carbon (TOC) 388 mS·m−1

Electrical conductivity (EC)

Reprinted from Li et al. [39] Copyright 2022, with permission from Elsevier

pH

Property

Table 5.8 Physical and chemical properties of soil samples

3241.2 mg·kg−1

Water soluble Cr(VI)

3900.8 mg·kg−1

Total Cr(VI)

18,959 mg·kg−1

Total Cr

5.4 Chemical-Microwave Combined Remediation of Strongly Alkaline … 281

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5 Chemical Remediation of Chromium-Contaminated Soil

Fig. 5.18 The remediation efficiency of FeS2 , ZVI, FeSO4 · 7H2 O and FeS2 /ZVI with 5% or 10% dosage on the SAHCR soil (a), the enhanced role of microwave irradiation on Cr(VI) removal at a 10% dosing rate (b), the remediation efficiency of the four materials at different dosing rates with and without microwave irradiation (c), magnetic hysteresis loop of FeS2 and FeS2 /ZVI (d), the zeta potential of the FeS2 , ZVI, FeS2 /ZVI under different pH conditions (e) and the pH of the remediated soil (f). Reprinted from Li et al. [39] Copyright 2022, with permission from Elsevier

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283

Fig. 5.19 Effect of FeS2 /ZVI dosage (a), microwave irradiation power (b), microwave irradiation time (c), and soil-water ratio (d) on the advanced remediation of SAHC soil. Reprinted from Li et al. [39] Copyright 2022, with permission from Elsevier

the reason why magnetic materials exhibit a more remarkable enhancement effect of microwave irradiation on soil Cr(VI). The magnetization hysteresis loops of FeS2 and FeS2 /ZVI are shown in Fig. 5.18d. The materials’ zeta potential is illustrated under various pH conditions in Fig. 5.18e. Both pyrite and ZVI had isoelectric points of 4.7, although pyrites were lower. ZVI’s was higher. Pyrite and ZVI were both ineffective as remediation agents for the SAHCR soil because their surfaces were prone to deprotonation and electrostatic repulsion when exposed to CrO4 2− in alkaline conditions. For instance, the potential of FeS2 /ZVI was positive from pH 1 to 10, with the exception of pH = 2, which indicates that the surface of FeS2 /ZVI was positively charged and favorable to electrostatic attraction towards the CrO4 2− anion, particularly under circumstances that were very alkaline. This explains why ZVI/FeS2 had higher efficacy in soil remediation when the soil pH was 10.1. Upon the remediation’s completion, the soil’s pH was also measured (Fig. 5.18f). The soil’s pH dropped significantly after applying FeS2 or ZVI, although it was still alkaline or very close to being extremely alkaline (pH = 9.15). Because of the relatively good Cr(VI) removal impact that FeSO4 had on the SAHCR soil, soil acidification occurred even though the soil pH

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5 Chemical Remediation of Chromium-Contaminated Soil

dropped dramatically from 10.1 to below 4.03 due to the application of FeSO4 . Interestingly, the soil pH for FeS2 /ZVI fell to a near-neutral level (pH = 7.49), which was conducive to restoring soil function. This occurred whether or not microwave irradiation was used. In addition to almost wholly eliminating soil Cr(VI), the microwave irradiation-assisted FeS2 /ZVI will also successfully regulate soil alkalinity, making this technique an exceptionally appealing option for restoring SAHCR soil.

5.4.2 Factors Affecting Microwave Irradiation-Assisted Reduction The study investigated the impact of soil-to-water ratio, microwave irradiation power, microwave irradiation period, and FeS2 /ZVI dosage on the removal of FeS2 /ZVI from SAHCR soil using microwave-assisted techniques, as depicted in Fig. 5.19. The enhancement of Cr(VI) reduction and the significant decrease in Cr(VI) content in the soil were both the result of an increase in the dose of FeS2 /ZVI, which caused there to be a more significant number of active sites of hot spots under the microwave irradiation. When the dose was fixed at 10%, the Cr(VI) concentration left behind was barely 2.38 mg kg−1 . Irradiation power from microwaves considerably impacted the elimination of Cr(VI), as well (Fig. 5.19b). As the strength of the microwave irradiation was raised from 280 to 560 mg kg−1 , the concentration of Cr(VI) decreased from 111.06 to 2.38 mg kg−1 in a span of only ten minutes. On the other hand, the concentration of Cr(VI) was as high as 3862 mg kg−1 for the same amount of time that it was subjected to microwave irradiation treatment when the intensity of the microwave irradiation was 560 W but there was no addition of FeS2 /ZVI. The findings above demonstrated that removing Cr(VI) could be improved by combining microwave irradiation and FeS2 /ZVI. The ideal microwave irradiation power of 560 W is also much lower than that previously reported (800 W) [41], which is conducive to energy savings because it reduces the energy needed. The effect of the microwave irradiation period on the removal of Cr(VI) is depicted in Fig. 5.19c. This figure shows that the two stages of Cr(VI) removal that occur the quickest are the beginning phase (0–3 min) and the ending phase (7–10 min). At the beginning of the process, a significant quantity of the water-soluble form of Cr(VI) was swiftly eliminated from the system. On the other hand, the encapsulated form of Cr(VI) that was more difficult to access required more extended energy to accumulate before the reduction reaction could be initiated. Both the reaction medium and the irradiation absorber for microwaves are water. Hence, raising the water content suitably can reduce the time required for microwave irradiation (Fig. 5.19d). In the FeS2 /ZVI system that was helped by microwave irradiation, the best ratio of soil to water was found to be 1:1 g: ml for removing Cr(VI).

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5.4.3 Mechanism of Microwave Irradiation Accelerating Cr(VI) Reduction In order to investigate the reaction process in the soil, three sets of experiments were designed to test the kinetics of the reduction of Cr(VI) at varying temperatures: (1) temperature in the FeS2 /ZVI system ranging from 293 to 308 K at normal conditions; (2) temperature in the FeS2 /ZVI system ranging from 333 to 353 K at heat conditions; and (3) temperature in the FeS2 /ZVI system ranging from 333 to 353 K at microwave irradiation-assisted conditions. The Langmuir–Hinshelwood first-order model [42] (Eq. 5.1) was utilized in order to characterize the kinetics of the process adequately. Calculations were made using the Arrhenius equation (Eq. 5.4) and the thermodynamic equations (Eq. 5.5–5.6) to determine the activation energies (Ea) and thermodynamic parameters that correspond to each system. Table 5.9 contains representations of all of the kinetic and thermodynamic characteristics. At temperatures of 293, 298, 303, and 308 K, respectively, the reaction rate constants were 9, 11, and 17.8 m min−1 . This can be seen in Fig. 5.20a, b. The activation energy (E a ) was 36.6 kJ mol−1 , which is significantly lower than the normal activation energy range of 60–250 kJ mol−1 for chemical reactions. This suggests that FeS2 /ZVI has a considerable affinity for Cr(VI) in the SAHCR soil, and the Cr(VI) reduction process still takes place with a fair amount of ease. In Series 2 (Fig. 5.20c, d), the heat temperature range was the same as the one used for the microwave irradiation procedure. This was done to remove the influence that temperature had on the amount of energy required to activate the reaction. In Series 2, the rate of Cr(VI) removal was much higher than in Series 1, which was the previous series. The reaction rate constants at 333, 343, and 353 K, respectively, were significantly greater than those at a lower temperature (293–208 K), which suggests that the heating impact was one of the reasons for the accelerating Cr(VI) reduction. The reaction rate constants were 175, 250, and 384 m min−1 , respectively. Accordingly, the Ea that was obtained was 38.4 kJ mol−1 , which was comparable to that which was found for Series 1. This demonstrates that the Ea of the reaction typically does not change with temperature [43]. The impact of temperature on the Cr(VI) removal rate in the microwave-irradiation-assisted FeS2 / ZVI system is seen in Fig. 5.20e, f, respectively. The reaction rate constants were found to be 1.5, 1.3, and 1.2 times higher at 333, 343, and 353 K, respectively, than those at the corresponding temperature without microwave irradiation. This suggests that there may be a synergistic effect between FeS2 /ZVI and microwave irradiation. The reaction rate constants were found to be 261, 326, and 458 m min−1 , respectively. The Ea was found to be 27.4 kJ mol−1 , which is 28.5% lower than that for Series 2, suggesting that microwave irradiation may really reduce the Ea of the reaction. This is the direct cause for the greater removal efficiency seen under microwave irradiation conditions. The large decrease in Ea that occurs as a result of microwave irradiation can be due to the catalytic effect that FeS2 /ZVI has when subjected to microwave irradiation. This effect lowers the potential barrier that exists during the Cr(VI) reduction process, which in turn simplifies the reaction. Therefore, the rapid reduction of Cr(VI) was attributed to both the thermal effect generated by microwave irradiation

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and the catalytic action of FeS2 /ZVI under microwave irradiation. This conclusion was reached as a result of the previous sentence. In order to investigate the reaction mechanism in the soil, three series of experiments on the kinetics of Cr(VI) reduction at varying temperatures were designed and carried out: (1) the average temperature in the FeS2 /ZVI system, which ranges from 293 to 308 K; (2) the heat temperature in the FeS2 /ZVI system, which ranges from 333 to 353 K; and (3) the temperature in the FeS2 /ZVI system that has been helped by microwave irradiation, which also ranges from 333 to 353 K. The Langmuir–Hinshelwood first-order model was used to characterize the kinetics of the process [42] (Eq. 5.1). Using the Arrhenius equation (Eq. 5.4) and the thermodynamic equations, we were able to determine the relevant activation energies (Ea) as well as the thermodynamic parameters (Eq. 5.5 and 5.6). The table containing all the kinetic and thermodynamic parameters is numbered 5.9. At temperatures of 293, 298, 303, and 308 K, respectively, the reaction rate constants were 9, 11, and 17.8 mmin−1 when examining Series 1 (Fig. 5.20a, b). The activation energy (Ea) was 36.6 kJ mol−1 , significantly lower than the conventional activation energy range of 60–250 kJ mol−1 for chemical reactions. This would imply that FeS2 /ZVI has a considerable affinity for Cr(VI) in the SAHCR soil and that the Cr(VI) reduction process still takes place at a reasonable rate despite this. The temperature range that was used for the heat treatment in Series 2 (Fig. 5.20c–d) was the same as the temperature range that was used for the microwave irradiation procedure. This was done in order to eliminate the effect that temperature has on the activation energy of the reaction. The clearance rate of Cr(VI) in the second series was much greater when compared to the rate seen in the first series. The fact that the reaction rate constants at 333, 343, and 353 K, respectively, were significantly greater than those at a lower temperature (293-208 K) lends credence to the hypothesis that the heating effect was one of the factors contributing to the accelerated reduction of Cr(VI). At each temperature, the reaction rate constants varied between 175, 250, and 384 min−1 . Therefore, the Ea that was determined to be 38.4 kJ mol−1 was found to be comparable to that which was discovered for Series 1. This indicates that the Ea of a reaction does not change much with the temperature in most cases [43]. Temperature’s influence on the Cr(VI) removal rate in the FeS2 / ZVI system that is helped by microwave irradiation is depicted in Fig. 5.20e, f. The reaction rate constants were 1.5, 1.3, and 1.2 times higher at 333, 343, and 353 K, respectively, when microwave irradiation was present. This suggests that there may be a synergistic effect between FeS2 /ZVI and microwave irradiation. The Ea was found to be 27.4 kJ mol−1 , which is 28.5% lower than that for Series 2, suggesting that microwave irradiation can reduce the E a value of the reaction. This is the direct cause for the greater removal efficiency seen under microwave irradiation conditions. The catalytic impact of FeS2 /ZVI under microwave irradiation can be related to the considerable reduction in E a that occurs due to microwave irradiation. This effect lowers the potential barrier during the Cr(VI) reduction process, making the reaction more straightforward. Therefore, it was determined that the fast reduction of Cr(VI) was due to both the thermal impact of microwave irradiation and the catalytic action of FeS2 /ZVI while it was being subjected to microwave irradiation. This conclusion was arrived at as a consequence of the statement that came before it.

11.0 16.2 17.8

298

303

308

261 326 458

353

384

353

343

250

343

333

175

333

1.51

2.13

2.66

1.80

2.78

3.96

38.9

42.8

63.0

77.0

The half-life periods (min)

Reprinted from Li et al. [39] Copyright 2022, with permission from Elsevier

Microwave irradiation

Heat treatment

9.00

293

Normal temperature

K obs (mmin−1 )

Kinetics parameters

Temperature (K)

Experiment Series

Table 5.9 Kinetics and thermodynamics parameters under three systems

0.997

0.998

0.999

0.995

0.991

0.995

0.996

0.995

0.975

0.996

R2

27.4

38.4

36.6

Ea (kJ mol−1 )

0.458

0.326

0.261

0.384

0.250

0.175

0.0178

0.0162

0.0110

0.009

lnK D (min−1 )

− 1.34 × 103

− 928

− 722

− 1.13 × 103

− 712

− 484

− 45.6

− 40.8

9.59

10.2

474

− 21.9 − 27.3

ΔH 0 (J mol−1 )

ΔG0 (J mol−1 )

Thermodynamics parameters

30.9

32.0

1.69

ΔS 0 (J mol−1 )

5.4 Chemical-Microwave Combined Remediation of Strongly Alkaline … 287

288

5 Chemical Remediation of Chromium-Contaminated Soil

Fig. 5.20 The Effect of temperature on Cr(VI) reduction rate (a, c, e) and kinetics at different temperatures (b, d, f) (a, b: normal temperature condition, c, d: heat treatment processes, e, f: microwave irradiation). Reprinted from Li et al. [39] Copyright 2022, with permission from Elsevier

5.4.4 Cr Speciation and Phase Transformation Under Microwave Irradiation At a variety of aging periods, analyses of soil Cr speciation were carried out with and without the use of microwave irradiation. The Cr species in the untreated contaminated soil had the following percentages: 31.46% for the exchangeable (EX), 5.34%

5.4 Chemical-Microwave Combined Remediation of Strongly Alkaline …

289

for the carbonate-bound (CB), 41.43% for the Fe–Mn oxide-bound (OX), 17.77% for the organic matter-bound (OM), and 4.00% for the residual (RS). This suggests that a significant amount of Cr(VI) may be in a more stable form (OX + OM + RS accounting for 63.2%) entrap. Within three days of treatment with FeS2 /ZVI alone, the proportion of the accessible state of Cr (the EX and CB fraction) reduced from 36.80% to 4.91%, suggesting that the available fraction of Cr was changed into more stable states. This was observed in the EX and CB fraction. Because Cr(III) and Fe(III) coprecipitated together to create Crx Fe1–x (OH)3 , the speciation distribution of chromium in the soil shifted when the time period was extended to 30 days. This shift occurred mostly in the form of oxygen, which accounted for 79.38% of the total. When further microwave irradiation was applied, the EX and CB fractions underwent a further transformation into stable species with a proportion of around 0.5% (Fig. 5.21b). This proportion was significantly lower than that which was achieved by treating FeS2 /ZVI by itself. Because there is only a minimal amount of accessible Cr after being subjected to microwave irradiation, the danger of Cr rerelease and oxidation is relatively minimal. In other words, microwave irradiation helps the stability of remediated Cr in the soil while also significantly improving the remediation performance of FeS2 /ZVI for SAHCR soil. It is hypothesized that an acceleration of the mineral phase transition of chromium-containing phases in response to microwave irradiation caused the increased stability of chromium in soil. This mineral phase transformation may have included Cr doping into the lattice of ferrite and the production of stable minerals. At a variety of aging periods, analyses of soil Cr speciation were carried out with and without the use of microwave irradiation. The Cr species in the untreated contaminated soil had the following percentages: 31.46% for the exchangeable (EX), 5.34% for the carbonate-bound (CB), 41.43% for the Fe-Mn oxide-bound (OX), 17.77% for the organic matter-bound (OM), and 4.00% for the residual (RS). This suggests that a significant amount of Cr(VI) may be in a more stable form (OX + OM + RS accounting for 63.2%) entrap. Within three days of treatment with FeS2 /ZVI alone, the proportion of the accessible state of Cr (the EX and CB fraction) reduced from 36.80% to 4.91%, suggesting that the available fraction of Cr was changed into more stable states. This was observed in the EX and CB fraction. Because Cr(III) and Fe(III) coprecipitated together to create Crx Fe1–x (OH)3 , the speciation distribution of chromium in the soil shifted when the time period was extended to 30 days. This shift occurred mainly in the form of oxygen, which accounted for 79.38 percent of the total. When further microwave irradiation was applied, the EX and CB fractions underwent a further transformation into stable species with a proportion of around 0.5% (Fig. 5.21b). This proportion was significantly lower than that which was achieved by treating FeS2 /ZVI by itself. Because there is only a minimal amount of accessible Cr after being subjected to microwave irradiation, the danger of Cr rerelease and oxidation is relatively minimal. In other words, microwave irradiation helps the stability of remediated Cr in the soil while also significantly improving the remediation performance of FeS2 /ZVI for SAHCR soil. It is hypothesized that an acceleration of the mineral phase transition of chromium-containing phases in response to microwave irradiation caused the increased stability of chromium in

290

5 Chemical Remediation of Chromium-Contaminated Soil

Fig. 5.21 Speciation changes of Cr in FeS2 /ZVI (a) and microwave-FeS2 /ZVI (b) systems after different times of maintenance, XRD patterns of samples collected after magnetic separation in FeS2 /ZVI and microwave-FeS2 /ZVI systems (c), and Cr 2p XPS pattern of sample collected after magnetic separation in microwave-FeS2 /ZVI system (d) (microwave irradiation power: 560 W, microwave irradiation time: 10 min, FeS2 /ZVI dosage: 10%, water content: 1:1 g: ml). Reprinted from Li et al. [39] Copyright 2022, with permission from Elsevier

soil. This mineral phase transformation may have included Cr doping into the lattice of ferrite and the production of stable minerals. When Cr(III) and Fe(III) coprecipitate (Crx Fe1-x (OH)3 ), the soil’s distribution of the Cr speciation underwent a shift when the period was increased to 30 days. This change occurred primarily in the form of OX (accounting for 79.38%). When additional microwave irradiation was applied, the EX and CB fractions were transformed into stable species with a proportion of roughly 0.5% (Fig. 5.21b). This was a significantly smaller percentage than that achieved through the treatment of FeS2 /ZVI alone. In the presence of microwave irradiation, the meager fraction of accessible Cr ensures that there is a low chance of Cr rerelease and oxidation. In other words, microwave irradiation considerably increases the remediation effectiveness of FeS2 /ZVI for SAHCR soil. It also contributes to the stability of the Cr that has been remediated in the soil. It is hypothesized that an acceleration of the mineral phase transformation of chromium-containing phases in response to microwave irradiation caused the improved stability of chromium in soil.

5.4 Chemical-Microwave Combined Remediation of Strongly Alkaline …

291

This mineral phase transformation includes the incorporation of chromium into the lattice of ferrite and forming stable minerals. Following the recovery of the product of the FeS2 /ZVI reaction using magnetic separation, it was analyzed using XRD and XPS. This was done so that the phase composition of soil Cr could be better understood. In the FeS2 /ZVI system, only the typical characteristic peaks of Fe0 were detected (Fig. 5.21c), which indicates that no crystalline Cr phase was produced without the use of microwave irradiation. This lends credence to the idea that the Cr(III)/Fe(III) coprecipitate that was generated in the FeS2 /ZVI system was amorphous. The creation of the FeCr2 O4 phase in the microwave-FeS2 /ZVI system was indicated by the appearance of peaks at 30.15, 35.5, 43.1, 57.1, and 62.7 degrees after incorporating microwave irradiation [44, 45]. These peaks were positioned at respective angles of 30.15, 35.5, 43.1, and 57.1. The presence of the Cr 2p peak at 576.97 eV in the XPS survey spectrum (Fig. 5.22) further indicated the deposition of Cr on the surface of the FeS2 /ZVI material subjected to microwave irradiation. The peaks at 579.0, 577.9, and 577 eV in the Cr 2p XPS spectra (Fig. 5.21d) corroborate the synthesis of FeCr2 O4 [46, 47], hence proving the development of the FeCr2 O4 phase. Crystalline FeCr2 O4 is more stable (Ksp = 1.8 × 10–36 ) when compared to amorphous Cr(III)/Fe(III) coprecipitation, and Cr ions are not liberated even in an exceedingly varied environment [48]. FeS2 /ZVI possesses a number of different reducing groups that are able to react with Cr(VI), including Fe0 , S2 2− , and Fe2+ (Eqs. 5.17–5.19). In addition to this, the better pH control ability of FeS2 /ZVI for alkaline soil was primarily due to the oxidation of FeS2 (Eq. 5.20), the cyclic process of regeneration of FeS2 with Fe(III) to Fe(II) (Eq. 5.21), and the coprecipitation of Cr(III)-Fe(III) in the presence of Fe(III) (Eq. 5.22). The reduced Cr(III) preferentially interacts with the active Fe(II) in the Fig. 5.22 XPS survey spectrum of the sample recovered after magnetic separation in the microwave-FeS2 /ZVI system. Reprinted from Li et al. [39] Copyright 2022, with permission from Elsevier

292

5 Chemical Remediation of Chromium-Contaminated Soil

microwave- FeS2 /ZVI system to create FeCr2 O4 , rather than Crx Fe(1−x) (OH)3 , in order to make FeCr2 O4 (Eq. 5.23). This is because of the alkaline environment and the reinforcing influence of microwave irradiation. 3Fe0 + 2CrO4 2− + 16H+ → 2Cr3+ + 3Fe2+ + 8H2 O

(5.17)

3S2 2− + 14CrO4 2− + 64H+ → 14Cr3+ + 6SO4 2− + 32H2 O

(5.18)

3Fe2+ + CrO4 2− + 8H+ → Cr3+ + 3Fe3+ + 4H2 O

(5.19)

2FeS2 + 7O2 + 2H2 O → 2Fe2+ + 4SO4 2− + 4H+

(5.20)

FeS2 + 14Fe3+ + 8H2 O → 15Fe2+ + 2SO4 2− + 16H+

(5.21)

(1 − x)Fe3+ + xCr3+ + 3H2 O → Crx Fe(1−x) (OH)3 + 3H+ MW

Fe2+ + 2Cr3+ + 8OH− → FeCr2 O4 + 4H2 O

(5.22) (5.23)

5.4.5 Long-Term Stability of Cr in Remediated Soil TCLP, SPLP, and PBET were the three methodologies utilized to evaluate the longterm stability of the remediated soil created by the microwave-FeS2 /ZVI system and the FeS2 /ZVI system. In order to facilitate a point-by-point comparison with FeS2 / ZVI, the sample nZVI, which possesses a high magnetic susceptibility, was used (Fig. 5.23). The untreated soil has a Cr(VI) value of approximately 154.27 mg L−1 after being leached with TCLP. After remediation with the nZVI and microwavenZVI systems, the concentration of TCLP leaching Cr(VI) steadily reduced to 92.71 and 73.6 mg L−1 within 30 days, respectively, and remained nearly unchanged for the following 330 days. The TCLP leaching Cr(VI) values from these two treatments were significantly more significant than the United States federal regulatory requirement of 5 mg L−1 , which further confirms that nZVI is not appropriate for the remediation of SAHCR soil even when it is subjected to microwave irradiation. After being treated with FeS2 /ZVI, the leaching concentration of Cr(VI) dropped significantly to 3.96 mg L−1 after three days, and it remained below 3 mg L−1 for the subsequent 360 days. After a period of three hundred and sixty days, the addition of microwave irradiation brought the total Cr(VI) concentration of the TCLP

5.4 Chemical-Microwave Combined Remediation of Strongly Alkaline …

293

leaching solution down to less than 0.5 mg L−1 . This suggests that the application of microwave irradiation led to an increase in the degree to which the remediated soil retained its stability. The leaching Cr(VI) concentrations from the SPLP and PBET tests in the FeS2 /ZVI system and microwave-FeS2 /ZVI system were much lower than those in the nZVI system and microwave-nZVI system (Fig. 5.23b, c); in particular, the leaching Cr(VI) concentrations in the microwave-FeS2 /ZVI system were consistently lower than 0.2 mg L−1 during 360 d, and its bioavailability was 42 times lower (Fig. 5.24).

Fig. 5.23 Extraction toxicities of Cr(VI) and Crtotal in four systems under different evaluation methods (MW-nZVI: Microwave-nZVI, MW-FeS2 /ZVI: Microwave- FeS2 /ZVI). Reprinted from Li et al. [39] Copyright 2022, with permission from Elsevier Fig. 5.24 Reduction of PBET-based bioaccessibility of soil-bound Cr(VI) for soil amended with FeS2 /ZVI at various treatment systems. Reprinted from Li et al. [39] Copyright 2022, with permission from Elsevier

294

5 Chemical Remediation of Chromium-Contaminated Soil

Bioaccessibilit y o f Cr(VI) =

(C P B E T , mg L−1 )(0.01 L) × 100% (Q, mg L−1 )(0.1 g)

(5.24)

where C PBET is the concentration of Cr(VI) in PBET solution and Q is the content of Cr(VI) in the soil sample before PBET extraction. In the TCLP, SPLP, and PBET tests, a phenomenon known as leaching Crtotal was found. This behavior was quite similar to the leaching Cr phenomenon (VI). The levels of Crtotal leaching in the soil that had been cleaned up by nZVI, with or without the use of microwave irradiation, were relatively high. These concentrations were noticeably higher than what was discovered in soil that had been treated with FeS2 / ZVI. It is interesting to note that the Total leaching concentration in the microwaveFeS2 /ZVI system was very similar to the equivalent Cr(VI) leaching concentration. This reveals that the reduction of Cr(VI) was expedited and encouraged by microwave irradiation and that the treatment successfully immobilized Cr(III). The creation of crystal FeCr2 O4 , which was the primary cause of the immobilization of Cr(III), was principally responsible for this result. In the TCLP, SPLP, and PBET experiments, the leaching concentrations of total Cr were less than 0.5 mg L−1 within 360 days (Fig. 5.24d–f), and there was no reoxidation of Cr(III). This was seen in the system, including microwaves and FeS2 /ZVI. This suggests that the remediated SAHCR soil exhibits more excellent long-term stability even in certain extremely unfavorable situations, such as rain, which is typically acidic.

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45. D. Lv, J. Zhou, Z. Cao, J. Xu, Y. Liu, Y. Li, K. Yang, Z. Lou, L. Lou, X. Xu, Mechanism and influence factors of chromium(VI) removal by sulfide-modified nanoscale zerovalent iron. Chemosphere 224, 306–315 (2019) 46. S.R. Chowdhury, E.K. Yanful, A.R. Pratt, Chemical states in XPS and Raman analysis during removal of Cr(VI) from contaminated water by mixed maghemite–magnetite nanoparticles. J. Hazard. Mater. 235–236, 246–256 (2012) 47. H. Yang, Z. Li, P. Fu, G. Zhang, Cr(VI) removal from a synthetic solution using a novel carbonaceous material prepared from oily sludge of tank bottom. Environ. Pollut. 249, 843–850 (2019) 48. C. Li, J. Yu, W. Li, Y. He, Y. Qiu, P. Li, C. Wang, F. Huang, D. Wang, S. Gao, Immobilization, enrichment and recycling of Cr(VI) from wastewater using a red mud/carbon material to produce the valuable chromite (FeCr2 O4 ). Chem. Eng. J. 350, 1103–1113 (2018)

Chapter 6

Case of Chrome-Contaminated Site Remediation Project

6.1 Project Overview The remediation project was carried out in a ferroalloy enterprise that began to produce element chromium in the early 1960s, with an annual output of 3000 tons of chromium. Due to serious chromium pollution, the chromium production workshop was shut down in 1996. After nearly 40 years of production, the company accumulated chromium slag over the years, and as a result, the amount of Cr(VI)contaminated soil was more than tens of thousands of tons, and the depth of the polluted soil was more than 10 m. In 2019, the microbial leaching method coupled with chemical stabilization was applied in the first phase of the treatment for chrome-contaminated soil.

6.2 Remediation Technology and Route The Cr(VI)-contaminated soil was broken and built into a pile, followed by bioleaching with Cr(VI)-reducing bacteria, and most of Cr(VI) in the soil was dissolved. The solution was pumped to the biochemical tank, Cr(VI) in the leaching solution was reduced to Cr(III) by Cr(VI)-reducing bacteria, and a chromium hydroxide precipitate was formed under alkaline conditions. The upper layer of bacterial culture solution was pumped back to the culture tank for recycling. A stabilizing agent was added to the soil after leaching to convert a small amount of residual Cr(VI) to Cr(III). The reducing bacteria retained in the soil continued to reduce trace amounts of Cr(VI) to ensure remediation effects. Figure 6.1 shows the technical process.

© The Author(s), under exclusive license to Springer Nature Singapore Pte Ltd. 2023 W. Yang et al., Remediation of Chromium-Contaminated Soil: Theory and Practice, Environmental Science and Engineering, https://doi.org/10.1007/978-981-99-5463-6_6

299

300

6 Case of Chrome-Contaminated Site Remediation Project

Contaminated soils Cr(VI)-reducing bacteria

broken

spray

pilot scale culture

leaehing tank

leachate

recycle by pump

Detoxification in Biochemical tank

supernatant

sediment

Leached soil

chemical stabilization

backfill

Safe Disposal Fig. 6.1 Technological process of bacterial bioleaching coupled with chemical stabilization for treating Cr(VI)-contaminated soils

6.3 Project Photos The project was initiated in October 2019 and terminated in September 2020. The remedied soil quantity amounted to 17,214.7 m3 (Fig. 6.2).

6.4 Remediation Effects Moderate, heavy and extremely heavy chromium-contaminated soils were selected for small-scale and pilot tests. The results showed that the leachate concentrations of Cr(VI) in TCLP after bioleaching were higher than the moderately contaminated soil were less than 0.5 mg L−1 . After bioleaching, the addition of a chemical stabilizer resulted in leachate concentrations of Cr(VI) in TCLP of less than 0.05 mg L L−1 . The optimal chemical stabilization agents were 2.2%, 3.3% and 4.7% for the moderate, heavy and extremely heavy chromium-contaminated soils, respectively. In the pilot tests with optimum conditions, the results showed that the leachate concentrations of Cr(VI) in TCLP were less than 0.05 mg L−1 in the moderately, heavily and extremely heavily chromium-contaminated soils after bioleaching and chemical stabilization,

6.4 Remediation Effects

Soil remediation tank

301

Microbial culture tank

Spraying with bacterial culutre solution

Spraying system

chemical stabilization

Fig. 6.2 Chromium-contaminated soil remediation site

which reached the class III standard of the “Surface Water Environmental Quality Standard” (GB3838) (Table 6.1). At the end of project completion, the remedied soils were monitored by the thirdparty testing agency. The results showed that the leachate concentrations of Cr(VI) in TCLP met the class III standard of the “Surface Water Environmental Quality Standard” (GB3838) in the moderately, heavily and extremely heavily chromiumcontaminated soils after bioleaching and chemical stabilization (Table 6.2).

302

6 Case of Chrome-Contaminated Site Remediation Project

Table 6.1 Remediation effects of chromium-contaminated soils by bioleaching and chemical stabilization in small-scale tests Samples

Bacterial Leaching leaching concentration of Cr(VI) in soil after leaching of bacterial solution (mg/ L)

Amount of Time reducing Leaching concentration of stabilization Cr(VI) (mg/L) agents (%) 1d 3d 5d 10d 30d

Moderate chromium contaminated soil samples(leaching concentration of Cr(VI): 4.45 mg/L)

Bacterial 0.15 solution

2.1

0.11 0.09 0.07 0.07 0.05

Bacterial solution

2.2

0.09 0.04 0.02 NDa ND

Bacterial solution

3.2

0.08 0.06 0.05 0.02 0.04

Bacterial solution

3.3

0.04 0.01 NDa NDa NDa

Bacterial 0.32 solution

3.2

0.15 0.11 0.09 0.08 0.06

Bacterial solution

3.3

0.12 0.04 0.04 0.02 0.02

Bacterial solution

4.2

0.13 0.07 0.05 0.05 ND

Bacterial solution

4.3

0.09 0.03 ND

Bacterial 0.45 solution

4.5

0.19 0.12 0.11 0.08 0.08

Bacterial solution

4.7

0.12 0.07 0.03 ND

ND

Bacterial solution

5.5

0.08 0.03 0.01 ND

ND

Bacterial solution

5.7

0.04 0.01 ND

ND

Heavy chromium contaminated soils (leaching concentration of Cr(VI): 27.16 mg/L)

Extremely heavy chromium contaminated soils(leaching concentration of Cr(VI): 69.58 mg/L)

a No

detection

ND

ND

ND

6.4 Remediation Effects

303

Table 6.2 Remediation effects of chromium-contaminated soils by bioleaching and chemical stabilization in pilot tests Samples

Bacterial leaching

Addition level of stabilization agents (%)

Treat time 1d

3d

5d

10d

30d

Moderate chromium contaminated soil samples (4.04 mg/L)

Bacterial solution

2.2

0.05

0.03

NDa

ND

ND

Heavy chromium contaminated soils (29.1 mg/L)

Bacterial solution

3.3

0.07

0.03

ND

ND

ND

Extremely heavy chromium contaminated soils (leaching concentration (62.4 mg/L)

Bacterial solution

4.7

0.09

0.02

0.02

0.02

ND